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al.. 2008). Many streams had ambient ANC < 20 ueq/L. Hindcast simulations suggested
that preindustrial ANC was above 50 ueq/L in all of the study streams.
+ -r

X

Z
/—v


+
i
1 CO
>>
o

z
1
E
+
"o
1
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Target year: 2050
Target ANC: 20 ^molc/L
Target year: 2150
Target ANC: 20 pmoUL
Target year: 2050
Target ANC: Preirdustrial-20 umolc/L
Target year: 2150
Target ANC: Preindustriat-20 pmoyL
o
o
o
0-50
50-200
>200
Source: Fakhraei et al. (2016).
Figure 8-15. Exceedance level of current NOs" + SO42" atmospheric deposition
for 387 stream sites in the GRSM. Exceedances were calculated
for the years 2050 and 2150 using two targets for modeled stream
ANC recovery of 20 |jmolc/L and 20 pmolc/L less than the
simulated preindustrial ANC.
8.5.4.2. International Critical Loads
1	Critical load concepts were initially developed in Europe and only more recently widely
2	applied in the U.S. and Canada. Work on CLs has continued internationally as a basis for
3	setting environmental policy. Some of the recent work conducted outside the U.S. is
4	summarized below. Previous regional CL assessments in Canada have often employed an
5	empirical clay-based soil texture approximation to estimate weathering as input for
6	aquatic and terrestrial steady-state CL modeling Krzvzanowski and Inncs (2010)
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estimated steady-state CLs, using the SSWC model, and associated exceedances for
protecting freshwater ecosystems in First Nations territory in British Columbia against
water acidification. Estimates of acidic deposition were not in exceedance of the CL at
any of the study lakes.
The 1999 Gothenburg Protocol, under the Long-Range Transboundary Air Pollution
(LRTAP) Convention of the United Nations Economic Commission for Europe
(UNECE) was aimed at reducing the total area of Europe where acidic deposition
exceeded CLs. In 2007, the Coordination Centre for Effects (CCE) of the International
Cooperative Programme for Modeling and Mapping (part of the LRTAP Convention)
issued a call to European countries for results from dynamic models of soil and water
acidification. In response to this call, Norwegian scientists used the MAGIC model to
project recovery of surface waters in Norway. Larssen et al. (2010) described the results
for water quality and fish population status and considered implications for policy
options. The modeling results suggested that surface waters will continue to recover
slowly under existing emissions controls, but about 18% of Norway will still have lakes
that receive acidic deposition that exceeds the CL to protect against aquatic acidification
and in which water quality will continue to be insufficient to support viable populations
of fish and other aquatic organisms.
Both measured and modeled data indicate past acidification and suggest some future
chemical recovery at acid-sensitive locations throughout northern Europe. Some surface
waters have experienced much greater levels of acidification than have acid-sensitive
sites in the eastern U.S., For example, the MAGIC model was used by Skeffington et al.
(2016) to simulate past and future trajectories in stream chemistry of a seriously acidified
small forest watershed in England. Hindcast simulations suggested that ANC decreased
from about 150 to -100 j^ieq/L and pH decreased from 7.1 to 4.2. Hypothesized decreases
in future acidic deposition suggested slow and prolonged chemical recovery over a period
of 250 years to ANC = 43 (ieq/L.
Bishop et al. (2008) applied the SSWC model to lakes in northern Sweden, many of
which are rich in organic matter. They argued that the SSWC model is not suitable for
judging the acidification of individual lakes in regions such as northern Sweden where
the extent of chronic acidification is relatively small and where organic influence on lake
chemistry is pronounced. A variant of the SSWC model predicted preindustrial lake
chemistry at 58 sites where paleolimnological reconstructions of lake chemistry were
available. The authors noted a large discrepancy between SSWC predictions and diatom
reconstructions, and this was attributed largely to short-term fluctuations in modern lake
chemistry.
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8.6. Aquatic Acidification Summary and Causal Determinations
In the 2008 ISA, the body of evidence was sufficient to infer a causal relationship
between acidifying deposition and changes in freshwater biota. Overall, the updated
research synthesized in this ISA reflects incremental improvements in scientific
knowledge of aquatic biological effects and indicators of acidification as compared with
knowledge summarized in the 2008 ISA. Studies indicate that aquatic organisms have
been affected by acidification at virtually all trophic levels in sensitive ecosystems and
that these responses have been well characterized for several decades. Research and
observations reported in the 2008 ISA showed consistent and coherent evidence for
effects on aquatic biota, especially algae, benthic invertebrates, and fish that are most
clearly linked to chemical indicators of acidification (pH, ANC, inorganic A1
concentration). Effects on fish species are especially well understood, and many species
have been documented to be adversely affected. Both in situ and lifestage experiments in
fish support previous thresholds of chemical indicators and biological effects.
Physiological perturbations at the organism level of biological organization can lead to
effects on reproduction, growth, and survival. More species are lost with greater
acidification, providing evidence of a biological gradient in effects. Biological shifts such
as loss of acid-sensitive organisms, population declines, and decreased species richness
have been reported from acid-sensitive regions of the U.S. and other countries. Despite
reductions in acidifying deposition, many aquatic ecosystems across the U.S. are still
experiencing effects on ecological structure and functioning at multiple trophic levels.
New information is consistent with the conclusions of the 2008 ISA that the body of
evidence is sufficient to infer a causal relationship between acidifying deposition and
changes in biota, including physiological impairment and alteration of species
richness, community composition, and biodiversity in freshwater ecosystems. Baker
et al. (^Oa) conducted a rigorous review of the effects of acidification on aquatic biota
for the 1990 National Acid Precipitation Assessment Program (NAPAP) State of
Science/Technology reports. In the review, hundreds of laboratory studies, in situ
bioassays, field surveys, whole-system field experiments, and mesocosm studies were
evaluated to provide a synthesis of the effects of acidification on aquatic biota. The
findings in that report, along with literature published after 1990 up to 2007, were
included in the 2008 ISA. The summaries below integrate information known at the time
of the 2008 ISA with newer studies. For most topics related to aquatic acidification, the
fundamental understanding of mechanisms and biological effects has not changed; rather,
additional studies lend further support to the findings from the previous ISA.
The effects of acidification on freshwaters in the U.S. and elsewhere were well
characterized at the time of the 2008 ISA. The sensitivity of a watershed depends on
watershed characteristics such as underlying geology and hydrological flowpaths
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(Appendix 7). and the sensitivity of species that make up the local biological community.
Changes in biota are linked to chemical indicators in surface water (Appendix 7;
Table 8-9). As stated in the 2008 ISA, biological effects are primarily attributable to low
pH and high inorganic A1 concentration. ANC is also used as a chemical indicator of
acidification because it integrates overall acid-base status and because surface water
acidification models do a better job projecting ANC than pH and inorganic A1
concentrations. However, ANC does not relate directly to the health of biota. The
usefulness of ANC lies mainly in the association between ANC and the surface water
constituents or parameters that directly cause or ameliorate acidity-related stress, in
particular pH, Ca, and inorganic Al.
Acid-sensitive freshwater systems can either be chronically acidified or subject to
periodic episodes of decreased pH and ANC and increased inorganic Al concentration.
Chronically acidic lakes and streams were traditionally defined as having ANC <0 (j,eq/L.
The ANC in freshwater systems where episodic acidification occurs may fall below
0 |icq/L for only a few hours to weeks in a given year (Driscoll et al.. 2001b). During
these acid episodes, water chemistry may exceed acid tolerance of resident aquatic biota.
Biological effects of episodes may include fish mortality, changes in species
composition, and declines in aquatic species richness across multiple taxa. Biological
effects of chronic and episodic acidification have been most clearly documented for
phytoplankton, zooplankton, aquatic invertebrates, and fish and can be linked to changes
in the chemical indicators of aquatic acidification (Table 8-9). An ANC of >50 ueq/L was
used as an indicator for acidification to an ANC level that may harm biota in the U.S.
EPA National Lake Surveys (U.S. EPA. 2009b).
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Table 8-9 Ecological indicators for aquatic acidification.
Key Indicators
Ecological Effect
Key References
Acid neutralizing
capacity
Commonly set at ANC values less than 0, 20, 50,
and 100 peq/Lto correspond with decreasing levels
of concern (Fiqure 8-4)
Driscoll et al. (2001b), MacAvov and
Bulqer (1995), Baker et al. (1990a),
U.S. EPA (2009b)
Base cation surplus
0 peq/L—risk of inorganic Al leaching to streams
Section 3.2.1.5 2008 Integrated
Science Assessment for Oxides of
Nitrogen and Sulfur-Ecological Criteria
U.S. EPA (2008a) Lawrence et al.
(2007)
PH
<6.0—reduced number of fish species
Driscoll et al. (2001b). MacAvov and
Bulqer (1995), Baker et al. (1990a)
Inorganic Al
>2 pmol/L (54 |jg/L)—toxic to brook trout and likely
other aquatic biota
Baldiqo et al. (2007), Driscoll et al.
(2001b). Wiqinqton et al. (1996a).
MacAvov and Bulqer (1995)
Al = aluminum; L = liter; |jeq = microequivalent; |jg = microgram.
Source: modified from Fenn et al. (2011b).
8.6.1. Phytoplankton
Phytoplankton, photosynthesizing forms of plankton, play important roles in freshwater
systems as primary producers at the base of the aquatic ecosystem food web. These
organisms, encompassing diatoms, cyanobacteria, dinoflagellates, and other algae, vary
in tolerance of acidic conditions. Studies reviewed in the 2008 ISA reported reduced
species richness of phytoplankton with the decreases in pH and increases in inorganic Al
concentrations that are associated with acid-affected surface waters (U.S. EPA. 2008a).
Effects were most prevalent in the 5 to 6 pH range (Baker etal.. 1990a). Since the 2008
ISA, several paleolimnological and field studies have further linked phytoplankton
community shifts to chemical indicators of acidification. For example, Lacoul et al.
(2011) reviewed information on the effects of acidification on plankton in Atlantic
Canada and observed that the greatest changes in phytoplankton species richness
occurred over a pH range of 4.7 to 5.6, just beyond the interval (pH 5.5 to 6.5) where
bicarbonate becomes depleted in the water.
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8.6.2.
Zooplankton
Zooplankton, the animal forms of plankton, comprise many groups of small freshwater
organisms including protozoans, rotifers, cladocerans, and copepods. Decreases in ANC
and pH and increases in inorganic A1 concentration were shown to contribute to the loss
of zooplankton species and/or abundance in lakes in studies reviewed in the 2008 ISA
(Keller and Gunn. 1995; Schindler et al.. 1985). Possible mechanisms for zooplankton
sensitivity to low pH and ANC include ion regulation failure, reduced oxygen uptake,
inability to reproduce, and inorganic Al toxicity (U.S. EPA. 2008a'). Reported pH
thresholds for zooplankton community alteration generally ranged from 5 to 6 in studies
reviewed in the 2008 ISA. For example, a decrease in pH from 6 to 5 caused decreased
species richness in zooplankton communities in lakes (Holt et al.. 2003; Holt and Yan.
2003; Locke and Sprules. 1994). Newer studies support effects in a similar pH range.
Vinebrooke etal. (2009) reported variations in phytoplankton and zooplankton
communities during a whole-lake experimental acidification of Lake 302S in the
Experimental Lakes area in Ontario. There was a negative effect on zooplankton species
richness as pH decreased from 6.8 to 4.5. Acidification often reduces Ca availability in
lake water and can affect growth and survival of Daphnia spp., an important prey item in
many freshwater food webs (Jeziorski et al.. 2012b). At ANC <0 (ieq/L, zooplankton
richness was low in Adirondack lakes [15 species in highly acidic lakes compared to
35 species at the highest values of ANC in the study (near 200 (j,eq/L) (Sullivan etal..
2006a)].
8.6.3. Benthic Invertebrates
Sediment-associated invertebrates such as bivalves, worms, gastropods, and insect larvae
can be impacted by acidification because H+ and Al can be directly toxic, causing
disruption of ion regulation and reduced reproductive success. As reviewed in the 2008
ISA, decreases in ANC and pH and increases in inorganic Al concentration have been
shown to contribute to the decline in abundance or loss of benthic invertebrates species in
streams. Typically, pH below 5 virtually eliminate all mayflies, a common taxa used to
assess water quality, along with other aquatic organisms from some streams (U.S. EPA.
2008a; Baker and Christensen. 1991). Since the 2008 ISA, a survey of benthic
macroinvertebrates by Baldigo et al. (2009) in 36 streams in the southwestern
Adirondack Mountains indicated that 44 to 56% of macroinvertebrate communities were
severely impacted by acidification at pH <5.1, moderately impacted at pH 5.1 to 5.7, and
unaffected at pH above 6.4 (Baldigo et al.. 2009). Thresholds of pH 5.2 to 6.1 were
identified for sensitive invertebrates from Atlantic Canada; below these pH values,
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changes in the abundance or presence of invertebrate taxa were observed (Lacoul et al..
2011).
8.6.4. Fish
The primary mechanism that controls the toxic effects of acidification on fish involves
disruption of normal ion regulation at the gill surface, resulting in increased rates of ion
loss, inhibition of ion uptake, and loss of gill function (Bergman et al.. 1988; Wood and
McDonald. 1987; Leivestad. 1982; McWilliams and Potts. 1978). Responses to
acidifying conditions include respiratory and circulatory failure. The effects of low pH,
low ANC, and high inorganic Al concentrations have been well characterized in fish for
many decades. Responses among fish species and lifestages within species to pH and Al
in surface waters have been variable. In general, early lifestages are more sensitive to
acidic conditions than the young-of-the-year, yearlings, and adults (Baker etal.. 1990a;
Johnson et al.. 1987; Baker and Schofield. 1985). Some of the most commonly studied
species include brown trout, brook trout, and Atlantic salmon.
Further characterization of physiological responses (ion regulation, stress responses, gill
Al accumulation) to acidification in fish, mostly Atlantic salmon, trout, and other
salmonids, adds to the existing information on sublethal effects on individual fish species.
Many of the newer studies were conducted in situ and reported varying sensitivity of
different lifestages. Findings are consistent with physiological alterations in fish reported
in the 2008 ISA. Several studies have assessed physiological changes associated with
migratory activities. For example, a recent study assessed gill Al and NKA activity in
smolts moving downstream in well-buffered and acid-impacted migration corridors in the
northeastern U.S. (Kelly et al.. 2015). In the acid-impacted river basin, the fish had
elevated gill Al and lower gill NKA activity.
As summarized in Baker etal. (1990a) and studies reviewed in the 2008 ISA, fish
populations in acidified streams and lakes of both Europe and North America have
declined, and some have been eliminated as a result of atmospheric deposition of acids
and resulting changes in pH, ANC, and inorganic Al concentrations in surface waters.
There is often a positive relationship between pH and number of fish species, at least for
pH values between about 5.0 and 6.5 (Cosby et al.. 2006; Sullivan et al.. 2006a; Driscoll
et al.. 2003b; Bulger et al.. 1999). Additional pH thresholds published since the 2008 ISA
generally reinforce these ranges, and several new studies consider the role of DOC in
modulating pH and subsequent effects on biota. New studies on fish responses to
chemical alarm cues show behavioral effects at pH <6.6. Characterization of ANC and its
levels of concern have not changed appreciably with the newly available information.
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Few or no fish species are found in lakes and streams that have very low ANC (near zero)
and low pH [near 5.0; (U.S. EPA. 2008a; Sullivan et al.. 2006aYI. The number of fish
species generally increases at higher ANC and pH. The pH largely controls the
bioavailability of Al (Driscoll et al.. 2001^. Al is very toxic to fish, and thresholds of
response to elevated concentrations of this metal in acidified waters are summarized in
Table 8-4.
Some of the most in-depth studies of the effects of acid stress on fish have been
conducted in streams in Shenandoah National Park, VA (Cosby et al.. 2006). and lakes in
the Adirondack Mountains, NY (Sullivan. 2015). Effects on fish have also been
documented in acid-sensitive streams of the Catskill Mountains of southeastern New
York and the Appalachian Mountains from Pennsylvania to Tennessee and South
Carolina (Bulger et al.. 2000; Bulger etal.. 1999; SAMAB. 1996; Charles and Christie.
1991V
8.6.5. Thresholds of Response
As reviewed above, new thresholds have been identified in aquatic organisms
(Table 8-10). However, this new information does not appreciably change the
understanding of biological effects associated with chemical indicators or the levels at
which effects occur. Evidence continues to strengthen findings in the 2008 ISA that high
levels of acidification (especially to pH values below 5) eliminate sensitive species from
freshwater streams.
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Table 8-10 Results of recent biological effects studies in surface waters
indicative of thresholds of biological response to changes in water
acidity.
Life Forms and Effects
Region
Potential Tipping
Points
Reference
Sensitive invertebrates present
Atlantic Canada
pH 5.2 to 6.1
Lacoul et al. (2011)
Littoral macroinvertebrates
Northern Europe
pH 5.8 to 6.5
Schartau et al. (2008)
Brook trout loss of whole-body Na
Great Smoky Mountains NP
pH 5.1
Neffet al. (2008)
Brook trout loss of whole-body Na
of 10 to 20%
Great Smoky Mountains NP
pH 4.9 to 5.1
Neffet al. (2009)
Fish damage in lakes
Norway
ANC 67 peq/L
Hesthaaen et al. (2008)
Juvenile brown trout mortality in
high DOC streams
Sweden
pH 4.8 to 5.4
Serrano et al. (2008)
Embryo and yolk sac fry survival
during episodes in DOC-rich lakes
Sweden
pH 4.0
Serrano et al. (2008)
Toxicity to brown trout in humic
streams
Northern Europe
pH 5.0; inorganic
aluminum 20 |jg/L
Andren and Rvdin (2012)
Presence of macrophytes
Atlantic Canada
pH 5.0
Lacoul et al. (2011)
Aquatic bird breeding
Atlantic Canada
pH 5.5
Lacoul et al. (2011)
ANC = acid neutralizing capacity; DOC
NP = National Park.
= dissolved organic carbon; L = liter; |jeq = microequivalent; |jg = microgram; Na = sodium;
8.6.6. Biological Recovery
1	Biological recovery (Chapter 1.11.1) can occur only if chemical recovery (Appendix 7) is
2	sufficient to allow growth, survival, and reproduction of acid-sensitive plants and animals
3	(Driscoll et al.. 2001b). Modeling studies in the northeastern and southeastern U.S.
4	suggest that full chemical recovery may take many decades or not occur at all due to the
5	dynamics of S adsorption and desorption and to long-term Ca depletion of soils. As
6	reported in the 2008 ISA, biological recovery lags behind chemical recovery in many
7	aquatic systems, and the time required for biological recovery after chemical recovery is
8	complete is uncertain (U.S. EPA. 2008a). Ecosystems deemed to be on a recovery
9	trajectory are those found to be moving towards a mix of species presence and abundance
10	that approximates the undisturbed state.
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Since the publication of the 2008 ISA, additional studies have assessed recovery of
benthic organisms, although most of this research has been conducted in Canadian and
European waters. New studies continue to support these earlier observations. In general,
recovery of plankton and benthic invertebrates is observed prior to recovery of fish
populations, although most biological communities studied to date have not returned to
preacidification conditions, even after recovery of chemical parameters. In a study
reviewed in the 2008 ISA, zooplankton recovery in experimentally acidified Little Rock
Lake in Wisconsin took one decade, with approximately 40% of the zooplankton species
experiencing a lag time of 1 to 6 years (Frost et al.. 2006). In an experimentally acidified
lake in Canada, zooplankton species diversity partially rebounded to preacidification
levels as pH returned back to 6.1, with rotifers recovering less than crustaceans (Mallcv
and Chang. 1994). Newer studies from the Sudbury Lakes region of Canada indicate
substantial recovery of copepods and cladocerans at pH 5.5 and higher in previously
acidified lakes (Valois etal.. 2011).
In the 2008 ISA, recovery of fish populations following liming or reduction of deposition
was reported in several studies. Newer studies have documented successful
reintroduction of brook trout in previously acidified Adirondack water bodies (Brooktrout
Lake) or recolonization (Honnedaga Lake) by populations in tributary refuges
(Sutherland et al.. 2015; Josephson et al.. 2014). Fish community shifts from historical
acidification have been observed in the upper mainstem of Hubbard Brook (Warren et al..
2008) and other locations.
8.6.7. Most Sensitive and Most Affected Regions
The extent and distribution of sensitive ecosystems and regions in the U.S. were well
known at the time of the 2008 ISA. In the U.S., surface waters that are most sensitive to
acidification based on ANC are largely found in the Northeast, southern Appalachian
Mountains, FL, the upper Midwest, and the mountainous West (McDonnell et al.. 2014b;
Greaver et al.. 2012; Campbell et al.. 2004a; Driscoll et al.. 2001b; Baker etal.. 1990b;
Omernik and Powers. 1983). Levels of acidifying deposition in the West are low in most
areas, acidic surface waters rare, and the extent of chronic surface water acidification that
has occurred to date has been limited (Charles and Christie. 1991). Episodic acidification
does occur in both the East and West at some acid-sensitive locations, and this is part
natural and part human-caused. Geographic patterns in acidification sensitivity vary in
response to spatial differences in geology, hydrologic flow paths, presence and depth of
glacial till, climate, and others. Sullivan (2017) mapped the locations of known lakes and
streams that had low ANC across the country (Figure 8-11).
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8.6.8.
Critical Loads
Since the 2008 ISA, considerable CL research has been conducted in the U.S. New
generalized empirical CL estimates include 571 eq N/ha/yr in the Northeast and
286 eq N/ha/yr in the West for episodic acidification of high elevation lakes under
high-flow conditions (Baron et al.. 201 lb). Heard et al. (2014) estimated CL = 74
eq/ha/yr to protect against chronic acidification in high-elevation lakes in the Sierra
Nevada. Steady-state CLs have been derived at many locations since the 2008 ISA.
Steady-state CLs of S and N for lakes in the Adirondack Mountains (1,620 eq/ha/yr) and
for the central Appalachian streams (3,700 eq/ha/yr) were calculated for maintaining a
surface water ANC of 50 |icq/L on an annual basis (NAPAP. 2011). Sullivan et al.
(2012b) calculated CL values in the Blue Ridge ecoregion for maintaining stream ANC at
50 (ieq/L. The calculated CLs were less than 500 eq/ha/yr at one-third of the study sites.
Observations showed that about one-half or more of the stream length in the study region
was in exceedance of the CL of S deposition for protecting aquatic resources to an
ANC = 50 (ieq/L over the long term. McDonnell et al. (2014b) calculated steady-state
aquatic CLs to protect southern Appalachian Mountain streams against acidification to
ANC = 50 (ieq/L and other critical benchmark values. Results showed that nearly
one-third of the stream length in the study region (mainly streams in Virginia, West
Virginia, North Carolina, and Tennessee) had a CL of S deposition <500 eq/ha/yr, which
was less than the estimated regional average S deposition (600 eq/ha/yr). Critical loads
for acid deposition to lakes in Class I and II wilderness areas of the Sierra Nevada were
estimated in 2008 to protect ANC to 0, 5, 10, and 20 (j,eq/L levels, which span the range
of minimum ANC values observed in those lakes. Median CLs were 217 (ANC = 0), 186
(ANC = 5), 157 (ANC = 10), and 101 (ANC = 20) eq/ha/yr. The median CL for granitic
watersheds based on a critical ANC limit of 10 j^icq/L was 149 eq/ha/yr. It was estimated
that slightly more than one-third of the lakes received acidic deposition higher than their
CL.
In addition to the steady-state and empirical CLs described above, CL estimates are
available from dynamic modeling. NOs leaching in stream water in California was both
simulated (by the DayCent model) and determined empirically to be approximately
1,214 eq N/ha/yr (Fenn et al.. 2008). Zhou et al. (2015a) simulated past and future effects
of N and S on stream chemistry of 12 watersheds in the Great Smoky Mountain National
Park. Three target levels of ANC (0, 20, and 50 j^icq/L) were used, based on a range of
protection of aquatic life from minimal to considerable. TLs of NO3 + SO42 deposition
for the 12 study streams ranged from 270 to 3,370 eq/ha/yr to reach ANC = 0 |icq/L by
2050, 0 to 2,340 eq/ha/yr to reach ANC = 20 (j,eq/L by 2050, and 0 to 1,400 eq/ha/yr to
reach ANC = 50 |icq/L by 2050. However, the majority of streams could not achieve the
ANC target of 50 j^icq/L. This was also true to a lesser extent for the target of
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ANC = 20 (ieq/L. Modeling studies also suggested that complete recovery may not be
possible in the Appalachian Mountains (Sullivan et al.. 201 lb). For some sites, one or
more of the selected critical ANC levels (0, 20, 50, 100 j^icq/L) could not be achieved by
2100, even if S deposition was decreased to zero and maintained at that level throughout
the simulation. MAGIC modeling based on simulations of past and future acid-base
chemistry of 14 streams in Shenandoah National Park identified a target load of about
188 eq kg S/ha/yr to achieve ANC = 50 (ieq/L in 2100 in the median modeled stream
located on sensitive (siliciclastic) bedrock, which was 77% lower than the S deposition in
1990 (Sullivan et al.. 2008). Many streams had ambient ANC <20 (ieq/L, although
hindcast simulations suggested that preindustrial ANC was above 50 j^ieq/L in all of the
study streams.
In the Adirondack Mountains, TLs were calculated for two time periods (2050 and 2100)
and three levels of protection (ANC = 0, 20, and 50 j^ieq/L). Results of simulated TLs,
and associated exceedances, were extrapolated to the regional population of lakes. About
30% of the lakes had TL <500 eq/ha/yr to protect lake ANC to 50 j^icq/L (Sullivan et al..
2012a). Also in the Adirondack Mountains, Zhou et al. (2015c) ran simulations using the
PnET-BGC model, which suggested that future decreases in SOr deposition would be
more effective in that region in increasing the lake water ANC than equivalent decreases
in N03 deposition. In another modeling study of 20 Adirondack watersheds, lake ANC
and fish and total zooplankton species richness were projected to increase under
hypothetical decreases in future acidic deposition, but model projections suggested that
lake ecosystems may not achieve complete chemical and biological recovery in the future
(Zhou et al.. 2015b). Estimates of preindustrial ANC for the study lakes ranged from 18
to 190 (ieq/L. The magnitude of simulated historical acidification represented by ANC
loss ranged from about 26 to 100 (ieq/L.
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APPENDIX 9. BIOLOGICAL EFFECTS OF
FRESHWATER NITROGEN
ENRICHMENT
This appendix characterizes the biological effects of nitrogen (N) nutrient enrichment
from atmospheric deposition to freshwater systems. Biogeochemical processes and
chemical indicators associated with nutrient enrichment of fresh waters are discussed in
Appendix 7. Atmospheric deposition constitutes only a portion of total N load in many
water bodies; however, it may be the dominant source of N in some remote aquatic
ecosystems, such as headwater and lower order streams and alpine lakes, which are more
affected by N deposition than other sources of N. Appendix 9J_ presents an overview of
freshwater nutrient enrichment in these systems and includes a discussion on the
characteristics of water bodies sensitive to N deposition, the effects of N deposition on
nutrient limitation, phosphorus (P) interactions, and climate modification of N response.
Inputs of P have direct implications for how ecosystems respond to N deposition, and
recent trends in increased atmospheric deposition of P are discussed in this context.
Indicators of biological responses to nutrient enrichment (Appendix 9.2) effects of N on
species diversity, ecosystem structure, and function (Appendix 9.3) and emerging
research on the links between nutrient enrichment and animal behavior and disease
(Appendix 9.4) provide a basis for identifying thresholds of biological response
(Appendix 9.5) and determining causation based on new information and evidence from
prior N assessments (Appendix 9.6).
9.1. Introduction to Nitrogen Enrichment and Eutrophication in
Freshwater Systems
In the 2008 Integrated Science Assessment for Oxides of Nitrogen and
Sulfur—Ecological Criteria (2008 ISA), the body of evidence was sufficient to infer a
causal relationship between N deposition and the alteration of species richness, species
composition, and biodiversity in freshwater ecosystems (U.S. EPA. 2008a). Increased
atmospheric N inputs to freshwater systems via runoff or direct deposition, especially to
N limited and N and phosphorus (P) colimited systems, can stimulate primary
productivity (Figure 9-1). Eutrophication is the process of enriching a water body with
nutrients resulting in increased growth and change in the composition of primary
producers (algae and/or aquatic plants). One of the consequences of eutrophication is low
oxygen levels in the water body when these primary producers decompose. Changes in
biological indicators of N enrichment, including chlorophyll a, phytoplankton
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(free-floating algae) biomass, periphvton (algae attached to a substrate) biomass, and
diatoms (major algal group with cell walls made of silica), provide evidence for N
effects. The transport of atmospheric N inputs downstream can exacerbate eutrophic
conditions in higher order streams, rivers, and lakes. Atmospheric deposition to
watersheds affects processes along the freshwater-to-ocean continuum, including coastal
and estuarine systems (Appendix 10). New information is consistent with the conclusions
of the 2008 ISA that the body of evidence is sufficient to infer a causal relationship
between N deposition and changes in biota, including altered growth and
productivity, species richness, community composition, and biodiversity due to N
enrichment in freshwater ecosystems.
Source: Modified from Baron et al. (2011 b).
Figure 9-1 Conceptual model of the influence of atmospheric nitrogen
deposition on freshwater nutrient enrichment.
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Based on the studies described throughout Appendix 9 and in the 2008 ISA, the
freshwater ecosystems in the U.S. most likely to be sensitive to N deposition are
headwater streams, lower order streams, and alpine lakes, which have very low nutrients
and productivity and are far from local pollution sources (U.S. EPA. 2008a'). Even small
inputs of N in these water bodies can increase nutrient availability or alter the balance of
N and P, which can stimulate growth of primary producers and lead to changes in species
richness, community composition, and diversity. Remote mountain lakes in the western
U.S. are naturally oligotrophic and are considered among the aquatic ecosystems most
sensitive to N deposition (Williams et al.. 2017b). A portion of these lakes and streams in
the western U.S. are in Class I wilderness areas (Williams et al.. 2017b; Clow et al..
2015; Nanus et al.. 2012). Survey data and fertilization experiments from studies
reviewed in the 2008 ISA have documented increases in algal productivity as well as
species changes and reductions in diversity at high-elevation lakes in the western U.S. in
response to increased availability of N (U.S. EPA. 2008a). Some examples include the
Snowy Range in Wyoming, the Sierra Nevada, Lake Tahoe, and the Colorado Front
Range. In studies reviewed in the 2008 ISA, atmospheric N deposition (approximately 1
to 5 kg N/ha/yr) can cause an increase in phytoplankton and periphyton biomass in some
alpine lakes. For example, in both the Beartooth Mountains of Wyoming and the Rocky
Mountains of Colorado, N deposition as low as 1.5 kg N/ha/yr affects algal productivity
(Baron. 2006; Saros et al.. 2003). The responses of high-elevation lakes can vary
considerably depending on catchment characteristics and the amount of deposition
(Appendix 9.1.1).
With increased characterization of nutrient inputs in remote high-elevation lakes and
streams, fate and transport processes, N and P dynamics, phytoplankton response, and
downstream effects in coastal/estuarine systems, the understanding of the role of N in
freshwater eutrophication has evolved in recent decades. As conveyed in the 2008 ISA,
the historical emphasis on P as a major cause of freshwater eutrophication was based on a
number of highly influential studies in the 1960s and 1970s that focused on the role of
wastewater, in particular phosphate detergents, in causing excessive algal blooms in Lake
Mendota (WI), Lake Washington (WA), Lake Erie, and many other locations
(Edmondson. 1991; Schindler. 1974; Schindler et al.. 1971; Edmondson. 1969;
Vollenweider. 1968; Hasler. 1947). These observations guided approaches to freshwater
lake and stream ecology for many years. Over time, an increasing recognition that N
inputs can stimulate phytoplankton growth under certain conditions, coupled with further
understandingof aquatic biogeochemistry (Appendix 7) and of the connectivity between
freshwater and receiving estuaries and coastal waters, has led to recommendations to
consider both N and P in nutrient reduction strategies (Dodds and Smith. 2016; Gobler et
al.. 2016; Paerl et al.. 2016b; Lewis etal.. 2011; Scott and McCarthy. 2010; Conlev et al..
2009; Paerl. 2009; Lewis and Wurtsbaugh. 2008).
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There is increasing understanding of the frequency and shifts in the types of nutrient
limitation in various systems. In oligotrophic clear water lakes where atmospheric N
inputs are low and light is not limiting, phytoplankton primary production is typically N
limited (Bcrgstrom et al.. 2015; Hcsscn. 2013). In the 2008 ISA, results from surveys,
paleolimnological reconstructions, experiments, and meta-analyses of hundreds of studies
have shown N limitation to be common in fresh waters, especially in remote areas, and a
nearly universal eutrophication response to N enrichment in lakes and streams that are N
limited (U.S. EPA. 2008a; Elser et al.. 2007; Bergstrom et al.. 2005; Elseretal.. 1990).
Tropical and subtropical lakes, and lakes having small watersheds relative to the lake
surface/volume, also tend to be N limited (Paerl and Scott. 2010; Elser et al.. 2007). As
reported in the 2008 ISA, some alpine lakes have exhibited shifts from N limitation to
between N and P limitation or to P limitation (U.S. EPA. 2008a'). In a meta-analysis
reviewed in the 2008 ISA, Elser et al. (2007) found that N limitation occurred as
frequently as P limitation in freshwater ecosystems. Newer studies published since the
2008 ISA add to the evidence that reservoirs, rivers, and freshwater lakes can exhibit N
limitation and N and P colimitation (Paerl et al.. 2014; Lewis etal.. 2011; Conlev et al..
2009; Sterner. 2008). N limitation appears to be increasingly common in freshwater
systems, probably because their nutrient dynamics are being altered significantly by
growing agricultural and urban P inputs (Paerl et al.. 2016b; Grantz et al.. 2014; Paerl et
al.. 2014; Finlav et al.. 2013).
A source of P to remote water bodies is deposition of dust. In an analysis of data from the
U.S. EPA National Lakes surveys and National Rivers and Streams surveys Stoddard et
al. (2016) found continent-wide increases in TP between 2000 and 2014, especially at
sites that exhibited low disturbance (Appendix 9.1.1.2). Among the large natural P
emission sources are soils, vegetation, and biomass combustion ash. Other notable
sources include industry, agriculture, and mining. Although P is not a criteria pollutant,
inputs of P may contribute to eutrophication and effect shifts in lake trophic status from P
to N limitation or to colimitation. In addition, Stoddard et al. (2016) observed that TN
was strongly correlated with TP in lakes and streams on a national scale. Although the
authors determined that TP was increasing at "minimally disturbed sites," they observed
that TN was not increasing at those sites. Responses of aquatic ecosystems to
atmospheric N deposition are heavily dependent on surface water P concentrations. Thus,
because P inputs can alter N response, the impact of recent trends in increased P
deposition is important to consider when evaluating nutrient status in water bodies
sensitive to atmospheric inputs (Appendix 9.1.1.4).
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9.1.1.
Deposition to Freshwater Systems
Both N and P inputs to fresh waters sensitive to atmospheric deposition can affect
nutrient limitation. The composition of N deposition is shifting in the U.S. from oxidized
to reduced forms of N, with implications for the receiving systems. Atmospheric
deposition of P may affect lake response to N inputs, and recent trends point to
widespread increases in P deposition in the U.S. (Stoddard et al.. 2016) and globally
(Brahncv et al.. 2015; Tipping et al.. 2014) (Appendix 9.1.1.2).
9.1.1.1. Nitrogen Deposition Sources and Trends
Sources of N and trends in atmospheric deposition are described in Appendix 2. Briefly,
N is deposited in various reduced and oxidized forms, including organic N. Deposition
can be wet (rain or snow), or dry. Atmospheric deposition of reduced N has increased
relative to oxidized N in parts of the U.S. including the East and Midwest in the last few
decades, shifting from a NOs dominated to a NH4 dominated condition, and this trend is
expected to continue under existing emissions controls (Li et al.. 2016d; Pinderetal..
2008; U.S. EPA. 2008a). In the soil or water, much of the deposited NFU+ is either taken
up by biota or nitrified to NO3 and leaches to water bodies mainly as NO3 (Hcsscn.
2013). Up to 70% of deposited NO3 in remote alpine lakes in the western U.S. is
anthropogenic in origin, with the largest sources being atmospherically delivered
fertilizers from agriculture (approx. 60%) and fossil fuel combustion (approx. 10%)
(Hundev et al.. 2016). N deposition to snow and glaciers are important sources of N to
alpine lakes and streams that are fed by meltwaters. For example, approximately 50% of
lakes in Glacier National Park and 10-20% of lakes in the central Rockies receive glacial
meltwater (Saros et al.. 2010). At high-elevation sites like those in the Rockies and Sierra
Nevadas, N deposition estimates are uncertain, especially for dry deposition
(Appendix 9.5).
Moving from lower order streams to higher order streams, atmospheric N from direct
deposition, runoff, and leaching from terrestrial ecosystems combines with other diffuse
and point sources of N. The contribution from other terrestrial sources ofN, such as
fertilizer, livestock waste, septic effluent, and wastewater treatment plant outflow, often
becomes much more important in downstream than in upland areas. About 75% of N
inputs are retained in the watershed or denitrified, and 25% are exported to surface waters
regardless of the dominant N input, including deposition (Howarth et al.. 2012; Howarth
ctal.. 1996a). Table 7-1 summarizes studies quantifying N deposition contribution to
total N loading in U.S. freshwater systems.
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9.1.1.2.
Characteristics of Freshwater Systems Sensitive to Atmospheric
Deposition of Nitrogen
Various factors affect the sensitivity of remote water bodies to atmospheric deposition.
These factors include the spatial and temporal patterns of nutrient limitation and the
physical and chemical attributes of the catchment (Williams et al.. 2017b; Hundev et al..
2014; Nanus et al.. 2012; U.S. EPA. 2008a'). Thus, the same amount of N deposition can
lead to different N loading depending on the characteristics of the receiving watershed
and water body (Hcsscn. 2013; Bergstrom. 2010).
In high-elevation lakes above the tree line in areas with steep slopes, sparse vegetation,
exposed bedrock, and shallow rocky soils, changes in productivity and biodiversity of
algal assemblages can occur with little or no lag time (Baron et al.. 201 lb). The
hydrology of these systems is dominated by spring snowmelt (Spaulding et al.. 2015).
Seasonal meltwaters can deliver a nutrient pulse to alpine lakes and streams, with glacial
meltwater contributing more NOs than snowmelt does (Slemmons et al.. 2015;
Slemmons et al.. 2013; Saros et al.. 2010; Baron et al.. 2009). Other freshwater systems,
such as some nonalpine freshwater lakes, reservoirs, and rivers that exhibit N limitation
and N and P colimitation, are sensitive to additional N inputs (Paerl et al.. 2016b; Paerl et
al.. 2014; Lewis et al.. 2011; Conlev et al.. 2009; Elser et al.. 2009b; Sterner. 2008). N
from atmospheric deposition represents a proportion of total N in these systems, although
agricultural and wastewater inputs are often predominant.
In the 2008 ISA, lakes and streams with high concentrations of NO, . indicative of
ecosystems being N saturated (Appendix 7.1.2.1). were reported at a variety of locations
throughout the U.S., including the San Bernardino and San Gabriel mountains within the
Los Angeles Air Basin (Fcnn and Poth. 1999; Fenn et al.. 1996; Riggan et al.. 1985). the
Front Range of Colorado (Williams et al.. 1996a; Baron et al.. 1994). the Allegheny
Mountains of West Virginia (Gilliam et al.. 1996). the Catskill Mountains of New York
(Stoddard. 1994; Murdoch and Stoddard. 1992). the Adirondack Mountains of New York
(Wigington et al.. 1996b). and the Great Smoky Mountains in Tennessee (Cook et al..
1994). All of these regions, except the Colorado Front Range, received more than about
10 kg N/ha/yr atmospheric deposition ofN throughout the 1980s and 1990s. The Front
Range of Colorado received up to about 5 kg N/ha/yr of total (wet + dry) deposition
(Sullivan et al.. 2005). less than half of the total N deposition received at many of these
other locations. Since the 2008 ISA, several long-term monitoring studies have shown
temporal decreases in surface water NO3 concentration corresponding to decreases in
atmospheric N deposition (Appendix 7.1.2.1). These regions include the Appalachian
Mountains, the Adirondacks, and the Rocky Mountains (Driscoll et al.. 2016; Kline et al..
2016; Strock et al.. 2014; Eshleman et al.. 2013; Elser et al.. 2009b; Bergstrom and
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Jansson. 2006). NOs in extremely high concentrations can have direct adverse effects on
fish, as well as invertebrates and amphibians. These effects are observed at
concentrations much higher than would commonly be attributable to atmospheric
deposition, are not included in this ISA, and were not defined as a primary biological
indicator in the 2008 ISA.
9.1.1.3. Nitrogen Deposition Effects on Nutrient Limitation
The 2008 ISA reported on gradient studies of undisturbed northern temperate, mountain,
or boreal lakes that receive low levels of atmospheric N deposition. These studies found
strong relationships between N limitation and primary productivity where N deposition
was low and between P and N + P limitations where N deposition was high (Bcrgstrom
and Jansson. 2006; Bcrgstrom et al.. 2005; Fenn et al.. 2003a'). As reviewed in the
2008 ISA, a comprehensive survey of 42 unproductive lakes (oligotrophic lakes with TP
less than or equal to 25 j^ig/L) along gradients of N deposition in Europe and North
America showed increased inorganic N concentration and productivity to be correlated
with atmospheric N deposition (Bcrgstrom and Jansson. 2006). Unproductive lakes
receiving low atmospheric N deposition were N limited (<2.5 kg N/ha/yr). At about 2.5
to 5 kg N/ha/yr, colimitation of N and P was observed, while at N deposition above
5 kg N/ha/yr, the lakes were P limited. Based on the study findings and paleolimnological
evidence, Bcrgstrom and Jansson (2006) suggested that most lakes in the northern
hemisphere may have originally been N limited and that atmospheric N deposition has
changed the balance of N and P in lakes.
Studies published since the 2008 ISA have continued to characterize nutrient
relationships and evaluate the potential for N deposition to contribute to the
eutrophication of water bodies (Table 9-1). Consistent with the 2008 ISA findings,
research literature after 2007 indicates that N deposition is correlated with a shift from N
to P limitation in certain high-elevation water bodies (Hessen. 2013). Elser et al. (2009b)
conducted a nutrient-limitation study across a gradient of lakes in the Rocky Mountains
of Colorado that receive low (<2 kg N/ha/yr; N = 20) and high (>6 kg N/ha/yr; N = 16) N
deposition. Nutrient enrichment bioassays indicated that P limitation or colimitation was
prevalent in the population of high-deposition lakes (9 of 16 lakes), while only one of the
lakes showed N limitation. In contrast, only 4 of 20 low-deposition lakes showed P
limitation. Based on relative response ratios (RR; chlorophyll concentration in a given
treatment normalized to chlorophyll in control), chlorophyll responded more strongly to
N relative to P in low-deposition lakes where N limitation was stronger (Figure 9-2).
These data were included in Elser et al. (2009a) who reported that lakes in Norway,
Sweden, and Colorado affected by N deposition showed a similar pattern. In lakes with
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high N deposition, phytoplankton was predominately P limited, whereas in lakes with
low N deposition, N limitation was more common.
Based on an analysis of lake water chemistry data from 106 alpine lakes in Sweden,
Slovakia, Poland, and the Rocky Mountains of Colorado and nutrient bioassays from high
mountain lakes in Sweden and the Rocky Mountains of Colorado, Bergstrom (2010)
documented how N:P mass ratios (TN:TP and DIN:TP) vary in oligotrophic lakes located
in northern Europe and the Rocky Mountains. Nutrients (N, P, N + P, control) were
added to water from 28 unproductive lakes. The majority of the lakes were N limited and
the shift from N to P limitation was strongly affected by N deposition. More than half
(54%) of the oligotrophic study lakes had a TN:TP mass ratio <25. A DIN:TP ratio of 1.5
was indicative of an N limited lake while a ratio of 3.4 was P limited. The DIN:TP ratio
was a better indicator than the TN:TP ratio for nutrient limitation of phytoplankton
because TN may include a large fraction of biologically unavailable N.
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Table 9-1 Summary of studies using diatoms as biological indicators of nitrogen enrichment evaluated in the
literature since the 2008 Integrated Science Assessment for Oxides of Nitrogen and
Sulfur-Ecological Criteria.
Study Site
Ambient N Deposition
Diatom Response
Species
Reference
Three national parks 0.25 ±0.15 to
in Washington State
(Mt. Rainier, North
Cascades, and
Olympic)
Overall, only one lake (Hoh Lake) of 10 study lakes where cores A. formosa, F.
1.10 ±0.21 kg/ha/yr NH4+-N;
0.34 ± 0.04 to
1.32 ±0.19 kg/ha/yr NO3--N;
0.59 ± 0.07 to
2.42 ± 0.27 kg/ha/yr inorganic
N
were collected showed clear evidence of impacts from N
deposition based on changes in sediment diatom communities.
crotonensis, and
F. tenera
Sheiblev et al.
(2014)
Western U.S.
(Glacial National
Park, Greater
Yellowstone
Ecosystem, and
eastern Sierra
Nevada)
Sierra Nevada = 2 kg/ha/yr
(current) to 4 kg/ha/yr (early
1990s) total wet inorganic N
deposition.
GNP = 0.5-1.5 kg/ha/yr
(1980-2006).
GYE = 0.5-1.1 kg/ha/yr
(1980-2006)
A critical load of 1.4 kg N/ha/yr wet deposition changed diatom
community structure in both the eastern Sierra Nevada and the
Greater Yellowstone Ecosystem, although N deposition rates
between the two regions and the timing of diatom community
shifts were different. No diatom community changes were
observed in Glacier National Park lakes.
A. formosa and
F. crotonensis
Saros et al. (2011)
Rocky Mountains
Central Rockies NO3 and
NhU+ = 1.4-2.5 kg Nr/ha/yr. N.
Rockies = 2.0-3.4 kg Nr/ha/yr.
Lakes fed by snowpack meltwater had greater sediment diatom
taxonomic richness over the last century (35 to 54 taxa)
compared with lakes that were fed by both glacial and snowpack
meltwater (12 to 26 taxa) which are higher in NO3".
(-50% of Glacier National Park lakes and -0-20% of lakes in
central Rockies receive glacial meltwater).
Multiple
Saros et al. (2010)
Rocky Mountains
Wet deposition ranged from
0.8 to 3.2 kg N/ha/yr at study
sites.
High abundance of key indicator species of N enrichment	Multiple
(A. formosa, F. crotonensis) in lakes with low to moderate NO3"
precluded an attempt to quantify the surface water NO3"
concentrations that elicit diatom community changes in
high-elevation lakes using a diatom-based transfer function.
Arnett et al. (2012)
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Table 9-1 (Continued): Summary of studies using diatoms as biological indicators of nitrogen enrichment
evaluated in the literature since the 2008 Integrated Science Assessment for Oxides of
Nitrogen and Sulfur-Ecological Criteria.
Study Site	Ambient N Deposition	Diatom Response	Species	Reference
Rocky Mountains >3 kg N/ha/yr	A maximum growth rate of 0.2 ± 0.04/day at 0.5 |jM N for the A. formosa	Nanus et al. (2012)
diatom A. formosa was measured following N addition to identify
a NO3" threshold for this species in surface water.
Central Rocky
Mountains/Beartooth
Mountains located
along the Montana
and Wyoming
borders
Not specified
Fossil-diatom assemblages from two lakes near each other (one
snowpack-fed and one glacier-fed) indicated greater diatom
assemblage turnover in the glacier-fed lake. The glacier-fed lake
showed evidence for mild N enrichment starting approximately
1,000 yr ago, and the increase in abundances of A. formosa and
F. crotonensis occurred much earlier, suggesting glacial
meltwater was a source of N.
A. formosa and F.
crotonensis
Slemmons et al.
(2015)
Central Rocky
Mountains/Beartooth
Mountains located
along the Montana
and Wyoming
borders
Not specified
Diatom species richness from the top of sediment cores was
1.8x higher in snowpack-fed lakes compared with glacier- and
snowpack-fed lakes, whereas richness did not differ between
core bottoms (ca. 1850) of the lakes. Compared with snow-fed
lakes, N enriched glacial snowpack-fed lakes were dominated by
A. formosa and F. crotonensis.
A. formosa and
F. crotonensis
Slemmons and
Saros (2012)
Grand Teton
National Park
Estimated at 2.5 kg N/ha/yr
Directional change in benthic diatom assemblages after 1960
that is correlated with atmospheric deposition.
Benthic diatoms
Spauldina et al.
(2015)
Uinta Mountains, UT 0.02-0.04 kg km2
Four of five lakes with recent increase in productivity (between
1940 and 1960) show comparable shifts in diatom community
composition linked to atmospheric deposition of N and P. Lake
sediment records show an increasing abundance of nitrophilous
A. formosa.
A. formosa
Hundev et al.
(2014)
Mount Rainier, North 0.38-3.24 kg N/ha/yr (wet)
Cascades, and
Olympic national
parks, WA
20 taxa of phytoplankton responded to N enrichment. F. tenera
and F. crotonensis are recommended as indicators of N
enrichment in the Pacific NW. The threshold for phytoplankton
biomass growth was 13-25 |jg DIN/L.
Multiple
Williams et al.
(2016a)
Northern Cascade
Mountains, WA
1.1 to 3.4 kg N/ha/yr
No significant difference in diatom biovolume or phytoplankton
community structure between snow-fed lakes and glacial
snowpack-fed lakes.
Multiple
Williams et al.
(2016b)
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Table 9-1 (Continued): Summary of studies using diatoms as biological indicators of nitrogen enrichment
evaluated in the literature since the 2008 Integrated Science Assessment for Oxides of
Nitrogen and Sulfur-Ecological Criteria.
Study Site
Ambient N Deposition
Diatom Response
Species
Reference
National parks of the
western Great Lakes
(Superior and
Michigan) region
1.5-5 kg N/ha/yr (1980-2010) In 63% of study lakes, change in diatom community correlates
with lake sediment 815N, which in turn relates to measured Nr
inputs (deposition and indirect watershed inputs). In about 36%
of the lakes, sediment 815N was statistically correlated to some
form of deposited Nr. A number of such lakes are in watersheds
believed to have high N retention associated with shallow
subsurface flow during snowmelt. The authors consider this to be
the primary path for NO3" transport. Diatom community change
is also expected to be related to nutrient inputs indirectly by
runoff and climate variability.
Multiple, including
F. crotonensis
and A. formosa
Hobbs et al. (2016)
U.S.
Not specified
Shifts in diatom community composition away from N intolerant Multiple
species.
Pardo et al. (2011a)
North America and
Greenland
0.8 to 2 (<50°N) to <0.5
(>50°N)
A meta-analysis of 52 alpine, Arctic, and boreal montane lakes
showed increased beta diversity in the 20th century. Observed
diatom assemblage turnover in alpine and Arctic lakes is related
to temperature changes, while in mid-latitude alpine lakes
changes are linked to N deposition.
Multiple
Hobbs et al. (2010)
Baptiste Lake,
Alberta, Canada
Not specified
Lake sediment cores indicate prevalence of diatom assemblages Multiple
that favor nutrient-rich conditions for at least the last 150 yr.
From -1980 to present, a distinct increase in Stephanodiscus
hantzschii was observed.
Adams et al. (2014)
George Lake,
Killarney Provincial
Park, Ontario,
Canada
6-7 kg N/ha/yr (2013)
Relative abundance of A. formosa increased over a 20-yr period
while lake total N concentration and regional N deposition
decreased. Increases in A. formosa were directionally related to
increases in air and epilimnetic temperatures.
Multiple, including
A. formosa
Sivaraiah et al.
(2016)
Whitefish Bay,
Ontario, Canada,
also meta-analysis
of 200 lakes in North
America and Europe
<5 to <1 kg N/ha/yr
No effects of N deposition were observed on diatom community Multiple	Ruhland et al.
structure in Whitefish Bay. Observed shifts in Arctic lakes (1850) planktonic and (2008)
and temperate lakes (1970) were in response to climate change, benthic diatoms
Ha = hectare; kg = kilogram; km = kilometer; |jM = micromole; N = nitrogen; NH4+-N = nitrogen as ammonium; N03 = nitrate; P = phosphorus; yr = year.
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cc
cc
cc
Low
High	Low
Atmospheric N deposition
High
N = nitrogen; P = phosphorus.
The response ratio (RR) has no units. (A) Response to P enrichment alone (RR-P); (B) response to N enrichment alone (RR-N);
(C) response to combined N and P enrichment (RR-NP); (D) relative response to N vs. P (RR-N/RR-P, equivalent to final chl N/final
chl P). For RR-N/RR-P, a value >1 indicates stronger N limitation while a value <1 indicates stronger P limitation. Error bars
indicate ± SE. The result of t tests comparing low- and high-deposition means for each parameter are given in each panel.
Source: Elser et al. (2009^
Figure 9-2 Phytoplankton responses to nitrogen and/or phosphorus
enrichment for Rocky Mountain lakes receiving low or high
atmospheric nitrogen deposition, given as the ratio of final
chlorophyll concentration in the enriched treatment
(+phosphorus, +nitrogen, or +nitrogen + phosphorus) to the
chlorophyll concentration in the unenriched control.
1	N deposition gradient studies conducted in Sweden support findings of shifts from N to P
2	limitation in lakes. Liess et al. (2009) compared a population of lakes in northern Sweden
3	with lower N deposition (N = 7) to lakes in the south with higher deposition (N = 6).
4	Atmospheric N deposition showed a strong positive correlation with TN. In the sampled
5	lakes, northern lakes received 2 to 6 kg N/ha/yr, while southern lakes received 10 to
6	12 kg N/ha/yr. Sampling of epilithic communities indicated a weakly positive correlation
7	of atmospheric N deposition to epilithon N:P ratios, suggesting that epilithic communities
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were generally more N limited in the northern lakes and more P limited in the southern
lakes, which received higher N deposition. Bergstrom et al. (2008) evaluated
phytoplankton responses to N and P enrichment in unproductive Swedish lakes along a
gradient of atmospheric N deposition. They found that regional and seasonal patterns in
nutrient limitation were related to the level of N deposition. In the south, high N
deposition was accompanied by high lake DIN concentrations during early summer, with
subsequent P limitation. Later in the summer, DIN concentrations declined, and lakes
switched to N and P colimitation, and then to N limitation. Bergstrom et al. (2008)
concluded that N limitation is probably the natural state of these unproductive lakes and
that P limitation has been induced by increased N availability caused by atmospheric N
deposition.
Other studies in high alpine systems have shown that N deposition does not necessarily
have a consistent effect on N or P limitation. In an assessment of 29 alpine lakes and
ponds in the Canadian Rocky Mountains (Banff National Park, Alberta and Yoho
National Park, British Columbia), certain ponds appeared to be N limited while most
water bodies sampled did not show evidence of N limitation (Murphy et al.. 2010).
DIN: TP ratio analysis indicated N limitation in only 14% of the sampled water bodies. N
limitation was observed more frequently in shallow ponds than lakes, and during the late
summer sampling period, the ponds had a significantly lower mean DIN:TP ratio. In
another study from the Rocky Mountains, surface sediment samples and surface water
N03 concentrations were collected from an N deposition gradient (1 to 3.2 kg Nr/ha/yr
in wet deposition) across 46 high-elevation lakes to characterize diatom community
changes (Arnett et al.. 2012). The researchers found that even in lakes with NOs
concentrations below quantification (<1 (ig/L). diatom assemblages were already
dominated by key indicator species characteristic of moderate N enrichment, and the
authors were unable to identify a threshold. They noted that small inputs of reactive N
can shift diatom species composition along the length of the NO3 gradient and that the
lakes switch from N limited oligotrophy to P limitation. In a study of 11 lakes in northern
subarctic Sweden situated along an altitudinal/climate gradient with low N deposition
(<1 kg N/ha/yr), the lakes did not show a shift from mainly N to mainly P limitation
reported from U.S. lakes where N deposition was higher. However, mainly P limitation
was observed in high alpine lakes with rocky catchments and high DIN:TP ratios, while
mainly N and N + P colimitation was occurring in subalpine and lower mid-alpine lakes
(Bergstrom et al.. 2013). The authors attributed these observations to climate and
catchment characteristics and suggested warming would have a greater impact than N
deposition on the subset of sparsely vegetated high alpine lakes, in contrast to the
majority of subarctic lakes with N or N + P colimitation, where both warming and N
inputs will alter phytoplankton response.
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9.1.1.4. Phosphorus Deposition Interactions
P deposition has direct implications for how ecosystems respond to N deposition.
Although this appendix is focused on the biological effects of N enrichment, inputs of P
may effect shifts in lake trophic status from P to N limitation or to colimitation and
increase the total nutrient supply to water bodies affecting how the system responds to N.
Since the 2008 ISA, several meta-analyses have reported an increase in P deposition to
water bodies (Stoddard et al.. 2016; Brahnev et al.. 2015; Tipping et al.. 2014). This
recent evidence suggests that the predominantly dry deposition of fine (<10 p.) and coarse
(<100 |a) particulates ("dust") containing P play a role in the enrichment effects of N
deposition to fresh waters and their catchments. Data from the U.S. EPA National Lakes
Assessment and National Rivers and Streams Assessment were analyzed by Stoddard et
al. (2016) to determine whether total P (TP) concentrations changed between 2000 and
2014. The authors found continent-wide increases in TP, especially at sites with low
disturbance. Median stream TP concentrations increased from 26 |ig/L (2000-2004) to
56 |ig/L (2013-2014); median lake TP increased from 20 |ig/L (2007) to 37 |ig/L (2012).
From the 2004-2014 surveys, the percentage of stream length where TP was <10 |ig/L
decreased from 24.5 to 1.6; the percentage of lakes where TP was <10 |ig/L decreased
from 24.9 to 6.7 between 2007 and 2012. Additional research has corroborated the
findings of Stoddard et al. (2016) in other locations (Zhu et al.. 2016a; Brahnev et al..
2015; Tipping et al.. 2014) and has investigated shifts from N to P limitation or
colimitation, as well as P deposition's role in prolonging N limitation (Appendix 9.2.2).
In a regression analysis of deposition and water quality data in 700 upland lakes across
21 alpine regions globally, Brahnev et al. (2015) suggested that P deposition may play a
large role in alpine lake trophic status. The authors evaluated the strength in N, P, and
N:P relationships in deposition and in lakes. Their 5-year, Community Atmospheric
Model (CAM4) simulation modeling indicates that P deposition may have increased
globally by 1.4 times the preindustrial deposition rate. TIN:TP of deposition and in lakes
showed a strong relationship (r2 = 0.82,p< 0.0001) with P deposition and lake water
(r2 = 0.64,p < 0.0001) and N deposition and lake water (r2 = 0.20,p < 0.05). Where the
deposition's molar ratio of N:P was less than 20, intermittent or persistent N limitation
was observed to be induced atmospherically. The authors concluded that the chemistry of
atmospheric deposition influences nutrient limitation of alpine oligotrophic lakes by
changing nutrient ratios as well as increasing the absolute supply of nutrients to these
ecosystems. A global-scale analysis by Tipping et al. (2014) supported the observation by
Brahnev et al. (2015) about oligotrophic lakes and indicated the importance of accounting
for sustained P deposition to assess the impact of productivity of anthropogenically
emitted N deposition. The authors analyzed data on P deposition measured globally (82%
of sites were in Europe and North America) from 1954-2012 at 250 sites. They found
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that "oligotrophic lakes, tropical forests, and ombrotrophic peatlands" are the most likely
to experience significant effects of atmospheric P deposition increase.
Zhu et al. (2016a) examined P deposition in China's forest, grassland, desert, lake, marsh,
and karst ecosystems and confirmed that China's wet P deposition in terrestrial
ecosystems (0.21 kg P ha/yr) is comparable to Tipping et al. (2014) TP deposition
estimates (2013 ambient mean wet deposition = 13.69 ± 8.69 kg N/ha/yr with fluxes
ranging from 0.47 to 47.71 kg N/ha/yr). The average N:P ratio for wet deposition in
China was reported as 77:1. The mean annual N concentration at 41 monitoring sites
correlated linearly with precipitation (r2 = 0.120; p = 0.0027), and N deposition also
correlated significantly with rainfall (r2 = 0.217./; = 0.002). The authors suggested that
the high N:P ratios in atmospheric wet deposition could shift systems to P limitation,
influencing ecosystem structure and function.
9.1.1.5. Climate Modification of Ecosystem Response to Nitrogen
Nutrient inputs to fresh waters are occurring within the context of physical, chemical, and
biological modifications caused by increased annual mean temperature and magnitude of
precipitation associated with climate change (Greaver et al.. 2016). Projected shifts in
runoff and timing and quantity of flushing will alter local and regional hydrology
(Whitehead et al.. 2009). Nutrient loads to surface waters are expected to increase due to
predicted increases in surface water flow (Adrian et al.. 2009; Whitehead et al.. 2009).
Air temperature increases will lead to warmer surface waters, altering the thermal
stratification of water bodies and affecting the community composition and distribution
of aquatic biota in streams and lakes (Adrian et al.. 2009; Keller. 2007). The increased
surface water temperatures will increase the rate of algal growth (Whitehead et al.. 2009).
In regions with no evidence of increased atmospheric nutrient inputs, warming trends are
observed to enhance competitiveness of planktonic diatoms like Asterionella formosa,
which are typically associated with elevated N, indicating that climate change has
significant direct and indirect effects on algal species composition. Climate change may
thus enhance the effects observed in areas with nutrient increases alone (Ri'ihland et al..
2015). Appendix 13 includes a more detailed discussion of how climate (e.g., temperature
and precipitation) modifies ecosystem response to N loading.
9.2. Biological Indicators
Increased NOs in surface water is a chemical indicator of freshwater N nutrient
enrichment (Appendix 7.1.2.1). Many of the critical loads discussed in Appendix 95_ are
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based on NOs in surface water. A critical load is a quantitative estimate of exposure to
one or more pollutants below which significant harmful effects on specified sensitive
elements of the environment do not occur according to present knowledge KSprangcr et
al.. 2004; Nilsson and Grennfelt. 1988); Chapter 1.2.2.31. Biological indicators of
freshwater N enrichment discussed in the 2008 ISA included chlorophyll a,
phytoplankton and periphyton biomass, and changes in lake nutrient status (U.S. EPA.
2008a). Paleolimnological records of shifts in diatom community composition were also
used to assess the effects of N deposition. Dose-response relationships between N and
biological indicators were reported in the 2008 ISA, and the new literature continues to
support these findings. Diatom community shifts (Appendix 9.2.1). and phytoplankton
biomass nutrient limitation shifts (Appendix 9.2.3) have been used as a basis for
determining critical loads for nutrient enrichment (Appendix 9.5). These same biological
indicators are discussed further below along with new studies. In the current review, the
response of microbial enzymes to N and P and increased incidence of harmful algal
blooms (HABs) in freshwater habitats are discussed as possible biological indicators of
the effects of elevated N.
9.2.1. Diatoms
Diatoms are commonly used to monitor environmental conditions in water bodies over
time. While many diatom studies explain patterns of variation in community composition
in relation to environmental (Sniol and Stoermer. 2010). as well as spatial factors
(Soininen et al.. 2016; Vilmi et al.. 2016). few studies have used experimental methods to
address the processes underlying the patterns (Smol and Stoermer. 2010; Pither and
Aarssen. 2006). Changes in diatom species assemblages in paleolimnological studies of
mountain lakes disturbed only by atmospheric deposition and climate change were
reported in the 2008 ISA. Chlorophytes, such as A. formosa and Fragilaria crotonensis,
generally prefer high concentrations of N and are able to rapidly dominate the flora when
N concentrations increase (U.S. EPA. 2008a; Findlav et al.. 1999). These two
nitrophilous species of diatom are used as biological indicators of the effect of N in water
bodies; however, increased relative abundance of A. formosa has also been attributed to
lake warming in some regions where N deposition is decreasing (Sivaraiah et al.. 2016).
Table 9-1 summarizes diatom studies published since the 2008 ISA. The majority of
these studies highlight the observed influence of 20th century N deposition on diatoms,
causing an increasing abundance of nitrophilous diatoms such as A. formosa and F.
crotonensis. Appendix 93_ considers the effects of N deposition on diatom diversity.
Thresholds of diatom response are reported in Appendix 9.5.
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9.2.2.
Ratios of Nitrogen and Phosphorus
Trophic status is a way to characterize bodies of water in terms of their productivity.
Increasing deposition of N to water bodies shifts element ratios, which in turn affect algal
growth, diversity, and community structure (Appendix 9.3.2). Algal species have
different nutrient optima for growth and cellular uptake of N, and P alters the elemental
composition of primary producers at the base of the freshwater food web. When N is the
limiting nutrient, atmospheric deposition of N will increase benthic algal (Liess et al..
2009) and phytoplankton (Hessen. 2013; Elser et al.. 2010) N:P ratios. Nutritional
responses of aquatic ecosystems to atmospheric N deposition are heavily dependent on
surface water P concentrations, which may also be affected by atmospheric P inputs
(Appendix 9.1.1.4). Thus, various chemical ratios of N to P can be useful for evaluating
biological responses of water bodies affected by deposition (Table 9-1).
In the 2008 ISA, several studies reported N:P ratios in which a shift in nutrient limitation
was observed. When DIN:TP values are greater than reference values, growth
stimulation, N and P colimitation, or P limitation commonly occur (Sickman et al.. 2003).
In a Swedish lake survey reviewed in the 2008 ISA, N limitation occurred in lakes where
the DIN:TP mass ratio was less than 7 (DIN concentrations <33 (j,M). Colimitation of N
and P was found in lakes with DIN:TP ratio between about 8 and 10, and P limitation
occurred at DIN:TP values greater than 10 (Bcrgstrom et al.. 2005). Other thresholds for
N limitation were reported in the literature to occur at DIN:TP ratios <4 (Lohman and
Priscu. 1992) and <10 (Wold and Hershev. 1999). Bergstrom et al. (2005) reported an
index (DIN: [chlorophyll a:TP]) to assess the eutrophication of lakes in response to N
deposition.
Hundev et al. (2014) assessed trophic status for six remote alpine lakes in the Uinta
Mountains, UT based on TP, TN, chlorophyll a, Secchi depth, and N:P relationships. One
of six lakes was P limited, two were N limited, one varied by month, and the limitation
was uncertain in the other two. Trophic status was difficult to determine in some of the
lakes because the ratios were values that could indicate either N or P limitation depending
on the threshold used, and some lakes were on the boundary between oligotrophic and
mesotrophic. In Green Lake in Colorado's Front Range, the ratio of DIN:TP in the
epilimnion over the course of the study was consistently above 4 and averaged
16.3 ± 2.76, suggesting the lake was P limited (Gardner et al.. 2008). Nutrient ratios
(TN:TP, DIN:TP, and NOs :TP) were compared to nutrient bioassays to evaluate the use
of these ratios in predicting the limiting nutrient. There was agreement in 29% of the
bioassay results, suggesting that nutrient ratios are not the best predictor of nutrient
limitation in this subarctic region.
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In an in situ mesocosm bioassay study of two subalpine lakes in the Sierra Nevada, CA,
Heard and Sickman (2016) modeled the effective dose of N addition along an N gradient
(N as KNO3 and NaNCh ranging from 0 to 50 (imol/L) with P addition as KH2PO4 held
constant at 1.5 (imol/L. DIN:TP ratios in the two lakes were <4.0. Given that DIN:TP
ratios <0.6 indicate N limitation and DIN:TP ratios > 0.6 indicate intermediate limitation,
both lakes were deemed N limited. The authors observed that the P addition delayed the
shift to P limitation (prolonged N limitation) for phytoplankton. Effective doses for N for
the two lakes were lower than the other study lakes receiving N addition only. They
acknowledged the spatial and temporal variation in lake P concentrations, as well as
concern for increased P deposition.
In a comparison of diatom-inferred lake nutrient records and surface water monitoring
data (collected since the early 1980s) from Baptiste Lake in the Canadian Rockies,
Adams et al. (2014) found total Kjehldahl nitrogen (TKN) was a significant predictor of
chlorophyll a, where chlorophyll a was independent of TP. They observed that the lake
experienced eutrophic conditions for >150 years based on diatom sediment core data.
Diatom-inferred TKN data tracked current TKN dynamics measured in the water column.
9.2.3. Phytoplankton Biomass Nitrogen (N) to Phosphorus (P) Limitation Shift
As described in Appendix 9.1.1.3. lakes may shift from N limitation to P limitation with
elevated N deposition and this can be assessed considering the RR-N compared to the
RR-P I Figure 9-2; (Elser et al.. 2009b)l. Phytoplankton biomass growth may shift from N
to P enrichment, altering lake primary production (Hcsscn. 2013). Williams et al. (2017a)
used the RR-N/RR-P to define a biological threshold of RR-N/RR-P = 1 above which
phytoplankton biomass P limitation is more likely than N limitation and developed
critical loads for western U.S. mountain lakes (Appendix 9.5).
9.2.4. Periphyton/Microbial Biomass
Periphyton mats are biofilms of algae, cyanobacteria, fungi, microinvertebrates, organic
detritus, inorganic particles, and heterotrophic microbes imbedded within a matrix and
attached to submerged substrates in aquatic systems (e.g., stream or lake bottoms). In the
2008 ISA, no studies reported resource requirements for periphyton, although several
papers described stimulated growth with N amendments in ecosystems throughout the
U.S. (Annex C of the 2008 ISA), including streams in Alaska, Arizona, Iowa, Texas,
Minnesota, and Missouri and lakes in California, Colorado, and Massachusetts (U.S.
EPA. 2008a). Additional lake bioassay experiments that enriched the water column down
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into the sediments found enhancement of periphyton growth on bioassay container walls
in experiments in California, Wyoming, and Massachusetts (Smith and Lee. 2006;
Nvdick et al.. 2004; Axler and Reuter. 1996). Strong N limitation of benthic algae has
also been inferred in streams of Arizona (Grimm and Fisher. 1986). California (Hill and
Knight. 1988). Missouri (Lohman et al.. 1991). and Montana (Lohman and Priscu. 1992).
Since the 2008 ISA, few studies on N effects on periphyton biomass have been identified.
In Ditch Creek, WY, a relatively undisturbed stream with N accumulation primarily from
high N2 fixation, biofilm growth on rocks increased after snowmelt and N2 fixers
dominated the algal assemblage (Kunza and Hall. 2014). A shift to non-N2 fixing taxa
was observed later in the season. In this stream with little to no anthropogenic influence,
N fixation was higher than denitrification. Burrows et al. (2015) assessed microbial
respiration and biomass using nutrient-diffusing substrata in 20 boreal streams in
Sweden. An increase in N availability led to increased microbial activity. Microbial
biomass was primarily N limited, and distinct microbial communities were associated
with inorganic N in stream water. In a series of nutrient-diffusing substrata studies across
rivers in the U.S. mountainous West, arid West, and the Midwest, Reisingeretal. (2016)
assessed nutrient limitation patterns of benthic biofilms. Regarding regional differences,
there was little evidence of N limitation in midwestern rivers due to high nutrient
concentrations, while nutrient limitation of biofilms was common in the summer in
mountainous western rivers. Increasing developed lands decreased the probability of
nutrient-limited river biofilms. In a meta-analysis of nutrient-diffusing substrate studies
from North America Beck et al. (2017). broad spatial factors such as ecoregion described
most of the variation in nutrient limitation. Variables affecting algal biomass response to
N included land use, riparian canopy cover, the presence of soluble reactive P, and
season.
Microbial communities involved in plant litter decomposition in streams have been
shown to be altered by nutrient concentrations. Most studies have examined the effects of
N and P in combination. However, Fernandes et al. (2014) conducted a series of N
addition studies in microcosms to assess leaf litter decomposition with increasing N
concentration and temperature in streams. In general, increased temperature led to an
increase in decomposer activity and a decrease in the amount of N needed. Field studies
in six low-order streams spanning an N gradient in the Ave River basin, Portugal,
suggested that eutrophication modulates leaf litter decomposition processes (Lima-
Fernandes et al.. 2015). Leaf litter diversity synergistically affected leaf litter
decomposition while biomass fungal and invertebrate decomposers increased with stream
eutrophication status but decreased in the most eutrophic of the six streams. Dunck et al.
(2015) reported decreased primary production and leaf litter decomposition in highly
eutrophic streams and streams with little human influence compared to streams that were
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intermediate along the trophic gradient. Kominoski et al. (20151 observed that microbial
litter breakdown rates increased across low to moderate nutrient enrichment gradients in
experimental stream channels where N:P ratios were varied to allow for examination of N
effects. A meta-analysis of plant matter decomposition in streams by Ferreira et al.
(2015) suggested that the effects of nutrient inputs might be strongest in oligotrophic
streams due to the low background nutrient concentrations and high magnitude of
nutrient enrichment in these systems.
9.2.5. Chlorophyll a
The concentration of chlorophyll a, a pigment present in all photosynthetic organisms, is
a common measure of algal productivity and is hence an easily documented biological
indicator of change in aquatic ecosystem productivity. Quantification of chlorophyll a is
one approach for estimating phytoplankton biomass. Chlorophyll a is used as the primary
indicator of trophic state in the U.S. EPA National Lakes Assessment and as a water
quality indicator in many state and federal monitoring programs (U.S. EPA. 2016h.
2009b). Chlorophyll a is being used as an indicator of nutrient enrichment in U.S. EPA's
National Nutrient Program lYU.S. EPA. 1998b); Appendix 7.1.61. The U.S. EPA is
working with the states to develop numeric nutrient criteria to better define levels of N
and P that affect U.S. waters. The numeric values include both causative (N and P) and
response (chlorophyll a, turbidity) variables to assess eutrophic conditions. The U.S.
EPA's National Nutrient Program to reduce eutrophication in water bodies developed
recommended nutrient criteria for rivers and streams in 14 U.S. ecoregions for the states
to use as a starting point to develop their own criteria (U.S. EPA. 1998b).
Surveys and fertilization experiments reported in the 2008 ISA show increased inorganic
N concentration and aquatic ecosystem productivity, as quantified by chlorophyll a
concentration, to be strongly related (U.S. EPA. 2008a). At the time of the 2008 ISA,
increases in lake phytoplankton biomass (as chlorophyll a) with increasing N deposition
were reported in several regions, including the Snowy Range in Wyoming (Lafrancois et
al.. 2003b) and across Europe (Bergstrom and Jansson. 2006). A meta-analysis of
enrichment bioassays in 62 freshwater lakes of North America reported in the 2008 ISA
found algal growth enhancement from N amendments to be common in slightly less than
half of the studies (Elseretal.. 1990). There was a mean increase in phytoplankton
biomass of 79% in response to N enrichment (Elseretal.. 1990). This meta-analysis was
repeated, incorporating study sites from multiple countries and a much larger data set,
with similar results (Elser et al.. 2007). Chlorophyll a continues to be a common
biological indicator of N nutrient enrichment in the research literature from 2008 to
present. Dodds and Smith (2016) conducted a review of N and P dynamics in stream
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ecosystems and concluded that both N and P are strongly correlated to chlorophyll and
algal biomass. In the western U.S., Elser et al. (200%) examined chlorophyll a response
in Rocky Mountain, CO lakes where atmospheric deposition ranged from 2 to
7 kg N/ha/yr. Concentrations of chlorophyll were 2 to 2.5 times greater in
high-deposition lakes relative to low-deposition lakes.
Since the 2008 ISA, nutrient threshold values for chlorophyll a responses have been
identified for a subset of western mountain lakes. Williams et al. (2016a) calculated an
average DIN threshold of 13 |ag DIN/L for increased chlorophyll a concentration across
nine western mountain lakes (25 |_ig DIN/L for increased chl a concentration beyond
interannual variation). More than 25% of 207 lakes in Mount Rainier and the northern
Cascades exceeded the 13 p.g DIN/L threshold. Heard and Sickman (2016) determined
threshold values for phytoplankton growth for Sierra Nevada lakes. The doses of N that
characterize the initial phytoplankton growth response (10% effective dose), rapid
phytoplankton growth (50% effective dose), and saturating nutrient level (90% effective
dose) in early- and late-season conditions within two high-elevation lakes were
established. These doses were then compared with monitoring data from lakes within
Yosemite, Sequoia, and Kings Canyon national parks. The range of threshold values for
the 10-50% effective doses were in the range of 0.3 to 4 (.iniol N/L (5-56 |_ig N/L). The
50% effective doses were exceeded by 18% (late season) to 29% (early season) of
monitored lakes, suggesting that N inputs via atmospheric deposition affect
phytoplankton in many lakes in this region. The lakes that exceeded effective doses
tended to be at high elevation on steep, north-facing slopes with limited vegetative
growth.
Chlorophyll a can also be used as a bioindicator of lake response over decadal and longer
periods. Hundev et al. (2014) took sediment cores from six remote alpine lakes in Utah to
assess trends in lake productivity. In five of the six lakes, sedimentary chlorophyll a and
percentage of organic matter was observed to be relatively constant from the beginning of
the record (mid 1800s using 210Pb dating; 1200 to 1800 estimated by linear regression)
until 1940-1960 when lake production progressively increased.
As reported in the 2008 ISA and discussed further in Appendix 7.1.2.9. dissolved organic
carbon (DOC) affects acidity and N cycling and is increasing in some U.S. surface waters
(Monteith et al.. 2007; Evans et al.. 2006). Recent studies indicate different
phytoplankton responses to N and dissolved organic matter (DOM) inputs depending on
nutrient status of the lakes and background DOC (Deininger et al.. 2017a; Daggett et al..
2015). Daggett et al. (2015) selected a low DOC, N and P colimited water body (Jordan
Pond in Acadia National Park, ME) and an N limited lake with higher DOC (Sargent
Lake in Isle Royale National Park, MI) to assess the effects of an N gradient on algal
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biomass following addition of DOM. Both of these Class I areas have a similar N
deposition rates (6-7 kg total inorganic N/ha/yr wet deposition). DOM addition
stimulated phytoplankton biomass in both lakes regardless of nutrient limitation or
background DOC concentration. Phytoplankton biomass in N limited Sargent Lake
increased with both N and DOM inputs, while in N and P colimited Jordan Pond, algae
were sensitive primarily to DOM addition. Appendix 9.3.2.3. discusses a whole-lake N
fertilization study along a gradient of DOC.
Not all studies have shown a positive relationship between N inputs and increased
chlorophyll a. In the Canadian Rockies, Murphy et al. (2010) performed nutrient
enrichment bioassays on 29 water bodies in Banff and Yoho national parks using
1 mg N/L additions. They found that N amendment did not significantly increase final
total phytoplankton chlorophyll concentration across all of the water bodies sampled
during either early or late summer 2007. Although phytoplankton collected from ponds
showed several responses to nutrient amendment, the effect of N on total chlorophyll did
not differ significantly between pond and lake communities. Harvested final total
chlorophyll concentrations from the bioassays were significantly higher for ponds than
lakes. In water bodies in Wapusk National Park, Manitoba, a significant linear
relationship occurred between chlorophyll a and TP across all ponds but not between
chlorophyll a and total N [TN; (Svmons et al.. 2012)1. In another study in Colorado's
Front Range in Green Lake Valley (part of the Niwot Ridge Long-Term Ecological
Research Project), Green Lake 4, a well-studied lake in the area, was sampled weekly
throughout the ice-free summer period to assess chlorophyll response (Gardner et al..
2008). Epilimnetic chlorophyll a concentrations peaked at 4 (ig/L in late July. In the same
study, N03 addition to an in situ mesocosm on the lake (930 (ig/L NOs . for a final
exposure of 1,240 (ig/L NO3 given background concentrations) did not increase algal
biomass significantly in comparison with the control, while phytoplankton chlorophyll a
increased in P and P + NO3 additions, indicating the lake was P limited during the
summer.
Although hypothesized to be driven by industrial point source emissions, an increase in
aquatic primary production in some lakes in the Athabasca Oil Sands Region in Alberta,
Canada appears to follow regional patterns of annual and seasonal changes in
temperature. In a comparison of spectrally inferred chlorophyll profiles of lake sediments
in 23 diverse undisturbed lakes, Summers et al. (2016) observed higher chlorophyll a
concentrations in surface sediments compared with concentrations inferred for the period
before oil sands development, irrespective of lake morphology, landscape position, N
deposition, or other limnological characteristics. Spatial and temporal patterns in inferred
chlorophyll a were positively correlated with annual and seasonal changes in temperature
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and were consistent across the region, suggesting a regional impact on productivity rather
than impacts associated with point source emissions.
Some research has been conducted regarding the biological impacts of changes in N
retention. Heini et al. (2014) examined chlorophyll a at various depths in a lake in
southern Finland, and found that algae and cyanobacteria can affect water chemistry
instead of just responding to it. Differences in phytoplankton strongly affected observed
short-term differences in chemical properties. The chemical conditions in the deeper
waters during summer were generally more stable than in the layers near the surface of
the lake.
Since the 2008 ISA lake surveys, fertilization experiments and nutrient bioassays
continue to show a strong correlation between N inputs and chlorophyll a. Determination
of additional thresholds for chlorophyll a in remote high-elevation lakes have further
linked this indicator to atmospheric inputs of N. The relative importance of DOC in
modulating lake response to nutrients is better understood than at the time of the
2008 ISA (Appendix 7.1.2.9). The use of chlorophyll a as an indicator in national lake
and stream assessments and for state numeric nutrient criteria (Appendix 7.1.6) is further
supported by studies reported in this ISA.
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Table 9-2 Summary of additional evidence for nitrogen limitation on productivity of freshwater ecosystems
that has been evaluated since the 2008 Integrated Science Assessment for Oxides of Nitrogen and
Sulfur-Ecological Criteria.
Region
N Deposition
N Addition
Biological Effects
Reference
Sierra Nevada,
CA
Not specified
Gradient of N (as
KNOs and
NaNC>3) ranging
from 0 to
50 pmol/L
(dose-finding
study), then range
used to model
biological
response was
only up to about
18 pmol/L
10, 50, and 90% effective doses for excess phytoplankton growth were
calculated for two high-elevation lakes. The modeled doses were compared
with lake chemistry monitoring data to assess nutrient status. The 50%
effective doses were exceeded by 18% (late season) to 29% (early season)
of monitored lakes, suggesting that N inputs via atmospheric deposition
affect phytoplankton in many lakes in this region. The threshold for
stimulation of phytoplankton (10 to 50% ED) was 5-56 |jg N/L.
Heard and Sickman
(2016)
Rocky Mountains, 2-7 kg N/ha
CO
None
Atmospheric N deposition increased the stoichiometric ratio of N and P in
lakes. Phytoplankton growth in 16 high N deposition lakes (~7 kg N/ha) was
P limited whereas in 20 low N deposition lakes (~2 kg N/ha) growth was
primarily N limited.
Elseret al. (2009a)
Rocky Mountains,
CO
High >6 kg N/ha/yr;
low <2 kg N/ha/yr
(NADP)
Enrichment of	Phytoplankton response to increased inputs of N was inferred from
7.5 pmol/L N (as chlorophyll changes in bioassay data from 20 low N deposition lakes and
NH4NO3) for N,	16 high N deposition lakes. Concentrations of chlorophyll and seston C
there was also a	were 2-2.5 times higher in the high N deposition lakes relative to the low N
P and N + P	deposition lakes, while high-deposition lakes also had higher seston C:N
treatment	and C:P (but not N:P) ratios.
Elseret al. (2009b)
Rocky Mountains,
CO
Not specified
Added 930 pg/L
NO3"; with
background,
exposure was
1,240 pg NO3-;
added 93 pg/L
TDP
In in situ mesocosm experiments with water from Green Lake 4, chlorophyll Gardner et al. (2008)
a did not increase significantly with addition of NO3" in comparison with the
control, while P and N + P treatments resulted in increases.
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Table 9-2 (Continued): Summary of additional evidence for nitrogen limitation on productivity of freshwater
ecosystems that has been evaluated since the 2008 Integrated Science Assessment for
Oxides of Nitrogen and Sulfur-Ecological Criteria.
Region
N Deposition
N Addition
Biological Effects
Reference
Utah
Not specified
None
Using sediment core data from six remote alpine lakes, chl a and
percentage of organic matter are relatively constant from the beginning of
the record until 1940-1960 in five of six lakes, when production
progressively increased. The authors found the limiting nutrient difficult to
identify.
Hundev et al. (2014)
Alberta, Canada Not specified
None
TKN was a significant predictor of chl a in Baptiste Lake where chl a was
independent of TP measured in the water column.
Adams et al. (2014)
Wapusk National Not specified
Park, Canada
N (NH4NO3) and
P (KH2PO4) were
added to increase
the nutrient
concentrations by
10x mean
ambient
concentrations
Nutrient enrichment bioassays were conducted on water samples from
21 lakes and ponds with chl a concentrations from 1.2 to 10.8 |jg/L. In total,
13% of the lakes and ponds were found to be N limited, 26% P limited, 26%
colimited, and 38% did not respond to either N or P additions.
Svmons et al. (2012)
Sweden
Gradient rates of N
deposition ranging
from 100 to
1,000 kg N/km2/yr
in four regions
during summers of
2004-2006
N: 1 mg/L and/or
the
concentrations of
P by 100 |jg/L
(molar N:P ratio,
23:1)
Phytoplankton were N limited in northern lakes (low N deposition), while in
the southern lakes higher N deposition was accompanied by increased lake
DIN concentrations and a switch from N to P limitation. N limitation was
common during the summer. As summer progressed, P limitation in these
systems switched to duel and colimitation of N and P, and to N limitation,
due to exhaustion of the DIN pool in the lakes.
Berqstrom et al. (2008)
Sweden
<1 kg N/ha/yr
Nutrients were
added to increase
[N] as NH4NO3 by
100 |jg/L
(7.2 pmol/L)
and/or [P] as
KH2PO4 by
10 |jg/L
(0.3 pmol/L)
In phytoplankton nutrient addition bioassays using water from
high-elevation lakes, phytoplankton was subject to P limitation and became
increasingly N and NP colimited at lower elevation. Chlorophyll
concentrations in the bioassays were lower with increasing elevation and
this pattern held over the whole growing season.
Berqstrom et al. (2013)
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Table 9-2 (Continued): Summary of additional evidence for nitrogen limitation on productivity of freshwater
ecosystems that has been evaluated since the 2008 Integrated Science Assessment for
Oxides of Nitrogen and Sulfur-Ecological Criteria.
Region
N Deposition N Addition
Biological Effects
Reference
Sweden
2 to 12 kg/ha/yr None
N deposition positively related to total N and total P. The highest proportion
of N fixing cyanobacteria (although only consisting of 5% of the algal
biovolume) was found where N deposition was lowest. Epilithic periphyton
N:P ratios increased with higher N availability from deposition.
Liess et al. (2009)
C = carbon; ED = effective dose; KH2P04 = monopotassium phosphate; N = nitrogen; NADP = National Atmospheric Deposition Program; NH4N03 = ammonium nitrate;
N03" = nitrate; P = phosphorus; TDP = total dissolved phosphorus; TKN = total Kjehldahl nitrogen; TP = total phosphorus.
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9.2.6. Potential Biological Indicators
In addition to widely used biological indicators of nutrient enrichment (chlorophyll a,
periphyton/microbial biomass, diatoms), freshwater HABs and enzymes are altered by N
availability. Formation of HABs are more relevant downstream where multiple sources of
N contribute to elevated nutrient levels sufficient for bloom formation. Only a limited
number of studies have used enzymes to date.
9.2.6.1. Harmful Algal Blooms
Freshwater HAB formation has become more prevalent in the U.S. in recent decades
(Lopez et al.. 2008). coinciding with increased N in surface waters. Although risk of
bloom formation remains low for high-elevation oligotrophic water bodies with
atmospheric N as the dominant source, atmospheric N may combine with other sources to
contribute to total N loading in downstream lacustrine and riverine systems. Freshwater
HABs can affect taste and odor of drinking water, lead to hypoxic conditions, impact
recreational uses of surface waters, and produce toxins harmful to humans, domestic
animals, and wildlife (Lopez et al.. 2008). The major harmful algal group in freshwater
environments are the cyanobacteria (blue-green algae). Cyanobacterial toxins are
produced by several genera including Microcystis, Anabaena, Nodularia,
Aphanizomenon, Cylindrospermopsis, and Oscillatoria. These toxins may target the liver
(hepatotoxins such as cylindrospermopsins and microcystins), the nervous system
(neurotoxins such as anatoxins and saxitoxins), or the skin (dermatoxins).
There is evidence of the role for N in freshwater HABs in water bodies across the U.S. In
western Lake Erie, cyanobacterial growth was N limited during bloom conditions in late
summer, and the N fixing cyanobacterium Anabaena became dominant following the
observed N limitation (Chaffin et al.. 2013). In a survey of eutrophic midcontinent lakes,
microcystins (common cyanobacterial toxins associated with HABs), were detected in all
blooms (Graham et al.. 2010). In another U.S. survey, TN was strongly correlated to
microcystin concentration in lakes and reservoirs (Beaver et al.. 2014). The highest
cyanobacteria abundance was observed in the Midwest where agriculture is a dominant
land use, and TN concentrations are higher. Using data from the U.S. EPA National
Lakes Assessment (U.S. EPA. 2009b). Yuan et al. (2014) modeled a threshold for the
probability of occurrence of Microcystis, a common non-N2 fixing cyanobacterial genus.
In their analysis, the frequency of occurrence of high microcystin concentrations
depended most strongly on TN, with weaker associations to chlorophyll a. The calculated
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threshold is based on the range of possible combinations of TN and chlorophyll a and the
World Health Organization drinking water provisional guideline for microcystin of
1 (ig/L (WHO. 1998). The recommended threshold ranges to protect against frequency of
occurrence of high microcystin concentrations (0.1 MC < 1 (.ig/L) were 570 or
1,100 (ig/L TN paired with chlorophyll a concentrations of 37 and 3 (ig/L, respectively
(Yuan et al.. 2014). Decreasing the frequency of microcystin occurrence to 5% gave TN
concentrations of 250 or 400 (ig/L paired with chlorophyll a concentrations of 14 and
1 (.ig/L. In the most recent U.S. EPA National Aquatic Resource Surveys, microcystin
was detected in 12% of wetlands where water depth was sufficient to allow for
microcystin sampling (U.S. EPA. 20161) and in 39% of lakes (U.S. EPA. 2016h). The
U.S. EPA has recently provided human health advisories on allowable limits of 0.3 (ig/L
microcystins and 0.7 (ig/L cylindrospermopsin in drinking water over a 10-day exposure
period (U.S. EPA. 2015a. b).
Since the 2008 ISA, additional studies have indicated that the availability and form of N
influences algal bloom composition and toxicity, and inputs of inorganic N selectively
favor some HAB species. A recent bloom in Lake Okeechobee in Florida was dominated
by Microcystis, which depends on DIN for growth (Paerl and Scott. 2010). Multiyear
monitoring data from western Lake Erie showed that microcystin concentration peaks
coincided with inorganic N and that microcystin was significantly lower in years with
less inorganic N loading (Gobler et al.. 2016). In nutrient amendment experiments with
water collected from Lake Agawam in New York, abundances of toxic strains of
Microcystis were enhanced more than nontoxic strains by inorganic N, while nontoxic
strains responded to organic N more frequently than inorganic N (Davis et al.. 2010).
Microcystis populations were stimulated more frequently by N than by P in the assays.
Similarly, in a series of nutrient addition experiments using water from Sandusky Bay,
Lake Erie, bloom growth and microcystin concentrations responded more frequently to
the addition of inorganic and organic forms of N than to P addition, indicating that N
inputs may affect bloom size and toxicity (Davis et al.. 2015). Donald etal. (2011)
reported differential responses of phytoplankton to various forms of N in mesocosm
experiments in Wascana Lake, Saskatchewan. In this naturally P rich lake, addition of
NIL+ and urea promoted nonheterocystous cyanobacteria and algae, while increased
chlorophytes and some cyanobacteria were observed with NOs and urea. Microcystin
production increased with added N, although the response varied by the form of N and
the predominant algal taxon. In the same mesocosms, species-specific analyses indicated
144 individual taxathat exhibited distinct responses to N addition: 45 species showed
stimulated growth, 93 species had a limited response, and 6 species had suppressed
growth (Donald et al.. 2013).
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9.2.6.2.
Enzymes
A recent study analyzed enzyme activity to characterize nutrient limitation in aquatic
systems. Microbial enzyme response to changes in N and P was found to vary in
terrestrial and aquatic compartments in Bear Brook watershed in Maine, the site of a
whole-watershed N enrichment experiment (Mincau et al.. 2014). Stronger effects were
typically observed in aquatic habitats. Although not explicitly the focus of the study, the
authors imply that P limitation is caused by N availability. Since 1989, ammonium
sulfate ([NH/J2SO4) has been applied bi-monthly to the watershed by helicopter,
(25.2 kg N/ha/yr). The activity of three enzymes (b-l,4-glucosidase [BG],
b-l,4-N-acetylglucosaminidase [NAG], acid phosphatase [AP]) in soil, leaf litter in both
terrestrial and stream habitats, and in stream biofilms were compared with the activities
of these enzymes in a reference watershed. In both terrestrial and aquatic habitats in the
Bear Brook watershed, BG and NAG activities were unaffected. Stream biofilms had
fivefold higher AP activity, while stream litter had eightfold higher AP activity. Increased
AP activity is indicative of enhanced P limitation in the treated watershed. Shorter
experimental P enrichments were used to characterize potential P limitation under
ambient and elevated N availability. In streams, the effects of acute P additions were
strongest with reduced AP activity and increased BG activity.
9.3. Community Composition, Species Richness, and Diversity
Increased N deposition to water bodies, either directly or via N leaching from soils,
increases available nutrients to aquatic biota and alters the balance of N and P. As
reported in the 2008 ISA, differences in resource requirements allow some species to gain
competitive advantage over others when nutrients are added, causing shifts in freshwater
community composition and structure as N concentration increases (Saros et al.. 2005;
Lafrancois et al.. 2004; Wolfe et al.. 2003). In eutrophic systems, water quality changes
associated with excess N nutrient inputs like lowered DO and algal blooms, which alter
habitat by covering up substrate, can also lead to declines in biodiversity (Hernandez et
al.. 2016). These pathways of N impacts have been identified as a threat for 50 aquatic
invertebrate species (mollusks) and 14 fish species that are listed, or candidates for
protection, under the U.S. Federal Endangered Species Act (Hernandez et al.. 2016). The
authors did not consider the sources of N in their analysis of biota affected by N
pollution.
Evidence for N effects on biodiversity in phytoplankton, zooplankton, and
macroinvertebrates in the 2008 ISA included observations from paleolimnological
studies, bioassays, and mesocosm and laboratory experiments. Community shifts and
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decreased species richness of phytoplankton have been described in multiple studies
while fewer studies have considered zooplankton. The new literature described below
continues to report the effects of N enrichment on algal biodiversity and shows limited
evidence of effects at higher trophic levels (Table 9-3).
9.3.1. Archaea and Bacterial Diversity
The recent identification of ammonia-oxidizing archaea (AOA) and the classification of
these microbes in a distinct and novel phylum within Archaea (Thaumarchaeota) has led
to a large research effort to characterize their prevalence, community dynamics, and
ecological distribution (Schlcpcr and Nicol. 2010; Spang et al.. 2010). Spatio-temporal
dynamics and community structure of AOA appear to be affected by lake trophic status
and form of N (Mukhcrjcc et al.. 2016; Bollmann et al.. 2014; Berdieb et al.. 2013). In a
survey of nitrifying microbes in Lake Erie and Lake Superior, AOA outcompeted
ammonia-oxidizing bacteria (AOB) where ammonium concentrations are lower
(i.e., oligotrophic Lake Superior 0.287 (.iM NH4 versus. 2.45 (.iM in mesotrophic Lake
Erie) (Mukhcrjcc et al.. 2016).
9.3.2. Phytoplankton Diversity
Survey data, paleolimnological studies, and fertilization experiments in the 2008 ISA
reported species changes and reductions in plankton diversity in sensitive high-elevation
lakes in the western U.S. in response to increased availability of N (U.S. EPA. 2008a).
Available data reviewed in the 2008 ISA suggested that the increases in total N
deposition do not have to be large to elicit an ecological effect. Interlandi and Kilham
(2001) demonstrated that the highest species diversity was maintained at very low N
levels (<3 (.iM N) in lakes in the Yellowstone National Park region. A hindcasting
exercise determined that a change in Rocky Mountain National Park lake algae occurred
between 1850 and 1964 at only about 1.5 kg N/ha/yr wet N deposition (Baron. 2006).
Similar changes inferred from lake sediment cores of the Beartooth Mountains of
Wyoming also occurred at about 1.5 kg N/ha/yr deposition (Saros et al.. 2003).
Preindustrial inorganic N deposition is estimated to have been only 0.1 to 0.7 kg N/ha/yr
based on measurements from remote parts of the world (Holland et al.. 1999; Galloway et
al.. 1995). In the western U.S., preindustrial, or background, inorganic N deposition was
estimated by Holland et al. (1999) to range from 0.4 to 0.7 kg N/ha/yr.
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Table 9-3 Summary of studies on nitrogen effects on species composition and biodiversity that have been
evaluated in the literature since the 2008 Integrated Science Assessment for Oxides of Nitrogen and
Sulfur-Ecological Criteria.
Region
Endpoint
Deposition
kg N/ha/yr
N Addition
Observation
Species
Ecosystem
Type
Reference
Isle Royale Phytoplankton 6-7 kg total 0, 5, 10, 20, or Chlorophytes had a reduced
National Park, community inorganic	40 [jg/L	response to DOM addition when N
Ml and Acadia structure	N/ha/yr wet	was added to the N limited lake
National Park.	deposition	(Sargent Lake in Isle Royale
ME	National Park). Increase in
abundance of diatoms (F.
crotonensis and Tabellaria
flocculosa) with DOM addition to
the lake.
Multiple
Boreal lakes
Daggett et
al. (2015)
Lake Tahoe,
CA and NV
Survey
Not specified None
Endemic benthic invertebrate taxa
have declined by 80 to 100%
since a survey of Lake Tahoe was
conducted in the 1960s. Changes
in the density and assemblage
structure of benthic invertebrates
mirrors increases in nutrient
enrichment and non-native
species in the lake. Macrophyte
occurrence has also declined.
Amphipods Stygobromus
lacicolus and S.
tahoensis declined
99.9%
No endemic turbellarians
but abundant in 1960s.
Endemic stonefly Capnia
lacustra decreased
93.5%
Endemic ostracods
Candona tahoensis
density decreased 83.4%
Subalpine,
oligotrophic
lake in
California
Caires et al.
(2013)
Ditch and Algal biofilm
Spread Creeks assemblage
in Grand Teton
National Park,
WY and Spring
Creek near
Wilson, WY
Not specified
0.5 M NaNOs;
0.5 M KH2PO4;
0.5 NaNOs, and
0.5 KH2PO4
Nutrient additions altered algal
biofilm assemblages in the
streams. NO3" addition inhibited
N2 fixer accumulation.
Phytoplankton
Streams
Kunza and
Hall (2013)
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Table 9-3 (Continued): Summary of studies on nitrogen effects on species composition and biodiversity that have
been evaluated in the literature since the 2008 Integrated Science Assessment for Oxides
of Nitrogen and Sulfur-Ecological Criteria.
Region
Endpoint
Deposition
kg N/ha/yr
N Addition
Observation
Species
Ecosystem
Type
Reference
Front Range of
the Colorado
Rocky
Mountains
Phytoplankton
community
composition
Not specified
Added 930 pg/L
NO3" (with
background,
exposure was
1,240 pg NO3-;
added 93 pg/L
TDP)
Diatom abundance increased and
phytoplankton species
composition shifted in
nutrient-enriched mesocosms.
Principal component analysis
suggested that 21 % of the
variance in phytoplankton
community composition was
related to added nutrients, while
34% of the variance was due to
seasonal changes.
Phytoplankton
Alpine lake Gardner et
al. (2008).
12 alpine
ponds in Banff
National Park,
Canadian
Rockies
Phytoplankton
and periphyton
biomass/
zooplankton
biomass/
community
diversity
6.5-2.3 kg N/ha
/yr (2000-2010)
Highest levels
(30-90 kg N/ha/
yr) downwind of
urban and
agricultural
sources.
1 mg N/L
30 pg P/L
Simulating
deposition = 20 k
g/ha/yr
Periphyton outcompeted
phytoplankton for limiting
nutrients, indicating the
importance of considering both
benthic and pelagic primary
producers. High grazing pressure
by fairy shrimp affected results
predicted from chemical inference
and bioassay results regarding
effects of added nutrients on algal
communities.
Phytoplankton,
zooplankton, invertebrate
grazers, (fairy shrimp,
Anostraca: Branchinecta
paludosa)
Fishless,
nonglacial
ponds located
above tree
line
Vinebrooke
et al. (2014)
Banff	Phytoplankton
National Park,	abundance,
Canadian	effect on
Rockies	zooplankton
1,200 mg/L N
added to aquaria
with and without
sediment
A taxonomic shift from
omnivorous, raptorial-feeding
copepods to more effective
filter-feeding herbivorous daphnids
was observed with warming and N
addition. Warming and N
fertilization increased
phytoplankton abundance while
herbivory by Daphnia decreased
but only in the presence of
sediment. Effects of warming and
N differ both between trophic
levels and aquatic alpine habitats.
Zooplankton: herbivorous Alpine lake
Daphnia middendorffiana
and omnivorous
Hesperodiaptomus
arcticus
Thompson
et al. (2008)
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Table 9-3 (Continued): Summary of studies on nitrogen effects on species composition and biodiversity that have
been evaluated in the literature since the 2008 Integrated Science Assessment for Oxides
of Nitrogen and Sulfur-Ecological Criteria.
Region
Endpoint
Deposition
kg N/ha/yr
N Addition
Observation
Species
Ecosystem
Type
Reference
Wapusk
National Park,
Canada
Phytoplankton
abundance,
effect on
zooplankton
Not specified
N (NH4NO3) and
P (KH2PO4) were
added to
increase the
nutrient
concentrations by
10x mean
ambient
concentrations
N limited lakes had statistically
significant different phytoplankton
community composition with more
chrysophytes and Anabaena spp.
compared to all other lakes.
41 phytoplankton taxa Subarctic
including	lakes and
Chlamydomonas spp., ponds
Sphaerocystis spp.,
Diatom a spp. and
Crugienella spp.
Svmons et
al. (2012)
Northern
Sweden
Phytoplankton
response to
whole lake
inorganic N
fertilization
<2	Dissolved KNO3
14 M N as KNOs
(in 2012) and 16
M HNOs (in
2013)
As DOC increased along a
gradient, community composition
shifted from nonflagellated toward
high DOC-adapted flagellated
autotrophs in the three fertilized
lakes. In the same set of lakes,
although phytoplankton biomass
increased, net zooplankton
responses were modest and
attributed by the authors to
incompatible stoichiometry of food
(phytoplankton) to consumers
(zooplankton)
Taxonomic groups:
chrysophytes,
crytophytes,
chlorophytes, diatoms,
dinoflagellates,
euglenophytes,
cyanobacteria,
picophytes.
Functional groups:
mixotrophic flagellates,
autotrophic flagellates,
nonflagellates,
heterotophic flagellates,
cyanobacteria,
picophytoplankton.
Boreal lakes
Deininqer
et al.
(2017a)
Northern
Sweden
Pelagic food
web response
to whole lake N
fertilization
<2	Dissolved
potassium nitrate
(14 M N as
KNOs) in 2012
and concentrated
nitric acid (14 M
N as HNOs) in
2013
Although phytoplankton biomass
increased, net zooplankton
responses were modest and
attributed to incompatible
stoichiometry of food
(phytoplankton) to consumers
(zooplankton). Therefore,
unconsumed phytoplankton could
accumulate in unproductive boreal
lakes with increased N deposition.
Crustacean zooplankton: Boreal lakes
calanoid copepods
(Eudiaptomus spp.),
cyclopoid copepods
(Cyclops spp.), and
cladocerans (mainly
Ceriodaphnia spp.,
Daphnia spp., Bosmina
spp., Diaphanosoma
brachyurum, Holopedium
gibberum, and Sida
spp.).
Deininqer
et al.
(2017b)
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Table 9-3 (Continued): Summary of studies on nitrogen effects on species composition and biodiversity that have
been evaluated in the literature since the 2008 Integrated Science Assessment for Oxides
of Nitrogen and Sulfur-Ecological Criteria.
Region
Endpoint
Deposition
kg N/ha/yr
N Addition
Observation
Species
Ecosystem
Type
Reference
Denmark	Macroinverte- Not specified
brate
occurrence
None
Macroinvertebrate communities
did not change significantly with
TN based on an analysis of a
large number of concurrent
samples of macroinvertebrate
communities and chemical
indicators of eutrophication and
organic pollution (TP, TN,
NH4+-N, biological oxygen
demand [BOD5]) from 594 Danish
stream sites.
Plecopteran Leuctra spp. Streams
Isopod Asellus aquaticus
Dipteran Chironomus
spp.
Friberq et
al. (2010)
French Alps
Phytobenthos 7
species
richness,
grazing
pressure
Nutrient-diffusing A shift in taxonomic composition Diatoms
substrata
N treatment:
N = 113 [jg/N/hr
N + P:
N = 181 [jg/N/hr
P = 13 [jg/P/hr
Alpine lakes
of phytobenthos toward green
algae (less palatable to grazers)
from cyanobacteria and diatoms
was observed with N enrichment.
Benthic grazing by
macroinvertebrates was not
reduced.
Green algae
Cyanobacteria
Lepori and
Robin
(2014)
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Table 9-3 (Continued): Summary of studies on nitrogen effects on species composition and biodiversity that have
been evaluated in the literature since the 2008 Integrated Science Assessment for Oxides
of Nitrogen and Sulfur-Ecological Criteria.
Region
Endpoint
Deposition
kg N/ha/yr
N Addition
Observation
Species
Ecosystem
Type
Reference
Bavaria,
Germany
Phytoplankton
community
composition
and
abundance
0, 1, 2, 8, 16, and Increasing N enrichment gradient Dominant phytoplankton Oligotrophic Poxleitner
32 mL of solution
of 30 g/L NOs"
and 10 mg/L
NH4+
affected phytoplankton
stoichiometry and community
composition and heterotrophic
nanoflagellate and ciliate
abundances, indicating N load
alters basal food web.
were Bacillariophyceae,
dinoflagellates,
Chrysophyceae, and
Cryptophyceae. Only a
few individuals of a few
Chlorophyceae and
Cyanophyceae (mainly
Anabaena spp.) were
observed.
Bacillariophyceae's most
abundant species were
Asterionella formosa,
Cyclotella spp., and
Fragilaria crotonensis.
Another abundant
Chrysophyceae species
Dinobryon divergens and
the dinoflagellate
Ceratium hirundinella.
lake
etal. (2016)
DOM = dissolved organic matter; KH2P04 = monopotassium phosphate; N = nitrogen; NaN03 = sodium nitrate; TN = total nitrogen; TDP = total dissolved phosphorus; TP = total
phosphorus.
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Some freshwater algae are particularly sensitive to the effects of added N and experience
shifts in species diversity and community composition with increased N deposition. A
general shift from chrysophytes that dominate lakes with low N to cyanophytes and
chlorophytes in lakes with higher N has been observed in studies reported in the
2008 ISA (Lafrancois et al.. 2004; Wolfe et al.. 2003; Jassbvetal.. 1994) and newer
studies (Svmons et al.. 2012; Pardo et al.. 2011a; Saros et al.. 2010). For example, in
phytoplankton sampling from 21 lakes and ponds in Wapusk National Park, Canada,
Svmons et al. (2012) observed that the N limited lakes had significantly different
phytoplankton community composition, with more chrysophytes and Anabaena spp.
compared to all other lakes.
Using data from the U.S. Geological Survey National Water Quality Assessment, Passv
(2008) assessed the responses of 2,426 benthic and 383 planktonic diatom communities
from 760 and 127 distinct localities, respectively, to nutrient limitation. As more
resources (basic cations, silica, iron, NH3, NO3 . and dissolved P) became limiting,
benthic diatom richness declined while phytoplankton richness increased.
9.3.2.1. Paleolimnological Studies
In the 2008 ISA, changes in diatom species assemblages, increases in cell numbers, and
pigment-inferred increases in whole-lake primary production were reported from
paleolimnological studies of mountain lakes that have only been disturbed by
atmospheric deposition and climate change. Sediment cores in oligotrophic lakes
receiving <5 kg N/ha/yr showed diatom community shifts that in part could be also
related to climate change. As reported in the 2008 ISA and Appendix 9.2.1 of this
appendix, two opportunistic species of diatom, A. formosa and F. crotonensis, now
dominate the flora of at least several lakes in the Rocky Mountains (Slcmmons etal..
2015; Arnett et al.. 2012; Saros et al.. 2010; Saros et al.. 2005; Saros et al.. 2003; Wolfe
et al.. 2003; Wolfe et al.. 2001; Interlandi and Kilham. 1998). In the southern Rocky
Mountains, this shift occurred in the 1950s in the south, with more recent shifts (1970s)
in the central region (Baron et al.. 201 lb). In most, but not all, of these studies, the
observed responses in phytoplankton were concordant with effects from increased
atmospheric N deposition. Increased abundance of A. formosa is also linked to changes in
lake temperature under conditions of declining N deposition and decreasing lake total N
concentration (Sivaraiah et al.. 2016).
Studies available since the 2008 ISA (summarized here and in Table 9-1) provide
additional evidence from historical records of lake sediments in assessing biological
changes associated with N enrichment over time. In a survey of 28 lakes in national parks
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across the western Great Lake region, 63% of study lakes experienced changes in the
diatom community correlated with lake sediment S15N (Hobbs et al.. 2016). About 36%
of sediment S15N was statistically correlated to some form of deposited N. Sediment
records from lakes in the Uinta Mountains, UT showed shifts in diatom community
composition and increasing abundance of the nitrophilous diatom A. formosa linked to
atmospheric deposition (Hundcv et al.. 2014). In sediment core sampling from
high-elevation lakes from national parks in Washington, Sheiblev et al. (2014) analyzed a
total of 56 sediment samples for diatom presence and abundance overtime. In the survey,
over 250 diatom species were identified; however, only Hoh Lake in Olympic National
Park showed clear evidence of impacts from N deposition based on changes in sediment
diatom assemblages and the presence of A. formosa, F. crotonensis, and Fragilaria
tenera. In high-elevation lakes in the Rocky Mountains with low atmospheric deposition
(<2 kg N/ha/yr) and low to moderate surface water NOs concentrations (<1 |_ig/L [below
detection] to 30 (ig/L). Arnett et al. (2012) observed that diatom assemblages were
already dominated by nitrophilous species like A. formosa or F. crotonensis. Because of
the abundance of species indicative of moderate N enrichment even at NO3
concentrations below the detection limit, it was not possible to identify a threshold of
reactive N response for diatom community change. In Baptiste Lake in Alberta, Canada,
a general shift in diatom assemblages away from N intolerant species and proliferation of
Stephanodiscus hantzschii, a nitrophilous diatom species, was observed (Adams et al..
2014). Paleolimnological analysis suggests that eutrophic conditions were present in the
lake for at least 150 years.
N deposition, along with climate change, was identified as a driver of diatom
compositional turnover (or beta diversity) in a synthesis of paleolimnological core
samples of 52 Arctic, alpine, and boreal montane lakes in North America and western
Greenland. Hobbs et al. (2010) stated that in all lakes, beta diversity was significantly
greater during the 20th century than the 19th century, with only a small and
nonsignificant difference in turnover between the 19th century and the 1550-1800
intervals (p = 0.86). Compared with forested montane boreal sites, both alpine and Arctic
lakes reveal greater diatom assemblage turnover in the 20th century. In Arctic lakes,
temperature appears to be the main factor affecting diatom composition turnover, while
in temperate lakes, N deposition appears to play a larger role. A meta-analysis of
200 lakes in Europe and North America showed diatom shifts in Arctic lakes around
1850 and temperate lakes around 1970 (Ruhland et al.. 2008). Observed changes were
primarily attributed to warming trends. The authors examined the diatom record from one
lake in detail (Whitefish Lake in Ontario) and reported that observations of diatom shifts
in this lake were not explained by N deposition but rather corresponded to temperature
changes.
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New information on relative NOs inputs from glacial versus snowpack meltwaters
reported in Appendix 7 indicates water of glacial origin has higher levels of NO3, which
may influence the interpretation of biological data from high-elevation lakes and streams
in some regions of the U.S. (Slemmons et al.. 2015; Slemmons et al.. 2013; Saros et al..
2010; Baron et al.. 2009). In the central Rockies and Glacier National Park, fossil diatom
richness in snowpack-fed lakes was found to be higher (34 to 54 taxa) relative to lakes
fed by both glacial and snowpack meltwaters (12 to 26 taxa) (Saros et al.. 2010). In the
central Rockies, N deposition was 1.4 to 2.5 kg Nr/ha/yr, with 10 to 20% of the lakes
there receiving glacial meltwater. In Glacier National Park in the northern Rockies,
approximately 50% of lakes receive glacial meltwater and deposition was
2.0 to 3.4 kg Nr/ha/yr. In another study comparing sedimentary fossil diatom
assemblages in two adjacent lakes (one glacier-fed and the other snow-fed) in the central
Rocky Mountains, increased abundances of A. formosa, and F. crotonensis were
observed in the glacier-fed lake starting 1,000 years ago along with a decrease in diatom
species richness (Slemmons etal.. 2015). Shifts in the planktonic diatom communities
occurred after 1970 in snow-fed lakes (Slemmons et al.. 2017). These observations
suggest increased N inputs associated with glacial meltwater have altered the fossil algal
record and continue to affect algal communities in some glacier-fed lakes. No significant
differences in phytoplankton biomass and community structure were observed between
snowpack-fed and glacier and snowpack-fed lakes in the northern Cascade Mountains in
Washington in a comparison of data collected in 1989 and 2013 (Williams et al.. 2016b).
9.3.2.2. Bioassay, Mesocosm, and Laboratory Studies
Several experimental nutrient additions (mesocosm and bioassay studies) described in
Appendix 9.1.1.3 show that N limitation is common in freshwater systems (Elser et al..
2009b; Elser et al.. 2009a; Bergstrom et al.. 2008). Since the 2008 ISA, several N
addition studies have explored the effects of increased nutrients on phytoplankton
biomass and community structure (Daggett et al.. 2015; Gardner et al.. 2008). In the
Colorado Front Range, diatom abundance increased, and phytoplankton species
composition shifted in nutrient-enriched mesocosms (Gardner et al.. 2008). Principal
component analysis suggested that 21% of the variance in phytoplankton community
composition was related to added nutrients, while 34% of the variance was due to
seasonal changes. In a bioassay comparison of phytoplankton community structure
following N and DOM addition to a low DOC and N and P colimited water body (Jordan
Pond in Acadia National Park, ME) and an N limited lake (Sargent Lake in Isle Royale
National Park, MI), Daggett et al. (2015) observed increased chlorophytes following
DOM inputs in both lakes. In the N limited lake, an increase in the abundance of F.
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crotonensis and Tabellaria flocculosa was observed. There was a reduced response of
chlorophytes to DOM addition when N was added to the N limited lake.
Most N addition studies of phytoplankton communities have focused on N limited
systems. Poxleitner et al. (2016) conducted an N addition mesocosm experiment during
the spring season in a P limited oligotrophic lake in Upper Bavaria, Germany. An
increasing N enrichment gradient affected phytoplankton stoichiometry and community
composition. Only small effects of N enrichment were documented using the biovolume
of phytoplankton, the amount of particulate organic carbon, and the concentration of
chlorophyll a as indicators. There was an effect, however, on phytoplankton community
composition and heterotrophic nanoflagellate and ciliate abundances. Thus, changes in
food web dynamics were suggested for P limited lakes when N levels are increased.
9.3.2.3. Whole Lake Studies
Altered phytoplankton community composition and increased phytoplankton biomass
resulted frominorganic N fertilization of Reindeer Lake in northern Sweden (Jansson et
al.. 2001). In a whole-lake N fertilization study of three small boreal lakes in northern
Sweden, Deininger et al. (2017a) observed that changes in community composition of
phytoplankton were related to DOC rather than N addition. As DOC increased along a
gradient, community composition shifted from nonflagellated toward high DOC-adapted
flagellated autotrophs in the three fertilized lakes. In the same set of lakes, phytoplankton
biomass increased but net zooplankton responses were modest, attributed by the authors
to incompatible stoichiometry of food (phytoplankton) to consumers (zooplankton)
(Deininger et al.. 2017b).
9.3.3. Benthic Algal Diversity
Benthic algae are typically less sensitive than planktonic algae to nutrient inputs because
the former are associated with nutrient-rich substrates where sunlight reaches the
sediment surface (Spaulding et al.. 2015). Periphyton is typically more abundant than
phytoplankton in these shallow habitats (Vinebrooke et al.. 2014; Nvdick et al.. 2004).
Several recent studies have considered the effects of N on benthic algae characteristic of
shallow lakes and littoral zones where primary production is mainly controlled by the
availability of light (Vadeboncoeur et al.. 2008). In a stable isotope tracer study in Bull
Trout Lake, a subalpine lake in Idaho, the largest portion of added N was retained within
the periphyton (Epstein et al.. 2012). A directional change in benthic diatom species after
1960 that correlates with atmospheric deposition was observed in lake sediment cores
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from high-elevation shallow lakes in Grand Teton National Park (Spaulding et al.. 2015).
Liess et al. (2009) conducted a survey of periphyton along an N deposition gradient (2 to
12 kg N/ha/yr) in Sweden. Benthic algal composition in lakes in the northern part of the
country where N deposition was lower had a higher contribution from N fixing
cyanobacteria (5% of algal biovolume) than in the south. Overall, lakes were more N
limited in the North and P limited in the South. In another study using nutrient-diffusing
substrata in two small Alpine lakes in the Rhone Alps, France, N enriched substrata had a
greater phytobenthic biomass, and the taxonomic composition of the phytobenthos was
different between the treatment and control (Lcpori and Robin. 2014).
In pond fertilization experiments in 12 Ashless, nonglacial ponds in Snow Pass within the
Cascade Valley catchment of Banff National Park, Alberta, periphyton outcompeted
phytoplankton for limiting nutrients, indicating the importance of considering both
benthic and pelagic primary producers (Vincbrookc et al.. 2014). In these N limited
ponds, phytoplankton biomass increased significantly only when N was applied in the
absence of fairy shrimp. High grazing pressure by fairy shrimp appeared to suppress the
effects of added nutrients on algal communities, suggesting that trophic interactions are
important to consider to avoid missing the effects of N in alpine water bodies.
The algal assemblage response following 6 weeks of nutrient amendments in Wyoming
streams indicated that both N and P altered community structure of epilithic biofilm
(Kunza and Hall. 2013). Depending on which nutrient was added, N2 fixers reacted
differently than non-N2 fixers; an increase in non-N2 fixing diatoms in the N and N + P
additions was observed compared to N2 fixer dominance in control and P treatment.
Other studies consider how changes in benthic algal biodiversity affect ecosystem
processes such as chemical uptake. NO3 -N uptake differed among benthic algal
assemblages on transplanted rocks in a stream in Boise National Forest in central Idaho
(Baker et al.. 2009). Uptake of NO3 was highest in the green filamentous algae,
(dominated by the chlorophytes Spirogyra spp. and Rhizoclonium spp.), lowest in the
yellow patch type (dominated by the chlorophytes Spirogyra spp. and Bulbochaete spp.),
and intermediate in the brown patch type (dominated by diatoms, including Synedra spp.,
Cymbella spp., Fragilaria spp., and Epithemia spp.). NO3 —N uptake normalized to
chlorophyll a increased with algal composition and species richness in the three patch
types.
9.3.4. Zooplankton Diversity
Changes to aquatic food webs in response to N enrichment have not been as thoroughly
explored as changes to algal assemblages, but a few studies in the 2008 ISA showed
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declines in zooplankton biomass (Lafrancois et al.. 2004; Paul et al.. 1995) in response to
N related shifts in phytoplankton biomass toward less palatable taxa with higher C:P
ratios (Elser et al.. 2001). New studies provide additional evidence for N deposition
effects on zooplankton through altered trophic interactions.
The nutritional status of zooplankton is influenced by the quantity and quality of food
items (Elser et al.. 2001). In lakes across an N deposition gradient in Norway,
zooplankton feeding on seston (organisms and nonliving matter in the water column)
were found to be affected by P limitation in phytoplankton (Elser etal.. 2010). The seston
in lakes with high N deposition had significantly higher C:P and N:P ratios due to
phytoplankton P limitation. Using an assay to measure alkaline phosphatase activity
(APA), which is overexpressed in animals with dietary P deficiency, Elseretal. (2010)
observed that the biomass-specific APA value differed considerably among taxa and that
N deposition was significantly associated with some species such as seston-feeding
zooplankton but not others, leading the authors to conclude that P limitation was
transferred up the food chain. In three P deficient lakes in Germany, N enrichment
experiments showed declines in mesozooplankton density and biomass, especially
cladocerans, with N additions in the range of projected increasing atmospheric deposition
(Trommcr et al.. 2017). An increase in seston C:P ratios was observed in one of the three
P deficient lakes in response to N enrichment.
A taxonomic shift from omnivorous, raptorial-feeding copepods to filter-feeding
herbivorous daphnids was observed with warming and N addition in a growth chamber
experiment (Thompson et al.. 2008). The study was conducted with water and sediment
collected from Pipit Lake in the Canadian Rockies in Banff National Park, Alberta. While
warming and N fertilization increased phytoplankton abundance, herbivory by Daphnia
middendorffiana decreased but only in the presence of sediment (pond conditions).
Findings demonstrated that the effects of warming and N differ both among trophic levels
and aquatic alpine habitats.
9.3.5. Macroinvertebrate Diversity
Macroinvertebrates such as aquatic insects, crustaceans, worms, and mollusks are
commonly sampled to assess water quality in federal and state monitoring programs (U.S.
EPA. 2016i). These organisms feed on algae and organic material and are an important
food source for fish and other wildlife. Few studies have considered the effects of
atmospheric N enrichment on biodiversity of freshwater macroinvertebrates. Due to the
existence of a survey of benthic invertebrates in Lake Tahoe, NV conducted in the 1960s,
it was possible to assess how the populations have changed with enrichment and the
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introduction of non-native species. In this high alpine lake, atmospheric N contributions
are a significant portion (approximately 57%) of the total N loading to the lake (Sahoo et
al.. 2013). Caires et al. (2013) resurveyed the lake in 2008-2009 and observed declines of
80 to 100% in endemic benthic invertebrate taxa and changes to the community structure
of benthic invertebrate assemblages. Corresponding to these changes in lake biota, a
decrease in water clarity over the four decades between studies has been associated with a
shift in the bottom of the euphotic zone (1% light penetration) from 80 to 57 m (Chandra
et al.. 2005).
A recent study from Denmark suggests that macroinvertebrates may be poor indicators of
N enrichment. Friberg et al. (2010) measured stream macroinvertebrate occurrence along
gradients in organic pollution and eutrophication in 594 streams. They observed that the
occurrence of many taxa showed a stronger relationship to habitat condition than to
chemical variables. Overall, macroinvertebrate occurrence appeared to be related
primarily to biological oxygen demand, NH4+-N, and TP rather than TN. In a principal
component analysis using the United States Geological Survey (USGS) National
Ambient Water Quality Assessment (NAWQA) data set, Carlisle et al. (2007) identified
nutrients (NH/, NOs . TP) as a factor affecting macroinvertebrate occurrence, although
specific conductance, pH, and SO42 explained the greatest effects on macroinvertebrate
abundance. A 20% loss of macroinvertebrate taxa was indentified as a threshold for
degraded streams based on benthic macroinvertebrate sampling data from a subset of
basins in the eastern and midwestern U.S. (Carlisle and Meador. 2007).
9.3.6. Macrophyte Diversity
No U.S. studies of N effects on macrophyte (aquatic plant) community biodiversity in
atmospherically N enriched lakes and streams have been identified in the recent
literature, although declines in total macrophyte occurrence were noted in a resurvey of
Lake Tahoe that compared current samples with those from a lake survey conducted in
the 1960s (Caires et al.. 2013). Benthic invertebrate declines were most severe in
sampling sites where macrophytes were present in the 1960s but now absent, suggesting
that macroinvertebrate and macrophyte assemblages are closely associated in the lake.
Barker et al. (2008) conducted a mesocosm experiment using water and sediment from a
stream in Norfolk, U.K. Experimental tanks were planted with 11 macrophyte species
from the local environment. Constant P loadings were given to all tanks (50 |_ig P/L).
Nitrate loading varied from 1 to 10 mg NO3 -N/L. Macrophyte species richness
decreased with increasing N during the first year of treatment, and decreased in all
treatments above 1 mg NO3 -N/L during the second year. Barker et al. (2008) estimated
a threshold of 1.5 mg N/L for maintaining a stable species richness.
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9.3.7. Amphibian Diversity
No studies on N enrichment effects on amphibian diversity were reviewed in the
2008 ISA or identified in the current literature.
9.3.8. Fish Diversity
No studies of direct effects of N enrichment on freshwater fish diversity were reviewed in
the 2008 ISA. Post-2007 literature includes several behavioral endpoints in fish.
(Appendix 9.4.1).
9.4. Animal Behavior and Disease
In addition to changes in biological indicators (i.e., chlorophyll a, periphyton/microbial
biomass, diatoms, nutrient limitation shifts) and altered biodiversity, there is increasing
evidence for a role of N in behavior and disease in biota. These indirect responses may
impact the fitness of organisms inhabiting nutrient-enriched waters.
9.4.1. Behavior
Nutrient enrichment of freshwater systems has recently been shown to alter behavioral
endpoints in an invertebrate, an amphibian, and a fish species in laboratory exposures.
N03 exposure (21.4, 44.9, 81.8, 156.1 mg NO, /L) was shown to decrease the velocity
of movement in the aquatic snail Potamopyrgus antipodarum (Alonso and Camargo.
2013). Reproductive impairments (decreased number of newborns) were observed at all
tested concentrations. The NOs concentrations used in this study are much higher than
typically measured in remote freshwater catchments affected by atmospheric deposition.
In the presence of chemical cues of the predator nymphs of the dragonfly (Anax
imperator), western spadefoot toad (Pelobates cultripes) tadpoles typically decrease
swimming activity by 44% (Polo-Cavia et al.. 2016). When predator chemical cues were
added to water containing NH4NO3 tadpoles did not alter swimming activity. The
concentrations of NH4NO3 used in the study (20 mg/L and 80 mg/L) were not lethal to
the tadpoles, and no altered swimming activity was observed with the added N except in
the presence of predator chemical cues.
In the three-spined stickleback Gasterosteus aculeatus, a fish that inhabits both
freshwater and brackish habitats, changes to water quality associated with eutrophication
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(i.e., turbidity associated with algal blooms) have impacted social and reproductive
behaviors in laboratory studies. These studies are reviewed in Appendix 10.
9.4.2. Disease
The interactions of N enrichment and disease in biota is a relatively new area of research
in N impacted freshwater systems. The interactions characterized to date are complex and
involve indirect effects on host-parasite interactions (Johnson et al.. 2010; Johnson ct al..
2007). A host-parasite relationship potentially impacted by increased N to aquatic
systems is that of the fungal pathogen Metschnikowia bicuspidata, which is parasitic to
crustacean zooplankton Daphnia dentifera (Dallas and Drake. 2014). In a series of
bioassays designed to assess the effects of N on host and pathogen, D. dentifera were
exposed to six NOs concentrations (0.4, 2, 4, 8, 16, and 32 mg NOs /L) and then
inoculated withM bicuspidata. NOs, decreased D. dentifera population size and
increased infection prevalence. Next, ambient levels of N (0.4 mg NO3 -N/L) and a
concentration representative of a moderately polluted system (12 mg NO, -N/L) were
used to test the influence of NO3 on pathogen dose, infection prevalence, and host
fitness (growth and fecundity). No effects were observed on the growth rate of D.
dentifera; however, greater infection prevalence was associated with increased NO, . and
in general, both host fecundity and infection intensity decreased with increasing pathogen
dose.
9.5. Summary of Thresholds, Levels of Deposition at Which
Effects Are Manifested, and Critical Loads
Since the 2008 ISA, additional thresholds of response to N have been identified that are
useful for critical load development in water bodies impacted by N nutrient effects.
Critical loads for N nutrient enrichment in U.S. freshwater ecosystems are summarized in
Table 9-4. Factors contributing to uncertainty in N deposition estimates for assessment of
critical loads according to Pardo etal. (2011a) include "(1) the difficulty of quantifying
dry deposition of nitrogenous gases and particles to complex surfaces; (2) sparse data,
particularly for arid, highly heterogeneous terrain (e.g., mountains); and (3) sites with
high snowfall or high cloud water/fog deposition, where N deposition tends to be
underestimated." N deposition estimates at high-elevation sites such as those in the
Rocky and Sierra Nevada mountains are associated with considerable uncertainty,
especially uncertainty for estimates of dry deposition (Appendix 2). For sensitive
receptors such as phytoplankton shifts in high-elevation lakes, N deposition model bias
may be close to, or exceed, predicted critical load values (Williams et al.. 2017a).
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Physical, chemical, and ecological variability across lakes affect their response to N
deposition and contribute to uncertainty of critical load estimates (Appendix 9.1.1.2V A
review by Bowman et al. (2014) noted that current N critical loads for sensitive alpine
systems may not be protective under future climate scenarios of warmer summer
temperatures and a shorter duration of snow cover.
Available data from the 2008 ISA suggest that the increases in total N deposition do not
have to be large to elicit an ecological effect in remote alpine lakes. A hindcasting
exercise determined that the shift in Rocky Mountain National Park lake algae
composition that occurred between 1850 and 1964 was associated with wet N deposition
that was only about 1.5 kg N/ha/yr (Baron. 2006). Similar changes inferred from lake
sediment cores of the Beartooth Mountains of Wyoming occurred at about
1.5 kg TN/ha/yr deposition (Saros et al.. 2003).
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Table 9-4 Summary of critical loads for nitrogen eutrophication for surface
water in the U.S. [adapted from Pardo et al. (2011c) with newer
studies added].
Ecosystem
Site
Critical Load for
N Deposition
kg N/ha/yr
Comments
Reference/
HERO ID
Alpine stream
Southern Rockies/Loch
Vale Rocky Mtn. Nat.
Park
2
Modeled
Baron et al.
(1994)
Alpine lake
Northern
Rockies/Beartooth Mtns.,
WY
1.5
Paleolimnological, shifts in
diatom assemblages with
Fragilaria crotonensis and
Cyclotella bodanica
increasing to comprise
approximately 30% each
of total assemblage
Saros et al.
(2003)
Alpine lakes
Rocky Mountains
2.5
Surveys and references
therein
Berqstrom and
Jansson(2006)
Alpine lakes
Eastern Sierra Nevada
and Greater Yellowstone
Ecosystem
1.4
Paleolimnological,
increase in the relative
abundances of
Asterionella formosa and
Fragilaria crotonensis
Saros et al.
(2011)
High-elevation
lakes of western
and eastern
U.S.
Rocky Mountains, Sierra
Nevada, Northeast
1.0 to 3.0 (western
lakes)
3.5 to 6.0
(northeastern lakes)
Based on N deposition
calculated in three
different ways and on lake
NO3" concentrations
Baron et al.
(2011b)
Remote lakes

2.0 (western lakes)
8.0 (eastern lakes)
Based on value of N
deposition at which
significant NO3" leaching
begins to occur
Pardo et al.
(2011c)
Alpine lakes
Rocky Mountain National
Park
>1.5
Based on NO3" threshold
of 0.5 |jM (concentration
which elicits a growth
effect in the diatom
A. formosa)
Nanus et al.
(2012)
Alpine lake
Hoh Lake, Olympic
National Park, WA
1.0-1.2
Increased relative
abundances of
Asterionella formosa and
Fragilaria tenera
Diatom assemblage shift
Sheiblev et al.
(2014)
Alpine lakes
Greater Yellowstone area
<1.5 to >4.0a
Modeled; based on
threshold value of NO3"
(0.5 to 2.0 peq/L)
Nanus et al.
(2017)
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Table 9-4 (Continued): Summary of critical loads for nitrogen eutrophication for
surface water in the U.S. [adapted from Pardo et al. (2011c)
with newer studies added].
Ecosystem
Site
Critical Load for
N Deposition
kg N/ha/yr
Comments
Reference/
HERO ID
Western U.S.
lakes
Remote mountain lakes
across the western U.S.
4.1a
Based on phytoplankton
biomass nutrient limitation
shifts
Williams et al.
(2017b)
Western U.S.
lakes
Remote mountain lakes
across the western U.S.
2.0a
Modeled to reduce
occurrence of false
negatives to near zero
Williams et al.
(2017b)
N = nitrogen; N03 = nitrate.
atotal N
Since the release of the 2008 ISA, work has continued on identifying thresholds of
response to N deposition that can be used to calculate critical loads in sensitive
freshwater systems. Threshold values for phytoplankton biomass growth have been
identified for nine western mountain lakes in North Cascades, Mount Ranier, and
Olympic national parks (13 to 25 (.ig DIN/L) (Williams et al.. 2016a) and in lakes in the
Sierra Nevada [0.3 to 4 (.iniol N/L (5 to 56 (.ig N/L); (Heard and Sickman. 2016)1.
Williams et al. (2017b) identified nutrient limitation shift thresholds (21 to 53 |_ig N/L)
for remote western U.S. mountain lakes. A nutrient threshold for surface water NOs for
growth of the diatom A. formosa was developed for high-elevation lakes in the Rocky
Mountains where N deposition is prevalent (Nanus et al.. 2012). The NO3 threshold of
0.5 (imol/L (31 |ig/L) was then used to estimate areas in Rocky Mountain National Park
that exceed critical load values.
Additional critical loads for nutrient enrichment of freshwaters developed since the
2008 ISA include those in the western and northeastern U.S. Pardo etal. (2011c)
estimated a critical load of 2.0 kg N/ha/yr for western lakes and 8.0 kg N/ha/yr for eastern
lakes based on the value of N deposition at which significant NO3 leaching begins to
occur. Using data from 125 lakes in the Greater Yellowstone area, including sensitive
lakes in both Grand Teton and Yellowstone national parks, Nanus et al. (2017) estimated
aquatic critical load values and identified locations with predicted critical load
exceedances based on a threshold value of NO3 concentration in lake water selected as
indicative of biological harm (0.5 to 2.0 j^ieq/L in this study). Critical loads of TN
deposition ranged from <1.5 ± 1.0 kg N/ha/yr to >4.0 ± 1.0 kg N/ha/yr. Exceedance
estimates were as high as 48% of the Greater Yellowstone area study region.
Baron etal. (2011b) found that in the western high-elevation lakes, increased primary
productivity and changes to algal diversity can occur with only minimal inputs of N
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deposition. They estimated that the thresholds, or critical loads, for nutrient enrichment
are 1.0 to 3.0 kg N/ha/yr for the western mountains (Sierra Nevada and Rocky
mountains) (Table 9-5). For minimally disturbed lakes in the Northeast, a critical load of
3.5 to 6.0 kg N/ha/yr was estimated, but independent biological measures for nutrient
enrichment are lacking in this region. In another study from the eastern Sierra Nevada
and Greater Yellowstone Ecosystem, Saros etal. (2011) determined a critical load of
1.4 kg N/ha/yr wet N deposition by modeling wet deposition rates from the period in
which diatom shifts first occurred. The shifts were identified from sediment cores
between 1960 and 1965 in the eastern Sierra Nevada, and starting in 1980 in the Greater
Yellowstone Ecosystem. Using core samples, Sheiblev et al. (2014) identified diatom
changes corresponding to N inputs from 1969-1975 in Hoh Lake in Washington and
established a critical load of 1.0-1.2 ± 0.01 kg N/ha/yr for the lake.
Using phytoplankton biomass N to P limitation shifts as the basis for critical load
calculations, Williams et al. (2017b) determined an empirical critical load of
4.1 kg/TN/ha/yr for remote high-elevation lakes across the western U.S. The critical
loads were calculated as the total (wet + dry) N deposition rate at which point below the
critical load, N limitation is more likely than P limitation and above the critical load, P
limitation is more likely than N limitation. Modeled critical loads using DIN:TP and DIN
response categories yielded an average critical load of 3.8 kg/TN/ha/yr for the lakes.
Modeled critical loads ranged from 2.8 to 5.2 kg/TN/ha/yr and correctly predicted
exceedances in 69% of lakes using NOs -N. The authors conducted a performance
evaluation using the NO3 -N univariate model and identified a critical load of
2.0 kg/TN/ha/yr to reduce the likelihood of false negatives to near zero.
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Table 9-5 Summary of mean lake nitrate (NO3-) concentrations, inorganic
nitrogen deposition amounts, and nutrient enrichment inflection
points where lake NO3- concentrations reflect increased nitrogen
deposition, [from Baron et al. (2011b)1.
Region
Mean
Lake
NOs"
Mean
1997-2006
NADP N
Deposition
Nutrient
Enrichment
Inflection
Point
(NADP)
Mean 1997-2006
(PRISM + NADP)
Deposition
Nutrient
Enrichment
Inflection Point
(PRISM + NADP)
2002 Total
N
Deposition
Nutrient
Enrichment
Inflection
Point
(Total N)
Sierra
Nevada
(n = 30)
2.7
(1.10)
1.5 (0.22)
1.5
2.5 (0.34)
1.5
3.4 (0.47)
2.0
Rocky
Mountains
(n = 285)
3.7
(1.90)
1.2 (0.27)
1.0
3.0 (0.28)
2.0
2.0 (0.24)
3.0
Northeast
(n = 216)
14.4
(0.76)
4.7 (0.18)
3.5
5.2 (0.18)
3.5
3.5 (0.20)
6.0
N = nitrogen; NADP = National Atmospheric Deposition Program; N03 = nitrate; PRISM = Parameter-Elevation Regressions on
Independent Slopes Model; PRISM + NADP = concentrations from NADP with PRISM precipitation values.
Note: Inorganic N deposition was calculated three ways, as was described in Baron et al. (2011b). Coefficients of variation are in
parentheses.
Mean lake N03" (in micromoles per liter), inorganic nitrogen (N) deposition amounts (in kg N/ha/yr), and nutrient enrichment
inflection points where lake N03" concentrations increase in response to increasing N deposition.
Source: Aber et al. (2003). Sickman et al. (2002). U.S. Forest Service's Air Resource Management online database
(www.fs.fed.us/ARMdata).
9.6. Summary and Causal Determination
In the 2008 ISA, the body of evidence was sufficient to infer a causal relationship
between N deposition and the alteration of species richness, community composition, and
biodiversity in freshwater ecosystems. New evidence from 2008 to the present, from
paleolimnological surveys, fertilization experiments, gradient studies, phytoplankton
community responses, and indices of biodiversity continue to show effects of N loading
to sensitive freshwater systems. As reported in the 2008 ISA, freshwater systems in
which N has been observed to influence ecological processes either received extremely
high inputs [e.g., (Dumont et al.. 2005)1. or had very low initial N concentrations and
responded rapidly to the additional inputs (Bcrgstrom and Jansson. 2006; Baron et al..
2000). As summarized in the 2008 ISA and supported by data in newer studies, nutrient
enrichment effects on freshwater ecosystems from atmospheric deposition of N are most
likely to occur in lakes and streams that have low primary productivity and low nutrient
levels and that are located in the most undisturbed areas with no local pollution sources.
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Even small inputs of N in these water bodies can increase nutrient availability or alter the
balance of N and P, which can stimulate growth of phytoplankton. As reported in the
2008 ISA and further strengthened with new evidence in this review, there is consistent
and coherent evidence for species composition changes and reductions in diversity,
especially for primary producers from high-elevation lakes in the western U.S., in
response to increased availability of N. New information is consistent with the
conclusions of the 2008 ISA that the body of evidence is sufficient to infer a causal
relationship between N deposition and changes in biota, including altered growth
and productivity, species richness, community composition, and biodiversity due to
N enrichment in freshwater ecosystems.
New data have not appreciably changed the consistent and coherent evidence at the time
of the 2008 ISA that freshwater systems that are likely to be most impacted by nutrient
enrichment due to atmospheric deposition of N are remote, oligotrophic, high-elevation
water bodies with limited local nutrient sources and with low N retention capacity. N
inputs to these ecosystems are linked to changes in biota, especially increased algal
growth and shifts in algal communities. Freshwater systems sensitive to N nutrient
enrichment include those in the Snowy Range in Wyoming, the Sierra Nevada
Mountains, and the Colorado Front Range. A portion of these lakes and streams where
effects are observed are in Class I wilderness areas (Williams et al.. 2017b; Clow et al..
2015; Nanus et al.. 2012).
Recent research further supports the 2008 ISA findings that N limitation is common in
oligotrophic waters in the western U.S. (Elser et al.. 2009b; Elser et al.. 2009a). Shifts in
nutrient limitation, from N limitation, to between N and P limitation, or to P limitation,
were reported in some alpine lake studies reviewed in the 2008 ISA and in this review.
Since the 2008 ISA, several meta-analyses have reported an increase in P deposition to
water bodies (Stoddard et al.. 2016; Brahnev et al.. 2015; Tipping et al.. 2014) and
highlight the need to account for how sustained P deposition can modify the effects of
anthropogenically emitted N deposition on productivity. P addition delayed the shift to P
limitation (prolonged N limitation) for phytoplankton.
The body of evidence for biological effects of N enrichment and altered N:P ratios in
remote freshwater systems (where atmospheric deposition is the predominant source of
N) is greatest for phytoplankton, the base of the freshwater food web. Lake surveys,
fertilization experiments, and nutrient bioassays reported in the 2008 ISA and in this
review show increased pelagic and benthic algal productivity (as indicated by
chlorophyll a concentration) to be strongly related to increased N concentration. An
increase in lake phytoplankton biomass with increasing N deposition was reported in
several regions, including the Snowy Range in Wyoming and across Europe. New studies
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in the Rocky Mountains of Colorado where atmospheric deposition ranged from 2 to
7 kg N/ha/yr support observations from the 2008 ISA which showed correlations between
greater chlorophyll a response and higher rates of deposition (Elser et al.. 2009a'). Several
recent studies have considered the effects of N on benthic algae characteristic of shallow
lakes and littoral zones where availability of light is the main factor controlling primary
production. A few studies have shown that periphyton outcompeted phytoplankton for
limiting nutrients, indicating the importance of considering both benthic and suspended
primary producers.
Changes in phytoplankton species diversity and community structure reported in the
2008 ISA (Lafrancois et al.. 2004; Wolfe et al.. 2003; Jassbvetal.. 1994) and newer
studies (Svmons et al.. 2012; Saros et al.. 2010) show a general shift from chrysophytes
that dominate lakes with low N to cyanophytes and chlorophytes in lakes with higher N.
In the 2008 ISA, diatom assemblage shifts were reported from the literature at total N
deposition as low as 1.5 kg N/ha/yr (Saros et al.. 2003). This evidence was further
supported by a hindcasting exercise that determined the change in Rocky Mountain
National Park lake algae that occurred between 1850 and 1964 was associated with an
increase in wet N deposition that was only about 1.5 kg N/ha/yr (Baron. 2006). Two
nitrophilous species of diatoms, A. formosa and F. crotonensis, are dominant in lakes
with higher N and serve as biological indicators ofN enrichment; however, increased
relative abundance of A. formosa has also been attributed to lake warming in some
regions where N deposition is decreasing (Sivaraiah et al.. 2016). Some shifts in algal
biodiversity observed in high-elevation waters are attributed to climate change or nutrient
effects and climate as costressors (Appendix 13).
The role of N in freshwater HAB formation has been further researched since the
2008 ISA. Additional evidence continues to show that the availability and form ofN
influences algal bloom composition and toxicity and that inputs of inorganic N
selectively favor some HAB species, including those that produce microcystin.
Microcystin is prevalent in U.S. waters as reported in recent regional and national
surveys. Although the risk of HAB formation is low in high-elevation oligotrophic water
bodies where N deposition is the dominant source ofN, transport of atmospheric inputs
can exacerbate eutrophic conditions in downstream water bodies. Increased
understanding of the role of N as a limiting nutrient in many freshwater systems has led
to recommendations to consider both N and P in nutrient reduction strategies (Dodds and
Smith. 2016; Gobler et al.. 2016; Paerl et al.. 2016b; Lewis etal.. 2011; Scott and
McCarthy. 2010; Conlev et al.. 2009; Paerl. 2009; Lewis and Wurtsbaugh. 2008).
Since the 2008 ISA, empirical and modeled critical loads for the U.S. have been
estimated based on surface water NOs concentration, diatom community shifts, and
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phytoplankton biomass growth nutrient limitation shifts. A critical load ranging from 3.5
to 6.0 kg N/ha/yr was identified for high-elevation lakes in the eastern U.S. based on the
nutrient enrichment inflection point [where NOs concentrations increase in response to
increasing N deposition; (Baron et al.. 201 lb)I. Another critical load of 8.0 kg N/ha/yr for
eastern lakes based on the value of N deposition at which significant increases in surface
water NO3 concentrations occur was estimated by Pardo etal. (2011c). In both Grand
Teton and Yellowstone national parks, critical loads for total N deposition ranged from
<1.5 ± 1.0 kg N/ha/yr to >4.0 ± 1.0 kg N/ha/yr (Nanus et al.. 2017). Exceedance
estimates were as high as 48% of the Greater Yellowstone area study region, depending
on the threshold value of NO;, concentration in lake water selected as indicative of
biological harm.
Additional critical loads have been identified since the 2008 ISA for eastern Sierra
Nevada lakes, Rocky Mountain lakes, the Greater Yellowstone Ecosystem, and Hoh
Lake, Olympic National Park ITable 9-4; (Nanus et al.. 2017; Sheiblev et al.. 2014; Pardo
et al.. 2011c; Saros et al.. 2011)1. The identified values fall near or within the range of 1.0
to 3.0 kg N/ha/yr for western lakes (Baron et al.. 201 lb). An empirical critical load of
4.1 kg/TN/ha/yr above which phytoplankton biomass P limitation is more likely than N
limitation was identified by Williams et al. (2017b) for the western U.S. Modeled critical
loads ranged from 2.8 to 5.2 kg/TN/ha/yr, and a performance analysis indicated that a
critical load of 2.0 kg/TN/ha/yr would likely reduce the occurrence of false negatives to
near zero.
The evidence for N effects on other freshwater biota is not as extensive as for
phytoplankton. Since the 2008 ISA, new research on archaea and bacterial diversity in
freshwater systems suggests that these organisms respond to lake trophic state and the
form ofN present. At higher trophic levels, zooplankton responses to N inputs are
attributed to changes in food quality which can potentially alter food web interactions
(Deininger et al.. 2017b; Meunier et al.. 2016; Elser et al.. 2009a). A few studies in the
2008 ISA showed declines in zooplankton biomass (Lafrancois et al.. 2004; Paul et al..
1995) in response to N related shifts in phytoplankton biomass toward less palatable taxa
with higher C:P ratios (Elser et al.. 2001). Limited studies since the 2008 ISA suggest
that atmospheric N inputs are linked to taxonomic shifts and declines in some
invertebrates, although the effects attributed to N are difficult to separate from other
stressors such as climate change and invasive species. Water quality changes associated
with excess N nutrient inputs, such as lowered DO and algal blooms that alter habitat by
covering up substrate, can also lead to declines in biodiversity in macroinvertebrates and
fish including species listed under the federal Endangered Species Act (Hernandez et al..
2016). Although little research has been conducted in freshwater systems, some evidence
suggests that increased turbidity associated with algal blooms may affect reproductive
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and social behaviors in fish (Appendix 10). Emerging research on disease in biota suggest
that N enrichment may modify relationships such as host susceptibility to parasites and
pathogens.
Since the 2008 ISA, further studies have shown that both trophic interactions and DOC
modify ecosystem response to N loading. In a whole-lake N fertilization study, Deininger
et al. (2017a) observed that changes in community composition of phytoplankton were
related to DOC rather than N addition to small N limited boreal lakes. As DOC increased
along a gradient, community composition shifted from nonflagellated toward high
DOC-adapted flagellated autotrophs in the three fertilized lakes. In the same set of lakes,
although phytoplankton biomass increased, net zooplankton responses were modest and
attributed by the authors to incompatible stoichiometry of food (phytoplankton) to
consumers (zooplankton) (Deininger et al.. 2017b). A study in Banff National Park,
Canada indicated that grazing pressure on algae may have negated the effects of nutrient
inputs on primary producers (Vinebrooke et al.. 2014). Limited evidence suggests a
decreased rate of leaf litter decomposition by microbial communities with N loading to
streams.
New studies show that the N contribution from glacial meltwater (which has higher NOs
relative to water from melting snow) affects diatom community composition in
high-elevation lakes and streams in some regions of the U.S. (Slemmons et al.. 2015;
Slemmons et al.. 2013; Saros et al.. 2010). N deposition to snow and glaciers are
important sources of N to alpine lakes and streams that are fed by meltwaters. In a
comparison of biological response in snowpack-fed lakes versus lakes with both glacial
and snowpack meltwater in the Rockies, fossil diatom richness was higher in the
snowpack-fed lakes (34 to 54 taxa) compared to lakes with both glacial and snow
meltwater inputs [12 to 26 taxa; (Saros et al.. 2010)1. Characterizing diatom community
responses to meltwater sources is important for interpreting the biological effects of N
deposition in high-elevation systems with both glacial and snow inputs.
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APPENDIX 10. BIOLOGICAL EFFECTS OF
NITROGEN ENRICHMENT IN
ESTUARIES AND NEAR-COASTAL
SYSTEMS
This appendix characterizes the biological effects of nitrogen (N) enrichment in estuaries
(areas where freshwater from rivers meets the salt water of oceans), coastal lagoons, coral
reef ecosystems, and open ocean areas near coastlines. An overview of N inputs to
coastal systems, including characteristics and identification of areas of the U.S. sensitive
to nutrient over-enrichment (Appendix 10.1) is followed by discussions of the indicators
of nutrient enrichment (Appendix 10.2). its effects on biodiversity and ecosystem
structure and function (Appendix 10.3). animal behavior, and disease (Appendix 10.4).
and the role of N in nutrient enhanced coastal acidification (Appendix 10.5).
Appendix 10.6 summarizes the thresholds of biological effects ofN in coastal regions,
and Appendix 10.7 reviews the causal determinations based on a synthesis of new
information and previous evidence from prior N assessments.
10.1. Introduction
In the 2008 Integrated Science Assessment for Oxides of Nitrogen and Sulfur-Ecological
Criteria (2008 ISA), the body of evidence was sufficient to infer a causal relationship
between N deposition and the alteration of species richness, species composition, and
biodiversity in estuarine systems (U.S. EPA. 2008a). Nitrogen pollution is the major
cause of harm to the majority of estuaries in the U.S. (Bricker et al.. 2008; NRC. 2000).
In estuaries, N from atmospheric and nonatmospheric sources contributes to increased
primary productivity, leading to eutrophication (Figure 10-1). Estuary eutrophication is a
process of increasing nutrient over-enrichment indicated by water quality deterioration,
resulting in numerous adverse effects including areas of low dissolved oxygen (DO)
concentration (hypoxic zones), species mortality, and harmful algal blooms (HABs).
Eutrophic systems are characterized by an increase in the rate of supply of organic matter
(primary production and organic carbon accumulation) in excess of what an ecosystem is
normally adapted to processing (Diaz et al.. 2013; Nixon. 1995). In the 2008 ISA the
strongest evidence for a causal relationship was from changes in biological indicators of
nutrient enrichment including chlorophyll a, macroalgal "seaweed" abundance, HABs,
DO, and submerged aquatic vegetation (SAV). Biological effects of increasing nutrient
enrichment also include changes in biodiversity in these systems (Chapter 1.2.2.4). For
this ISA, new information is consistent with the 2008 ISA and the causal determination
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has been updated to reflect more specific categories of effects. The body of evidence is
sufficient to infer a causal relationship between N deposition and changes in biota,
including altered growth, total primary production, total algal community biomass,
species richness, community composition, and biodiversity due to N enrichment in
estuarine environments.
Since the 2008 ISA, N loading has been recognized as a possible contributing factor to
acidification of coastal waters (Appendix 10.5). Dissolution of atmospheric
anthropogenic CO2 into the ocean has led to long-term decreases in pH. Increased
production of CO2 from microbial degradation of organic matter associated with
eutrophication and respiration of algae and seagrasses during the night along with
atmospheric anthropogenic CO2 inputs can make the ocean water more acidic (Wilson et
al.. 2016; Wallace et al.. 2014; Wetz and Yoskowitz. 2013; Sunda and Cai. 2012; Cai et
al.. 2011c; Howarth et al.. 201IV Organisms that produce calcium carbonate shells such
as calcareous plankton, oysters, clams, sea urchins, and coral may be affected by
long-term decreases in pH (Mostofa et al.. 2016; Pfister et al.. 2014; Kroeker et al.. 2013;
Sunda and Cai. 2012). The body of evidence is suggestive, but not sufficient to infer, a
causal relationship between N deposition and changes in biota, including altered
physiology, species richness, community composition, and biodiversity due to
nutrient-enhanced coastal acidification.
In the previous ISA, there was a strong scientific consensus that N is the principal cause
of coastal eutrophication in the U.S. (U.S. EPA. 2008a; NRC. 2000). On average, human
activity has likely contributed to a sixfold increase in the N flux to the coastal waters of
the U.S. over the past several decades, and excessive N represents the most significant
coastal pollution problem (Bricker et al.. 2008; Bricker et al.. 2007; Howarth and Marino.
2006; Bricker etal.. 2003; Howarth et al.. 2002; NRC. 2000). Many coastal areas receive
high enough levels of N input from human activities to cause eutrophication (Bricker et
al.. 2007; Howarth et al.. 1996a; Vitousek and Howarth. 1991). As reported in the 2008
ISA, N driven eutrophication of shallow estuaries has increased over the past several
decades, and environmental degradation of coastal ecosystems is now a widespread
occurrence (Bricker et al.. 2007; Paerl et al.. 2000).
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Atmospheric Deposition
iJvAV
~ - - •-«
% ** j Phytoplankton Bloom
• thrives on nutrients
-o
Dissolved Oxygen
from wave action
and photosynthesis
1 1"
I
4
~
~ 4 "
Dead ~
materials
#
settles m
Dissolved Oxygen
trapped in the upper,
lower-salinity layer
Decomposition
*
. Dissolved Oxygen used up
1 by microorganism respiration
1 i»				.
* * i	Nutrients
'» released by bottom sediments
Lower-density
surface water
Higher-density
bottom water
Dissolved Oxygen consumed
Ir~
Fish will avoid
hypoxia if possible
Shellfish
and other
benthic
organisms
unable
to escape
hypoxia
Decomposition of organic
matter in sediments
N = nitrogen.
Notes: atmospheric N deposition is one of the sources of nutrient enrichment to coastal areas.
Source: modified from U.S. EPA (2012b).
Figure 10-1 Eutrophication can occur when the availability of nutrients
increases above normal levels.
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Coastal systems are linked to terrestrial N processes along the freshwater to ocean
continuum as nutrients deposited to the watershed move downstream. As described in the
2008 ISA, in coastal ecosystems, the nutrients most commonly associated with
phytoplankton growth are N, phosphorus (P), and silicon (Si). The relative proportions of
these nutrients are important determinants of primary production, food web structure, and
energy flow through the ecosystem (Turner et al.. 1998; Justic et al.. 1995b; Justic et al..
1995a; Dortch and Whitledge. 1992). In general, estuaries tend to be N limited [(Paerl
and Piehler. 2008; Elser et al.. 2007; Howarth and Marino. 2006; NRC. 2000; Nixon.
1995; Howarth. 1988; D'Elia et al.. 1986) Appendix 10.1.31; however, some estuaries are
P limited, or colimited by N and P, or switch seasonally between N and P (Howarth et al..
2011; Paerl and Piehler. 2008; Howarth and Marino. 2006).
Altered biogeochemical processes associated with N loading (Appendix 7) may affect
aquatic biota in a diversity of coastal habitats. Habitats associated with coastal areas
include shallow open waters, sandy beaches, mud and sand flats, rocky shores, oyster
beds, coral reefs, mangrove forests, river deltas, tidal pools, and sea grasses (U.S. EPA.
2016g). Appendix 11 will cover wetland ecosystems, including those located on coasts in
which soils and/or sediments are periodically inundated by tides or flooding. Estuaries in
the U.S. are located on both coasts with varying levels of eutrophication occurring
lYBricker et al.. 2007):Figurc 10-21. SAV, including the eelgrass Zostera marina, are
important ecological communities found within some coastal bays and estuaries that are
sensitive to elevated nutrient loading (Appendix 10.2.5). Estuaries provide breeding
grounds, nurseries, and shelter for aquatic biota. Near-coastal coral reefs have a more
limited distribution in the U.S. occurring off south Florida, Texas, Hawaii, and U.S.
territories in the Caribbean and Pacific. Elevated N loading appears to play a role in
susceptibility of coral species to disease and bleaching (Appendix 10.4.2).
In the 2008 ISA and the National Estuarine Eutrophication Assessment (NEEA), a
comprehensive survey of eutrophic conditions in the Nation's estuaries conducted by the
National Oceanic and Atmospheric Administration (NOAA), the most eutrophic estuaries
in the U.S. occurred in the mid-Atlantic region, and the estuaries with the lowest degree
of eutrophication were in the North Atlantic lYBricker et al.. 2007) and Figure 10-21. In
the NEEA, 65% of assessed estuaries had moderate to high overall eutrophic conditions
(Bricker et al.. 2007). In the U.S., Chesapeake Bay is perhaps the best-documented case
study of the effects of human activities on estuarine eutrophication. Human disturbances,
such as landscape changes, have exacerbated the negative impacts of N deposition by
reducing N removal and retention in the upper watershed region. Anthropogenic N inputs
have substantially altered the trophic condition of Chesapeake Bay over the last 50 to
100 years. Signs of eutrophication in the bay include high algal production, low
biodiversity, and large hypoxic and anoxic zones. Other estuaries identified in the 2008
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1	ISA where the extent of hypoxia and algal blooms have increased included Long Island
2	Sound, the Pamlico Estuary in North Carolina, and along the continental shelf adjacent to
3	the Mississippi and the Atchafalaya River discharges to the Gulf of Mexico (U.S. EPA.
4	2008a). This appendix updates the state of the science on N-nutrient enrichment, focusing
5	on U.S. waters, since the release of the 2008 ISA.

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the atmosphere. In estuaries adjacent to areas of coastal upwelling, such as some
locations along the Pacific coast of the U.S., oceanic inputs of nutrients may also
represent an important source of N (Brown and Ozretich. 2009). N inputs can be
attributed to point sources from a single outfall or discharge point and nonpoint sources,
which are diffuse (U.S. EPA. 2008a). Both point and nonpoint sources have been
identified as targets for control of N inputs in coastal systems (Stephenson et al.. 2010;
Paerl et al.. 2002).
10.1.2. Trends in Atmospheric Deposition of Nitrogen (N)
Trends in atmospheric deposition are discussed in Appendix 2. In summary, in many
parts of the U.S., including the Southeast, Mid-Atlantic, and Midwest, deposition of
reduced N has increased relative to oxidized N in last few decades, shifting from a nitrate
(NO, ) dominated to an ammonium (NH4) dominated condition. This trend is expected
to continue in the future under existing emission controls and current projections of
atmospheric deposition (Li et al.. 2016d; Ellis et al.. 2013; Pinder et al.. 2008; U.S. EPA.
2008a). Wet deposition is now primarily NH4+ at nearly 70% of U.S. air monitoring
locations and reduced N dominates dry deposition in most parts of the country (Li et al..
2016d). This relative increase in NH4 is attributed to intensified industrial-scale animal
operations, increased application of fertilizer, and successful NOx emission controls (Li
et al.. 2016d; Xing et al.. 2013). Mobile source emissions, especially emissions from
diesel vehicles, which increasingly use urea to control NOx emissions, also contribute to
total reduced N loading as well as near-road runoff in regions receiving heavy traffic
(Bettez et al.. 2013). The increase in highly bioreactive reduced N from deposition and
other sources is often a preferred form of N for phytoplankton, including harmful species
(Appendix 10.3.3). In coastal areas of the U.S., atmospheric inputs are heterogeneous
ranging from <10 to approximately 70% of the N inputs (Table 7-8).
10.1.3. Nitrogen Limitation
As reported for several decades prior to the 2008 ISA, N enrichment of marine and
estuarine waters can alter the ratios among nutrients such as P and Si and affect overall
nutrient limitation. N is the most common limiting nutrient in estuarine and coastal
waters and continued inputs have resulted in N over enrichment and subsequent
alterations to the nutrient balance in these systems (U.S. EPA. 2008a; Elser et al.. 2007;
Howarth and Marino. 2006; NRC. 2000; Paerl et al.. 2000; Nixon. 1995; Howarth. 1988;
D'Eliaetal.. 1986). As described in the 2008 ISA, nutrient limitation may shift along the
estuarine to ocean continuum (U.S. EPA. 2008a). For example, Chesapeake Bay exhibits
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a shift in nutrient limitation with P most commonly limited at upstream freshwater
locations. At the transition between fresh and salt water, N and P may be colimiting,
whereas the saltwater environments in the lower bay and sound regions are usually N
limited (Fisher et al.. 1998; Rudek et al.. 1991). Levels of N limitations are affected by
seasonal patterns in estuaries, with N limited conditions likely occurring during the peak
of annual productivity in the summer (U.S. EPA. 2008a').
Since the 2008 ISA, N limitation in estuarine and marine systems has been further
evaluated, with recognition of the shifting nature of nutrient limitation based on relative
nutrient inputs and other ecosystem conditions. The rate of nutrient delivery, especially
N, to coastal waters is strongly correlated to primary production and phytoplankton
biomass (Paerl and Piehler. 2008). Strong input controls causing low riverine P inputs
have, in some cases, exacerbated N limitation downstream, with implications for
estuarine eutrophication (Paerl. 2009). Likewise, if N inputs are high enough, some
estuaries and coastal marine ecosystems can become P or Si limited (Howarth et al..
2011; Paerl and Justic. 2011; Paerl and Piehler. 2008). Under this scenario, if a system
becomes P limited, the N input may travel farther away from its sources and contribute to
eutrophication at greater downstream distances (Howarth et al.. 2011).
The role of N inputs from upstream and the connectivity between freshwater and
receiving estuarine and coastal waters have led to recommendations to reduce both N and
P in upstream waters (Glibert and Burford. 2017; Paerl et al.. 2016b; Woodland et al..
2015; Paerl et al.. 2014; Conlev et al.. 2009; Paerl. 2009). Increased N inputs may be
affecting N limitation in the open ocean as well as in near-coastal areas. (Kim et al..
2014a) reported a detectable increase in NOs concentration in the upper Pacific Ocean.
The rate of increased N relative to P was highest near the source of anthropogenic
emissions in northeastern Asia with rates decreasing eastward across the upper North
Pacific Ocean. The authors suggest that increased N deposition may enhance primary
production and potentially lead to a shift from N to P limitation in this region.
Additional studies support observations reported in the 2008 ISA of N and P dynamics in
estuaries. P limitation can play an important role, particularly seasonally if the water
body has a significant freshwater source (Kemp et al.. 2005; Malone et al.. 1996). P
limited conditions often exist in the low salinity regions while N limitation is more
common downstream in higher salinity waters (Paerl and Otten. 2013). Studies in the
Chesapeake Bay and in Europe have shown that the phytoplankton community in coastal
ecosystems can be P limited over several months in the spring and switch to N limitation
for periods of time as short as 1 week later in the season (Trommer et al.. 2013; Malone
et al.. 1996). In the low-nutrient waters of North Carolina's Alligator River estuary,
phytoplankton primary productivity and biomass (chlorophyll a) increased with N
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additions, but both indicators showed the greatest increases after treatment with
combined N and P, indicating that this system was colimited by N and P (Rossignol et al..
2011). N additions in this region are known to lead to poor water quality and related
complications, such as algal blooms, hypoxia/anoxia (water with DO that is too low to
support marine biota), fish kills, and to impact ecosystem services, including fisheries
and recreation (Pacrl and Piehler. 2008).
10.1.4. Characteristics of Coastal Systems Sensitive to Eutrophication
Each coastal system has site-specific characteristics that influence ecological response to
nutrient loading. A variety of factors that govern the sensitivity of estuaries and
near-coastal marine waters to eutrophication from atmospheric N deposition are
summarized in the 2008 ISA. Of critical importance is the total N input from all sources,
including both atmospheric and nonatmospheric sources (Appendix 10.1.1). Other key
elements include the flushing rate and dilution capacity of the watershed which reflects
the volume of water available to dilute added N (NRC. 2000; Brickeretal.. 1999). The
NEEA defined susceptibility as an estimate of the natural tendency of an estuary to retain
or flush nutrients (Bricker et al.. 2007). In estuaries that have longer residence times,
nutrients are more likely to be taken up by algae and lead to eutrophic conditions.
As described in the 2008 ISA, the principal watershed features that control the amount of
increased N flux to estuaries in the U.S. include human population, agricultural
production, and the size of the estuary relative to its drainage basin (Fisher et al.. 2006;
Caddy. 1993; Peierls et al.. 1991). A study included in the 2008 ISA reported a strong
correlation between population density (persons/km2) and the total N loading from
watershed to estuary (r2 = 0.78) for coastal watersheds in the U.S. (Turner et al.. 2001).
This finding is likely due to the prevalence of automobiles in heavily populated areas,
along with their associated N emissions and deposition, plus the myriad of
nonatmospheric sources of N from human activities, particularly wastewater discharges.
The study authors also determined that direct atmospheric deposition becomes
increasingly more important as a contributor to the total N loading to an estuary as the
water surface area increases relative to total watershed area (terrestrial plus water
surfaces). Because of human population growth and the preference for living in coastal
areas, there is substantial potential for increased N loading to coastal ecosystems from
both atmospheric and nonatmospheric sources.
Freshwater inputs to coastal areas and subsequent response to nutrient inputs depend on
residence time of the nutrient-laden freshwater and the degree of tidal exchange within
the estuary (Zaldivar et al.. 2008). Residence time is defined by physical and hydrological
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characteristics of the watershed, such as surface area, volume, depth, and flushing rate
(Pacrl et al.. 2002). In water bodies with short residence times, there is little opportunity
for nutrients to be taken up and for algal blooms to develop (Bricker et al.. 2007). For
example, in the heavily N loaded lower Hudson River estuary, phytoplankton are flushed
away as fast as they can grow due to high input of freshwater and high rates of flushing
(Howarth and Marino. 2006; Howarth et al.. 2000). In the NEEA, systems with longer
flushing times were considered more susceptible to eutrophication (Brickcr et al.. 2007).
Other factors within the highly variable estuarine environment (Appendix 7.2.2) that
influence the composition of biological communities include salinity, DO, and suspended
solids, which vary spatially and temporally along the estuary continuum (Borja et al..
2012). Mixing depth, temperature, and light penetration depth also affect biological
response (Pacrl et al.. 2002). A stratified water column that separates well-oxygenated
surface water from bottom water and sediments is generally required to form hypoxic
zones (Jcwctt et al.. 2010). In contrast to the seasonal or persistent nature of hypoxic
zones, diel-cycling hypoxia typically occurs over hours to days and does not require
stratification. Precipitation events like storms and floods or drought can also modulate
nutrient effects I (Pacrl and Piehler. 2008); Appendix 131.
In the 2008 ISA and NEEA, estuaries characterized as eutrophic were generally those that
had large watershed-to-estuarine surface area, high human population density, high
rainfall and runoff, low dilution, and low flushing rates (U.S. EPA. 2008a; Brickcr et al..
2007). In literature reviewed for this ISA, many studies have re-emphasized the important
role that physical and hydrologic factors play in determining which estuaries and coastal
ecosystems are the most sensitive to N enrichment (Hart et al.. 2015; Wilkerson et al..
2015; Glibert et al.. 2014; Rothenberger et al.. 2014; Howarth et al.. 2011; Kennison et
al.. 2011). The hydrodynamics of a system may play an overriding role in controlling
phytoplankton growth (Hart et al.. 2015; Yang et al.. 2008). These factors, which affect
estuarine response to nutrient loading, should be taken into account when establishing N
input thresholds so that eutrophication can be controlled for different ecosystem types,
hydrologic conditions, and future climate scenarios (Paerl et al.. 2014).
Bricker et al. (2014) evaluated N inputs to the Potomac River Estuary and found that
despite some improvements in the upper estuary (i.e., increased DO and decreased
chlorophyll a in the tidal fresh zone; continued regrowth of sea grasses), eutrophic
conditions have worsened in the lower estuary since the early 1990s. Eutrophic
conditions in the Potomac were found to be representative of the Chesapeake Bay region
and other U.S. estuaries, with moderate to high levels of nutrient-related degradation,
particularly compared with similar river-dominated low-flow systems (Bricker et al..
2014). These river-dominated watersheds with low flow (>10 days residence time), high
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population density (i.e., >100 people/km2), and >40% of land use classified as urban
and/or agriculture were predicted to be likely impacted by eutrophication.
(Scavia and Liu. 2009) evaluated a nutrient-driven phytoplankton model that provides a
first-order screening tool for estuarine susceptibility classification. Using data from
75 estuaries, they found that the susceptibility of an estuary to nutrient loading could be
estimated based on the ratio of river inflow (Q) to estuarine volume (V). In this analysis,
efficiency appeared to decrease roughly with the inverse square root of OIV\
s = 0.908(6>/F) "47(/^2 = 0.53), where s represents mean values arising from the
75 estimated normal distributions. Model results showed that estuaries with a 0: F value
greater than 2.0/year are less susceptible to nutrient loads, and those with Q: V values
between 0.3 and 2.0/year are moderately susceptible. Case studies showed that Q:V—and
thus estuarine sensitivity to nutrient loading—can vary between seasons and with storm
events due in part to fluctuations in river inflow.
In some estuaries, especially in the Pacific Northwest, upwelling and oceanic exchange
caused by regional wind patterns likely control primary production rather than
anthropogenic nutrient loading (Brown and Ozretich. 2009; Hickev and Banas. 2003).
Nutrient inputs from local and regional upwelling in these systems can be difficult to
discern from anthropogenic sources. Upwelling-dominated areas are characterized by
short water residence time and have a moderately low expression of eutrophication
symptoms, although nutrient concentrations are high (Kaldv et al.. 2017; Brown and
Ozretich. 2009; Bricker et al.. 2008; Bricker et al.. 2007). Transfer of hypoxic water from
upwelling to estuaries can also occur but is not linked to anthropogenic nutrient additions
(Brown and Power. 2011).
10.1.4.1. Climate Modification of Ecosystem Response to Nitrogen
Climate-related changes including temperature, precipitation, wind patterns, extreme
weather events, stronger estuary stratification, increased metabolism and organic
production, and sea level rise are all expected to modify coastal habitats (Altieri and
Gedan. 2015; Statham. 2012; Rabalais et al.. 2009). These interacting factors will alter
sensitivity to N loading and ecosystem response to nutrient inputs. For example,
eutrophic conditions and the extent and duration of hypoxia are predicted to increase with
changes in temperature and precipitation (Altieri and Gedan. 2015; Rabalais et al.. 2009;
Boesch et al.. 2007). Freshwater and nutrient contributions to estuaries are expected to
rise due to predicted increases in surface water flow and runoff from watersheds
(Rabalais et al.. 2010; Adrian et al.. 2009; Whitehead et al.. 2009). Howarth et al. (2012)
demonstrated larger N fluxes (larger percent delivery of human N inputs) in wetter
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climates with more discharge, across 154 different watersheds in the U.S. and Europe.
High organic loads and freshwater inputs associated with extreme weather events may
enhance stratification and contribute to hypoxia ("Wetz and Yoskowitz. 2013).
Temperature modification leading to sea level rise and inputs of freshwater will likely
alter salinity gradients and increase stratification within estuaries (Statham. 2012; Najjar
etal.. 2010V
In estuaries affected by nutrient enrichment and climate-associated alterations described
above, conditions favor a shift in phytoplankton toward greater abundance and
distribution of toxic cyanobacteria associated with increased prevalence of HABs and
declines in SAV (Pacrl et al.. 2016a; Najjar et al.. 2010). Research indicates thatN and
climate change will interact to drive losses in biodiversity that will be more than additive
compared to each independent force (Porter etal.. 2013). Decreases in pH associated with
nutrient-enhanced coastal eutrophication combined with elevated atmospheric CO2 could
increase susceptibility of fauna to ocean acidification I (Cai et al.. 2011c).
Appendix 10.51. Coral reef ecosystems are particularly susceptible to combined effects of
acidification, rising sea level, warming trends, and eutrophication. Appendix 13 includes
a more detailed discussion of how climate (e.g., temperature and precipitation) modifies
ecosystem response to N loading.
10.2. Indicators of Nutrient Enrichment
In the complex environment of the freshwater-to-ocean continuum there are many
chemical and biological indicators of eutrophic conditions (Table 7-8). One approach is
to measure total nutrient loading and concentrations; however, these data need to be
interpreted in the context of the physical and hydrological characteristics that determine
ecosystem response. Water quality measures (water clarity, DO) and quantification of
nutrients along with biological indicators such as chlorophyll a, phytoplankton
abundance, HABs, macroalgal abundance, SAV, and fish kills can all be used to assess
responses to nutrient loading. Some biological indicators such as chlorophyll a are
directly linked to nutrient enrichment and provide evidence of early response to added
nutrients, while other indicators such as low DO and decreases in SAV indicate that the
degree of eutrophication has progressed (Borja et al.. 2012). A summary of key indicators
of estuarine eutrophication is provided in Table 10-1. and these indicators are discussed
in the following sections. The 2008 ISA used the five ecological indicators shown in
Figure 10-3 (chlorophyll a, harmful/nuisance/toxic algal blooms, macroalgae, DO, and
SAV) included in the Assessment of Estuarine Trophic Status (ASSETS) categorical
Eutrophication Condition Index (ECI) in the NEEA to estimate the likelihood that an
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estuary is experiencing eutrophication or will experience eutrophication (Brickcr et al..
2007V
Table 10-1 Indicators of Estuarine Eutrophication.
Indicator	Description
Chlorophyll a	Excess N input will stimulate primary productivity, and chlorophyll a concentration is an
indicator of phytoplankton biomass. Phytoplankton biomass is strongly controlled in
coastal waters by the availability and supply rates of nutrients, especially N (Paerl and
Piehler. 2008).
Excess N input can cause nuisance or toxic algal blooms, which release toxins in the
water that can poison aquatic animals and threaten human health (Baron et al.. 2012).
Problem occurrences of HABs increased 13% from 1997 to 2007 in Mid-Atlantic
estuaries, the only region for which data were available, along with increasing N loading
(Bricker et al.. 2008).
Macroalgal abundance Macroalgal blooms can cause the loss of important submerged aquatic vegetation by
blocking sunlight, and opportunistic macroalgae can outcompete many seagrass
species (Kennison et al.. 2011). They can also cause hypoxia and smother seagrass,
coral, clams/oysters, and other benthic organisms. High biomasses also cause "stinky
beach" by rafting onto beaches and decaying.
Dissolved oxygen (DO) Dissolved O2 concentration decreases with increasing algal abundance under elevated
N because microbes consume Chas they decompose dead algae. Oxygen depletion
mainly occurs in bottom waters under stratified conditions. Increased atmospheric N
deposition combined with N loading from other sources will likely affect the size,
frequency, and severity of hypoxia events (Rabalais et al.. 2010). Hypoxia also
contributes to ocean acidification, which is detrimental to the growth and development
of calcifying organisms (Howarth et al.. 2011). The largest documented zone of hypoxic
coastal water in the U.S. is located in the northern Gulf of Mexico (Dale et al.. 2010).
Excess N input stimulates algal growth, and opportunistic macroalgae may block the
penetration of sunlight into the water column and outcompete seagrasses, leading to
reduced SAV coverage. Increased epiphyte loads on the surface of macrophytes from
nutrient enrichment may reduce biomass, shoot density, percentage cover, production
and growth of SAV (G.Nelson. 2017). Reduced extent of eelgrass was found to
correspond to increased loading of N (from wastewater, fertilizer, and atmospheric
deposition) to small to medium shallow estuaries in New England (Latimer and Reqo,
2010), and the distribution of SAV in Chesapeake Bay is used as an indicator of
diversity and biological balance in the U.S. EPA Report on the Environment (U.S. EPA.
2016k).
Chi = chlorophyll; DO = dissolved oxygen; EPA = Environmental Protection Agency; HAB = harmful algal blooms; N = nitrogen;
02 = oxygen; SAV = submerged aquatic vegetation.
Harmful/nuisance/toxic
algal blooms (HABs)
Submerged aquatic
vegetation (SAV)
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S.R Brickcr et at./Harmful Algae S (2008) 21-32
Impact: No Problem /low Moderate low
3
UJ
Moderate
Moderate high
High
\.\Y
MM
Few symptoms Symptoms occur Symptoms occur Symptoms occur Symptoms occur
occur 31 more than episodically and/or less regularly less regularly and/or periodically or
minimal levels over a small to and/or over a over a medium to persstendy arxl/or over
medium area medium area extensive area an extensive area
Influencing factors
ds and susceptibility)
Key to symbols
Submerged

aquatic vegetation
Chlorophyll a
Nuisance/toxic
blooms (HAB)
Macroalgae
Dissolved oxygen
Source: Bricker et al. (2008).
Figure 10-3 Biological indicator responses to nutrient enrichment.
10.2.1. Chlorophyll a
Chlorophyll a is part of the light harvesting complex (group of pigments) used by
photosynthetic organisms to convert light energy into carbohydrates through a series of
biochemical reactions. The concentration of Chlorophyll a is often used as a proxy for
phytoplankton biomass. Algae form the base of the coastal food web and excess algal
growth is often directly linked to nutrient enrichment. This indicator can be easily
quantified, and linked to aircraft and satellite-based remote sensing and autonomous
monitoring platforms to assess effects at the regional and ecosystem level
(Appendix 7.2.7). Chlorophyll a is widely used to assess eutrophic conditions because of
its sensitivity to nutrient inputs and was one of the indicators of overall eutrophic
condition of U.S. coastal areas in the 2008 ISA (U.S. EPA, 2008a). Increased levels of
chlorophyll a can signal an early stage of water quality degradation related to nutrient
loading and measurements of chlorophyll a are incorporated into water quality
monitoring programs (Appendix 7.2.7 and Appendix 7.2.10). High concentration of
chlorophyll a suggests that algal biomass is sufficiently high to contribute to low DO
concentration from increased decomposition of dead algae. Due to the strength of
hydrologic forces (i.e., freshwater inputs and tidal flushing) in some types of coastal
systems, benthic chlorophyll a may be a more sensitive indicator of ecosystem response
to N enrichment than planktonic chlorophyll a in well-flushed systems (Paerl and Piehler.
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2008). Phytoplankton blooms can also be advected into estuarine systems from the
coastal ocean (Brown and Ozretich. 2009).
Chlorophyll a concentrations are commonly included in standardized frameworks of
eutrophic condition [Appendix 10.2.6; (Borjaetal.. 2012)1. Many of these indices define
spatial extent and temporal sampling periods and statistical measures to determine
representative concentrations. Several chlorophyll a thresholds have been identified for
U.S. water bodies (Table 10-2). The U.S. EPA National Coastal Condition Assessment
(NCCA; formally known as National Coastal Assessment) measures chlorophyll a from
an annual index period (June to October) and compares the samples to reference
conditions to determine a rating [>20 |ig/L. poor; 5-20 |ig/L. fair; <5 |ig/L. good; (U.S.
EPA. 2016g. 2012bVI. For the NEEA ASSETS, the 90th percentile of annual values for
chlorophyll a combined with spatial coverage and frequency of occurrence of blooms are
reported for distinct salinity zones (tidal fresh 0-0.5 ppt), mixing zone (0.5 to 25 ppt),
and seawater (>25 ppt); 0-5 |ig/L is a water body with low risk of eutrophication,
5-20 |ig/L is moderate, >20 |ig/L is high (Bricker et al.. 2007).
Table 10-2 Chlorophyll a thresholds used in methods to evaluate the status of
phytoplankton in U.S. coastal and estuarine water bodies.
Method/
Approach
Chlorophyll a
Thresholds
and Ranges
(ng/L)
Sample
Time
Frame
Statistical
Measure
Other
Characteristics
Community
Composition
Indicators in
Overall
Eutrophication
Index
U.S. EPA
NCCA
Poor >20
Fair 5-20
Good 0-5
Index
period
(June-Oct.)
Concentration
percentage of
coastal area in
poor, fair, and
good condition
based on
probabilistic
sampling design
for 90% conf. in
areal result

No
Chlorophyll a,
water clarity, DO,
DIP, DIN
ASSETS
(eutrophic
condition
component
only)
Poor >20
Fair 5-20
Good 0-5
Annual
90th percentile
Chlorophyll a
concentration of
annual data
Spatial
coverage,
frequency
occurrence
Nuisance and
toxic bloom
occurrence,
frequency,
duration
Chlorophyll a,
macroalgae, DO,
seagrasses,
nuisance/toxic
algal blooms
ASSETS = Assessment of Estuarine Trophic Status; conf. = confidence; DIN = dissolved inorganic nitrogen: DIP = dissolved
inorganic phosphorus; DO = dissolved oxygen; EPA = Environmental Protection Agency; L = liter; NCCA = National Coastal
Condition Assessment.
Source: Boria et al. (2012).
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In the NEEA, elevated chlorophyll a concentration was the most widespread documented
symptom of eutrophication (Bricker et al.. 2007). Half of the estuaries for which data
were available exhibited high chlorophyll a concentration (Bricker et al.. 2007). In the
2008 ISA, San Francisco Bay, CA was an example of an estuary that has experienced
considerable increases in chlorophyll a concentrations in recent years.
Phytoplankton biomass, as indicated by chlorophyll a concentration, is strongly
controlled in estuaries by the availability and supply rates of nutrients, especially N (Pacrl
and Piehler. 2008). Bioassays conducted in the low-nutrient Alligator River estuary in
North Carolina showed that N enrichment is directly related to increasing chlorophyll a
concentration. Although the highest increase occurred in response to addition of both N
and P, dissolved inorganic N (DIN) treatment alone stimulated chlorophyll a in some
treatments (Rossignol et al.. 2011). A 3-year data set from Raritan Bay, NJ indicates that
nutrient loading contributed to high concentrations of chlorophyll a from 2010-2012
(Rothcnbcrgcr et al.. 2014). Winter and spring N loading to Mattawoman Creek estuary
and other shallow estuaries in the Chesapeake Bay region was highly correlated to
summer chlorophyll a concentrations (Bovnton et al.. 2014). However, after point source
nutrient (N and P) reductions were enacted, a decrease in chlorophyll a was observed in
Mattawoman Creek. In North Carolina's Neuse River estuary, which is part of the
Albemarle-Pamlico estuary, it appears that the elevated loading of total N (TN)
contributed to higher annual average chlorophyll a values from 2000-2009 (Lcbo et al..
2012).
While recent research indicates that N loading remains a strong predictor of chlorophyll a
concentrations under most conditions, there has been increasing discussion regarding the
role of other factors in altering the strength of this relationship. The impact of nutrient
inputs can at times be overtaken by hydrologic features, seasonal variations, climatic
changes, and oscillations exerting greater control over phytoplankton dynamics (Pacrl et
al.. 2010). In Buzzard's Bay, MA, both nutrient loading and shape of the embayment
were factors in determining spatial and temporal water quality trends including the
observation that chlorophyll a is increasing at a faster rate than N enrichment (Rheubanet
al.. 2016). It appears that more chlorophyll a per unit of TN was produced as the 22-year
time series progressed. In Corpus Christi Bay, TX, rainfall events altered nutrient N:P
ratios but did not affect chlorophyll a (Turner et al.. 2015). Instead, chlorophyll a
concentration in weekly samples mainly followed seasonal trends by increasing in spring
and summer and decreasing in fall and winter. No significant relationships were observed
between annual TN, total phosphorus (TP) load, and annual mean chlorophyll a
concentrations in the Guana Tolomato Matanzas estuary in Florida, rather phytoplankton
biomass in this well-flushed system was influenced by temperature, precipitation, water
residence times, and tidal exchange (Hart et al.. 2015). (Glibert et al.. 2014) found
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significant increases in chlorophyll a over time at only three out of seven study areas in
the Maryland/Virginia coastal lagoon, although regionally chlorophyll a concentrations
increased due to increasing anthropogenic nutrient loads and increased freshwater flow in
the early 2000s. In the Neuse River estuary, NC, which has a long residence time
allowing for detection of nutrient stimulation, nutrient addition bioassays along the
estuary indicated strong N limitation at the chlorophyll a maximum and downstream
where there was a strong preference of NH44" over NOs as a DIN source (Paerl and
Piehler. 2008).
Using satellite and meteorological data from U.S. Atlantic coastal waters, (Kim et al..
2014b) found that precipitation events were associated with increased levels of
chlorophyll a in low nutrient areas (defined as having NO3 concentrations of less than
1 |iM). but precipitation was correlated with lower chlorophyll a concentrations in high
nutrient areas (defined as having NO3 concentrations greater than 1 |iM). This is likely
because in low nutrient areas, new N input from precipitation stimulated phytoplankton
growth. In areas already high in nutrients, wind associated with precipitation events may
have deepened the mixed layer and the resulting loss of light availability caused
chlorophyll a concentrations to decline (Kim et al.. 2014b). Two studies modeling
historical data from Chesapeake Bay indicate that variation in climatic conditions
dominated phytoplankton dynamics in the bay in recent years (Harding et al.. 2016a;
Harding et al.. 2016b). Much of the bay is producing more chlorophyll a per unit TN than
in the past, leading Harding and others to suggest that return to the historical relationship
between N and chlorophyll a is unlikely.
In four coastal estuaries that shifted from eutrophication to oligotrophication (lower
nutrients), the degree of return of chlorophyll a to reference status varied, likely due to
concurrent changes in the estuaries from costressors and the time elapsed from
eutrophication to nutrient reduction (Duarte et al.. 2009). Due to these shifting baselines
the authors suggest that the current paradigm of nutrient reduction to a historical level
needs to be replaced by targets that maintain key ecosystem functions in the context of
changing conditions in the estuaries over time.
10.2.2. Harmful/Nuisance/Toxic Algal Blooms
Nuisance or toxic algal blooms reflect the proliferation of a toxic or nuisance algal
species that negatively affects natural resources or humans. Blooms are increasing in
outbreak frequency and extent in the U.S. and other countries, and N is one of the
nutrients known to promote HAB formation (Baron et al.. 2012; Heisler et al.. 2008). A
concurrent increase in atmospheric sources of N to coastal areas and proliferation of
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harmful algal bloom formation have been recognized for several decades (Paerl et al..
2002; Paerl and Whitall. 1999; Paerl. 1997). The form of N delivered to coastal regions
of the U.S. from atmospheric and other sources is changing from primarily NOs to an
increase in reduced forms of N (Appendix 10.1.2). which are favored by some HAB
forming species (Glibert et al.. 2016). HABs caused by N enrichment can release toxins
that are harmful to fish and shellfish and that may accumulate in predators and organisms
higher in the food web (Johnson et al.. 2010). For example, cyanobacteria blooms
(cyanoHAB) have been documented in estuaries along the U.S. Mid-Atlantic and
Southeast coasts as well as Gulf coast estuaries (Preece et al.. 2017). Other blooms are
not toxic but may cause low DO events due to very high biomass. In addition to effects
on biota, HABs cause a range of responses in humans from direct dermatitis, such as
swimmers itch, to severe food poisoning, liver and kidney toxicity, and paralysis (Peel et
al.. 2013; Johnson et al.. 2010).
The frequency and duration of algal blooms is one of five indicators used in the NEEA
and was included as an indicator of eutrophic conditions in the 2008 ISA (U.S. EPA.
2008a; Bricker et al.. 2007). Of the 81 estuary systems for which data were available in
the NEEA, 26 exhibited a moderate or high symptom expression for nuisance or toxic
algae (Bricker et al.. 2007). In the previous 2008 ISA review, the frequency of
phytoplankton blooms and the extent and severity of hypoxia were documented to have
increased in the Chesapeake Bay (Officer et al.. 1984) and Pamlico Sound estuaries in
North Carolina (Paerl and Piehler. 2008) and along the continental shelf adjacent to the
Mississippi and Atchafalaya river discharges to the Gulf of Mexico (Eadie et al.. 1994).
New tools and monitoring approaches to further characterize HABs have become
available since the 2008 ISA (Appendix 7.2.7). For example, solid phase adsorption toxin
tracking (SPATT), a passive sampling tool which simulates the contamination of
filter-feeding bivalves, has been used recently in the field to detect the presence of algal
toxins with greater accuracy than grab-sampling methods (Gibble and Kudela. 2014;
Kudela. 2011; Lane et al.. 2010). Remote sensing systems are increasingly being used to
forecast and monitor HABs in coastal waters (Klemas. 2012).
Harmful effects of HAB toxins on fish and wildlife are readily transferred through the
food web and may persist even when the bloom conditions have passed. Microcystin, a
class of toxins produced by many cyanoHABs is found in all trophic levels from shellfish
to finfish to top level predators and can persist in the foodweb for months (Preece et al..
2017).(Wood et al.. 2014) found that microcystin persisted in overwintering populations
of estuarine finfish, common wedge clam (Rangia cuneata) and blue crab (Callinectes
sapidus) in the James River estuary, VA although the highest tissue concentrations and
greatest percentage of individuals affected were observed when toxins in the water
column were at maximum levels. The toxin was present in both muscle and viscera of
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blue crabs at concentrations that have been shown to have physiological effects on other
species of estuarine crab (Wood et al.. 2014). Accumulation of microcystin measured
over the course of 2 years in the James River estuary, near Chesapeake Bay, was found to
be highest in suspension feeding animals while top predators (piscivores), scavengers,
and benthic feeders all had lower levels of microcystin (Bukavcckas et al.. 2017). In
Monterey Bay, CA, deaths of 21 sea otters (Enhydra lutris), a federally listed threatened
species, were attributed to hepatotoxic shellfish poisoning due to trophic transfer of
microcystin observed in this study to have originated from nutrient-impaired lakes and
rivers discharging to the bay (Miller et al.. 2010). Other HAB toxins, such as domoic
acid, are also known to be transferred up the food web (Trainer et al.. 2012). Domoic acid
from recent blooms of the diatom Pseudo-nitzschia along the U.S. West Coast has
impacted razor clam (Siliqua patula) and Dungeness crab (Metacarcinus magister)
fisheries, led to symptoms of domoic acid poisoning in California sea lions (Zalophus
californianus), and was detected in additional marine mammals from southern California
to northern Washington (Mccabe et al.. 2016).
Research on HAB-forming species have shown that the form of N supplied affects
phytoplankton growth (Glibert et al.. 2016). Generally, NH4+ is considered to be the
preferred form of N for some phytoplankton due to lower energy requirements for uptake
and assimilation; however, diatoms specialize in use of oxidized N forms (Glibert et al..
2016). For example, the HAB-forming dinoflagellate species Heterosigma akashiwo is
able to grow well with a pulsed supply of NH/, NOs . and urea under low light
conditions and assimilate stores ofN nutrients, suggesting that this species could
outcompete diatoms in silicate-limited, N enriched coastal areas (Kok et al.. 2015). In
Pseudo-nitzschia spp., laboratory experiments showed that both of the species studied
were able to grow on NH4+, NO3 . and urea, although urea was the preferred form of N
(Melliti Ben Garali et al.. 2016). These findings were supported by a field experiment in
which chlorophyll a concentration significantly increased and exponential growth
occurred in all N-enriched in situ microcosms until the end of the experiment, with
specific growth rates highest in the urea and NO3 additions. The dinoflagellate Akashiwo
sanguinea showed different growth profiles and N assimilation with form of N and
concentration, growing faster in NH4 and with greater enzyme affinity for urea (Liu et
al.. 2015). Differential growth responses of HAB species to reduced N can alter
phytoplankton community composition and biodiversity (Appendix 10.3.3).
There is also evidence, mostly from freshwater systems, that the form of N affects toxin
production of some HAB species (Appendix 9.2.6.1). In nutrient amendment experiments
with water collected from the tidally influenced Transquaking River, which flows into
Chesapeake Bay, Microcystis was stimulated by N more frequently than P, and
abundances of toxic and nontoxic strains were enhanced to different degrees by inorganic
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and organic N (Davis et al.. 2010). Response of Alexandrium fundyense to nutrient
addition varied throughout the course of bloom events in Northport-Huntington Bay, NY
(Hattcnrath et al.. 2010). Addition of NH44" to bloom water most frequently resulted in
statistically significant increases of A. fundyense density and toxin concentration
compared to other forms of N (glutamine, NOs . and/or urea). Davidson et al. (2012)
found that, in general, laboratory studies show that toxin production can be influenced by
nutrient ratios, but extrapolation of those results to the specific species and conditions in
the field is difficult.
Modeling studies have reported on potentially altered future scenarios of phytoplankton
community changes and HAB formation, intensity, duration, and toxicity due to changes
in N deposition (Lee and Yoo. 2016; Glibert et al.. 2010b). Pinder et al. (2008) reported
on modeled scenarios that predict alterations in the frequency, intensity, toxicity, and
species composition of algal blooms due to higher rates of N deposition and changes in
the reduced N to oxidized N ratio. Future changes in HAB dynamics will be affected by
climate change (Appendix 10.1.4.1) and increased N loading, and integrated ecosystem
models that couple the atmosphere, land, and coastal ocean are needed to estimate these
HAB responses (Glibert et al.. 2010a). Long-term monitoring (1987-2005) of
phytoplankton populations in the Bay of Fundy in southwestern New Brunswick, Canada
did not link HABs to nutrients, rather many species abundances are explained by climate
and weather patterns (Martin et al.. 2009). Vox A. fundyense there was a negative
relationship with cell density and NO3 .
Table 10-3 summarizes new studies from U.S. waters on levels and forms of N at which
effects are manifested in phytoplankton.
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Table 10-3 Levels and forms of nitrogen at which effects on phytoplankton are manifest in U.S. coastal waters
evaluated in the literature since the 2008 Integrated Science Assessment for Oxides of Nitrogen and
Sulfur-Ecological Criteria.
Study Site
Ambient N
Deposition
N Additions
Ecological/Biological Effects
Species
Reference
Northport-
Huntington Bay
complex, NY
Addition of NH4+
(10 |jM-40 |jM)
Addition of NhV significantly increased A. fundyense Phytoplankton
densities compared to the control. The addition of NHV (Alexandrium
(40 |jM) yielded a significant increase in both A.	fundyense)
fundyense densities and toxin concentrations four and
8x, respectively, compared to controls.
Hattenrath et al.
(2010)
Raritan Bay, NJ
Ambient levels Multivariate analyses of a 3-yr data set indicated that the Phytoplankton
abundance of HAB species Heterosigma akashiwo is	(Heterosigma
positively associated with NO3" in Raritan Bay. Both	akashiwo and
climatic conditions and nutrient concentrations affect	13 other HABs
phytoplankton bloom composition in the bay.	identified)
Rothenberqer et
al. (2014)
Raritan Bay, NJ
Nutrients (N and
Fe) added alone or
in combination to
different treatments.
N was added as a
single pulse sodium
nitrate (NaNOs) to
increase NO3"
concentrations by
approximately
10 |jM. The N
additions lowered
the Si:N ratios from
~3 to <1
Dinoflagellates and HAB-forming taxa increased to a
greater extent when NO3" levels were high (which led to
a low Si:N ratio of less than one). Centric and
chain-forming diatoms resulted from enriched NO3"
concentrations, differing from the pennate diatoms and
green flagellates that accompanied ambient NO3"
concentrations. Dinoflagellates in the genus Dinophysis,
which could be HAB-forming, also resulted from enriched
NO3" and lowered Si:N.
Phytoplankton
Dinoflagellates,
diatoms
Rothenberqer
and Calomeni
(2016)
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Table 10-3 (Continued): Levels and forms of nitrogen at which effects on phytoplankton are manifest in U.S.
coastal waters evaluated in the literature since the 2008 Integrated Science Assessment
for Oxides of Nitrogen and Sulfur-Ecological Criteria.

Ambient N




Study Site
Deposition
N Additions
Ecological/Biological Effects
Species
Reference
Ocean surface
Based on direct

Precipitation events in coastal waters of the eastern U.S.
Phytoplankton
Kim et al.
from 28°N to
measurements of

increased the chlorophyll a concentration up to 15% in

(2014b)
44°N and from
wet deposition along

low-nutrient areas (<1 |jM NO3") but decreased the


the East Coast
the East Coast of

chlorophyll a concentration in nutrient-replete areas


of the U.S. to
the U.S., the N

(>1 |jM NO3"). The authors suggested that in


60-70°W
supply through wet
deposition was
estimated to be
25-45 mmol N/m2/yr

nutrient-depleted areas (south of 36°N), the added
nutrients were a dominant factor increasing the
chlorophyll a concentration, whereas in the
nutrient-replete areas (north of 36°N), where
phytoplankton growth was light limited, reduced light
availability was the dominant factor determining reduced
chlorophyll a concentration.


Matta woman
Used atmospheric

Strong relationships were found between N loading and
Phytoplankton and
Bovnton et al.
Creek,
deposition data from

algal biomass and between algal biomass and water
SAV
(2014)
Chesapeake
Bovnton et al.

clarity. Winter-spring N loading and summer chlorophyll


Bay
(2008). includina N
from both wet and
dry deposition
(0.81 mg N/L as an
annual average
concentration).
Direct atmospheric
deposition to surface
waters of the creek
contributed about
6,000 kg N/yr or
about 16 kg N/day to
the creek system.

a were found to be highly correlated, a relationship which
appears to be linear.


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Table 10-3 (Continued): Levels and forms of nitrogen at which effects on phytoplankton are manifest in U.S.
coastal waters evaluated in the literature since the 2008 Integrated Science Assessment
for Oxides of Nitrogen and Sulfur-Ecological Criteria.
Study Site
Ambient N
Deposition
N Additions
Ecological/Biological Effects
Species
Reference
Tidally
influenced
Transquaking
River which
flows into
Chesapeake
Bay
20 |jM (NOs"),
20 |jM NH4+ (NH4+),
10 |jM (= 20 |jM N)
urea, 10 pM
(= 20 |jM N),
L-glutamine (GA), P
(1.25 |jM
orthophosphate), or
a combined
treatment of NO3"
and P
Microcystis was simulated by N more frequently than P,
and abundances of toxic and nontoxic strains were
enhanced to different degrees by inorganic N and
organic N. Toxic Microcystis abundance increased more
with inorganic N than organic N.
Toxic and nontoxic
strains of Microcystis
Davis et al.
(2010)
Maryland and
Virginia coastal
bays
Measurements of
atmospheric
deposition since
2000, based on the
NADP, suggest that
NO3" is decreasing
and NH4+ from
deposition is stable
for the coastal bays
(NADP data
http://nadp.sws.uiuc.
ed u/sites/ntn/NTNtre
nds.html?sitelD=MD
18).
N in the water column is dominated by reduced N
(primarily NH4+) with low concentrations of NO3" also
present, resulting in phytoplankton community shifts to
those species that can do well under such conditions.
Submerged aquatic vegetation has decreased.
Phytoplankton and
SAV
Glibert et al.
(2014)
Alligator River
estuary, NC
Potential for
atmospheric
deposition of N from
nearby farm (wet
NH4+ concentrations
increased greatly
close to farm).
DIN additions:
140 |jg N-NhV/L,
140 |jg N-NO3-/L,
and 70 pg N-NhV/L
plus 70 pg
N-NCV/L
Significant increase in phytoplankton biomass (chl a) and Phytoplankton
rates of primary productivity due to N enrichment. DIN
treatments alone significantly stimulated chl a in two out
of five tests for all three DIN addition treatments,
although DIN + DIP treatments provided the largest
increase in chl a.
Rossiqnol et al.
(2011)
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Table 10-3 (Continued): Levels and forms of nitrogen at which effects on phytoplankton are manifest in U.S.
coastal waters evaluated in the literature since the 2008 Integrated Science Assessment
for Oxides of Nitrogen and Sulfur-Ecological Criteria.

Ambient N




Study Site
Deposition
N Additions
Ecological/Biological Effects
Species
Reference
Neuse River

Ambient, acute DIN
Study noted increase in phytoplankton bloom frequency
Phytoplankton
Paerl et al.
estuary, NC

inputs via runoff
and magnitude (indicated by increasing chl a variability)

(2010)


from hydrologic
overtime in response to acute DIN inputs from




pulses (hurricanes,
hydrologic pulses. Control of algal bloom duration,




tropical storms,
thresholds, taxonomic composition, and spatial extent




heavy rainfall
may be dictated by climatic changes and oscillations




events)
instead of nutrient inputs.


New River

N addition to
Varying the form of nutrients promoted growth of
Phytoplankton
Altman and
estuary, NC

estuary water in the
different phytoplankton groups based on photopigment

Paerl (2012)


form of DIN, organic
analysis. Dinoflagellates, chlorophytes, and




N from river water
cyanobacteria responded to dissolved organic N while




(with dissolved
cyanobacteria increases were most frequent with




inorganic P) or urea
inorganic N addition.


Four sites with

NH4+, NOs", and
Along a gradient from highly developed to undeveloped
Phytoplankton
Reed et al.
tidal influence,

urea treatments
sites, phytoplankton communities at the more developed

(2016)
South Carolina

applied separately
sties had higher biomass and growth rate with N


coast

to bioassays. N
(particularly urea) additions and potentially HAB forming




forms were added
species were more often found at the more developed




at Redfield ratios
sites.




(N:P = 16:1)



Ten Mile Creek,

High median
Chl a was negatively correlated with N concentrations
Phytoplankton
Yana et al.
Indian River

concentrations of
and the highest chl a concentrations were related to the

(2008)
Lagoon, FL

total N
conditions of static and open water with long residence




(0.988 mg/L),
time. Hydrodynamics of a system may play an overriding




NO3--N
role in controlling phytoplankton growth.




(0.104 mg/L),





NH4+-N





(0.103 mg/L),





and total Kjeldahl N





(0.829 mg/L)



San Francisco

NOs" and NH4+
A comparison of N isotopes in cells of the HAB species
Phytoplankton
Lehman et al.
Bay/estuary

measured from two
Microcystis aeruginosa with N in rivers flowing into the
(cyanobacteria
(2015)


source estuaries
bay suggested that NH4+, not NCVwas likely the primary
Microcystis




source of N that supported the bloom.
aeruginosa)

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Table 10-3 (Continued): Levels and forms of nitrogen at which effects on phytoplankton are manifest in U.S.
coastal waters evaluated in the literature since the 2008 Integrated Science Assessment
for Oxides of Nitrogen and Sulfur-Ecological Criteria.
Ambient N
Study Site	Deposition	N Additions	Ecological/Biological Effects	Species	Reference
San Francisco
Ambient nitrate
The greatest chl a concentration and cell density
Phytoplankton
Lehman et al.
Bay/estuary
ranged from 0.19 to
occurred in the San Joaquin River estuary which had a
(cyanobacteria
(2008)

0.36 mg/L. Ambient
high average nitrate concentration (0.36 mg/L). The
Microcystis


ammonia ranged
second highest chl a concentration and cell density
aeruginosa)


0.02 to 0.06 mg/L.
occurred in the Old River estuary which had relatively
low N concentration (mean 0.19 mg/L). Differences in chl
a were correlated with environmental conditions,
particularly streamflow and water temperature and
secondarily to nutrient concentrations and ratios.


DIN = dissolved inorganic nitrogen; DON = dissolved organic nitrogen; HAB = harmful algal blooms; N = nitrogen; NADP = National Atmospheric Deposition Program;
NH4+ = ammonium ion; N03" = nitrate; P = phosphorus; SAV = submerged aquatic vegetation; W = west.
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10.2.3. Macroalgal Abundance
The abundance of macroalgae, which are generally referred to collectively as "seaweed,"
was an indicator of eutrophic condition in the 2008 ISA (U.S. EPA. 2008a'). Macroalgae
have been combined with other indicators to classify estuarine condition (Borja ct al..
2012). For example, NEEA includes an assessment of macroalgae (Brickcr et al.. 2007).
Macroalgal blooms can contribute to loss of important SAV by blocking the penetration
of sunlight into the water column. Although macroalgal data for estuaries in the U.S.
were generally sparse, the NEEA reported high macroalgal expression in 15 of the
64 estuaries evaluated (Brickcr et al.. 2007). In some lagoons with limited oceanic
exchange, macroalgae may be a more sensitive biological indicator than phytoplankton
(e.g., McLaughlin et al.. 2014; Nobre et al.. 2005). For example, in most estuaries of the
Southern California Bight, macroalgae is the dominant primary producer and a key
indicator of eutrophication (McLaughlin et al.. 2014). Macroalgae may not be a good
indicator of eutrophication in some upwelling-influenced estuaries in the Pacific
Northwest because an increase in macroalgal biomass in these systems does not appear to
be associated with temporal declines in eelgrass (Hessing-Lewis et al.. 2015; Hessing-
Lewis and Hacker. 2013; Hessing-Lewis et al.. 2011).
Opportunistic, fast-growing macroalgae can exhibit very high rates of N uptake during
periods of high N availability and can often outcompete or block out light for other
macrophytes when N loads are high and variable (Abreu etal.. 2011). This growth of
macroalgae can also smother corals, clams, oysters, and other biota (Bricker et al.. 2007)
and contribute to declines in seagrasses (Olvarnik and Stachowicz. 2012). These
macroalgae also tend to preferentially uptake N in the form of NH4+. In a Danish study,
sea lettuce (Ulva lactuca), which is a common species of macroalgae in U.S. coastal
waters, grew faster and exhibited greater biomass when subjected to NH4 as the N
source compared to NOs (Ale et al.. 2011). This difference was thought to be due to
reduced N being more easily assimilated and used by algae. Similarly, in experimental
manipulations with Gracilaria tenuistipitata, an opportunistic macroalgal species from
China, when both NH4 and NO3 were available, NH4 was assimilated more rapidly and
algal biomass was higher than with NO3 addition alone (Wang et al.. 2014a). Growth of
Caulerpa cylindracea, an invasive macroalga in the Mediterranean Sea, was not inhibited
by high NH3, and was able to outcompete native macroalgae in experimental plots with
nutrient addition (Gennaro et al.. 2015).
Southern California estuaries, some of the most nutrient enriched in the world, have not
been well studied in the past, but recent work is providing important information about
macroalgal growth in response to N loading to these systems. Opportunistic macroalgae
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(Ulva intestinalis and Ulva expansa) were shown to take up NOs, from the water very
efficiently at many concentration levels, giving these species the ability to outcompete
other algae in estuaries subject to varying nutrient loads (Kcnnison ct al.. 201 1). Algae
that were depleted in N took up NO3 at higher rates than enriched algae, and uptake rates
slowed as the algae became saturated with N.
However, macroalgal abundance was not found to be directly related to water or sediment
N concentrations in Southern California estuaries (Kcnnison and Fong. 2014). Rather,
macroalgal blooms occurred throughout the year and appeared to be influenced by the
unique physical and hydrological differences between estuaries. For instance, in one
estuary with high nutrient levels in both water and sediment, little to no macroalgal
biomass was present, possibly due to high water velocities that prevented young algal
filaments from attaching(Kcnnison and Fong. 2014). In other estuaries, comparatively
high algal biomass was found year-round even when nutrient supplies were lower than
other sites (although still N enriched), perhaps due to low rates of tidal flushing and
longer water residence times. Internal nutrient processing, variable hydrological regimes,
and multiple external nutrient sources may all contribute to eutrophic conditions
(Kennison and Fong. 2014).
10.2.4. Dissolved Oxygen
DO was included in the 2008 ISA, the NCCA, and the NEEA as an indicator of eutrophic
condition (U.S. EPA. 2008a). Additional information on DO is provided in
Appendix 7.2.3. Oxygen depletion largely occurs only in bottom waters under stratified
conditions, not throughout the entire water column. The decomposition of organic matter
associated with increased algal abundance consumes DO and can reduce DO
concentrations in eutrophic waters to levels that cannot support aquatic life (Jcwctt ct al..
2010). Respiration of microbes, macrophytes, and animal biota can also reduce DO to
very low levels such as primary producers in the dark, deeper portions of the well-mixed
Hudson River estuary (Howarth et al.. 1996b) and as seen every night in a eutrophic
seagrass-dominated system on Cape Cod (Howarth et al.. 2014). In the Cape Cod
seagrass system, hypoxia is common at dawn, following hours of darkness, yet oxygen
levels are supersaturated at the end of the daylight period.
Generally, some biota are impacted at DO levels from 3 to 4 mg/L, and increasingly
adverse effects are observed on biota at lower DO concentrations (Figure 10-4).
Decreased DO can lead to the development of hypoxic or anoxic zones that are
inhospitable to fish and other life forms and can impact ecosystem processes (Diaz et al..
2013; Levin et al.. 2009; Diaz and Rosenberg. 2008). For example, in Chesapeake Bay,
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Sturdivant et al. (2014) observed that macrobenthic production was 90% lower during
hypoxic conditions, resulting in abiomass loss similar to 7,320-13,200 metric tons C
over an area of 7,720 km2. The authors estimated this change represented a displacement
of 20 to 35% of macrobenthic activity during the summer.
Mobile Fauna Begin
to Migrate to Higher
DO Areas
100% Saturation
Avoidance by
Fishes
Fishes Absent
Shrimp & Crabs
Absent
Burrowing Stops
Stressed Fauna
Emerge & Lay on
Sediment Surface
Fauna Unable to Escape
Initiate Survival
Behaviors
Mortality of Sensitive
Fauna
Mortality of Tolerant
Fauna
Formation of Microbial
Mats
Sediment Geochemistry
Drastically Altered
No Macrofauna
Survive
Hydrogen Sulfide Builds Up
\ Water Column
DO = dissolved oxygen; I = liter; mg = milligrams.
Source: Diaz et al. (2013V
Figure 10-4 The range of ecological impacts exhibited as dissolved oxygen
levels drop from saturation to anoxia.
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In the U.S., the incidence of hypoxia has increased almost 30-fold from 1960 to 2008 and
is now reported in more than 300 systems (Jcwett et al.. 2010). Climate change and
increasing N loading to coastal ecosystems are both expected to widen the distribution
and increase the size of areas affected by eutrophication-induced hypoxia, while also
increasing the frequency and persistence of these events KRabalais et al.. 2010);
Appendix 131. Areas of eutrophication-related hypoxia are found on the U.S. East and
West coasts and the Gulf of Mexico (Figure 10-5). Eutrophication-induced hypoxia,
which has been documented globally, can be characterized by both the duration of the
event and the ecosystem response (Diaz et al.. 2013; Diaz and Rosenberg. 2008).
Summer hypoxia is most common, followed by periodic oxygen (O2) depletion that may
occur in some systems more often than seasonally. In these hypoxic and anoxic areas
("dead zones") only organisms that can live with little or no O2 are present (Diaz et al..
2013; Jewett et al.. 2010). Once an ecosystem reaches severe seasonal hypoxia, there is a
shift to benthic organisms with shorter life spans and smaller body size (Diaz and
Rosenberg. 2008).
Zhou et al. (2014b) estimated the hypoxic volume of Chesapeake Bay over a 25 year
period. The date of onset of hypoxia occurs earlier in the summer and the end of hypoxia
also shifted to earlier in the fall with no trend in the seasonal-maximum hypoxia itself
from 1985 to 2010. Nutrient loading from the Rappahannock, Susquehanna, and Potomac
Rivers explains >85% of the seasonally averaged interannual variability in hypoxic
volumes. Testa and Kemp (2012) investigated interactions between hypoxia and nutrient
cycling in Chesapeake bay, based on analysis of long-term monitoring data collected
during two time periods: 1965 to 1980 and 1985 to 2007. They found that bottom water
in the upper Bay region, where seasonal hypoxia first develops, was enriched in NH4+
and PO4 relative to other regions of the bay. This might help explain the occurrence of
extensive and persistent hypoxia during the month of June even during years of lower N
loading.
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Eutrophic arid Hypoxic Coastal Areas of North America and the Caribbean
Source: http://www.wri.ora/resources/maps/coastal-eutrophic-and-hvpoxic-areas-north-america-and-caribbean modified from Diaz et
al. (2013).
Figure 10-5 Coastal eutrophic and hypoxic areas of North America the
Caribbean.
Recent research has shown that the duration of a hypoxic event is one of the most
important factors affecting macroinvertebrate density and species composition. (Baustian
and Rabalais. 2009) found that hypoxia duration (the number of days during which O2
concentrations were below 1 mg/L) was highly correlated with both reduced species
density and diversity in the northern Gulf of Mexico. Very low O2 levels can be lethal to
fish with no means of escape, and it has been suggested that even sublethal hypoxia may
lower the breeding rate and affect fish populations (Moran et al.. 2010). Hypoxia has
been shown to act as an endocrine disruptor in Atlantic croaker (Micropogonias
imdidatiis) in laboratory studies and biomarkers of reproductive function and endocrine
disruption are observed in field-collected individuals (Murphy et al.. 2009; Thomas and
Rahman. 2009; Thomas et al.. 2007). Hypoxic conditions increase nitric oxide and super
oxide radicals in brain tissue in croakers, causing cellular oxidative damage and inhibited
protein expression in the hypothalamus and leading to neuroendocrine effects (Rahman
and Thomas. 2015). Low O2 conditions have also been shown to alter animal behavior
(Appendix 10.4.1). O2 content influences hatching rate and parental effort among other
reproductive behaviors in three-spined sticklebacks (Gasterosteus aculeatus), so
Eutrophic
#	Hypoxic
#	Sy9tems in Recovery
Dill RwWW Sel*Bn 2D1D.
www oil
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eutrophication-induced hypoxia may alter reproductive output in some fish populations
(Candolin. 2009). However, hypoxia does not appear to negatively affect fisheries below
what would be predicted from N loadings alone, except under circumstances in which
raw sewage is released or when critical habitat is lost for very sensitive species (Breitburg
ct al.. 2009). In a review of long-term chronic effects of hypoxia on commercially
important fishery species (Townhill et al.. 2017) noted that effects range from positive to
negative, and from physiological to ecosystem-level. Some fish can acclimate to hypoxic
conditions and/or take advantage of more susceptible prey, affecting food web dynamics.
Other species cannot avoid hypoxia (especially shellfish) and suffer physiological stress,
mortality, and/or increased predation. Other species have been shown to move away from
the area to avoid stress (Appendix 10.4.1). which also has cascading food web
implications.
The effects of low DO are influenced by the presence of multiple stressors. For example,
(Gobler et al.. 2014) examined concurrent effects of low DO and acidification on the
early lifestages of bay scallops (Argopecten irradians) and hard clams (Mercenaria
mercenaria). Observations in later lifestages of the clams indicated that growth rates
decreased by 40% in combined exposures to hypoxia and acidification. Additional studies
with earlier lifestages indicated effects were more severe with costressors than with either
hypoxia or acidification alone. Juvenile oysters (Crassostrea virginica) grown under
varying conditions of low pH and duration of hypoxia seemed to acclimate and were only
negatively affected by the most severe pH experiment (Keppel et al.. 2016). Growth rates
were reduced 30-37% initially by both brief, repeated hypoxia and long moderate
hypoxia events; however, at the end of the study most oysters were the same size
regardless of treatment. This study also reported that the initial effects on oyster growth
were more pronounced by constant moderate hypoxia (1.3 mg/L) than they were by
severe but cyclical hypoxia (0.5 mg/L).
At the time of the last review, it was documented that the largest zone of hypoxic coastal
water in the U.S., and the second largest in the world, was the northern Gulf of Mexico
Hypoxic Zone on the Louisiana-Texas continental shelf (Dale et al.. 2010; Jewett et al..
2010; U.S. EPA. 2008a). Since the 2008 ISA, observations continue to indicate a large
zone of hypoxia in the northern Gulf of Mexico during the summer months associated
with increased N loading from the Mississippi and Atchafalaya River Basins. The size of
the midsummer bottom-water hypoxia area (<2 mg/L DO) in the northern Gulf of Mexico
has varied considerably since 1985, with a long-term average of 13,751 km2 (5,240 mi2)
(U.S. EPA. 2015c). In the summer of 2017, the hypoxic zone in the Gulf was the largest
ever measured at 14,123 km2 (8,776 mi2) (U.S. EPA. 2017e). Alexander et al. (2008) used
the SPARROW water quality model to show that atmospheric deposition to watersheds in
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the Mississippi River Basin is the second largest source of N (16%) to the Gulf, after
effluents from corn and soybean production (52%).
The Hypoxia Task Force (Mississippi River/Gulf of Mexico Watershed Nutrient Task
Force) is a federal/state partnership (five federal agencies, the National Tribal Water
Council and 12 states bordering the Mississippi and Ohio rivers) established in 1997 with
the goal of reducing nutrient inputs and the size of the hypoxic zone in the Gulf of
Mexico (U.S. EPA. 2017d. 2015c). The task force is working toward the goal of reducing
the areal extent of the Gulf of Mexico hypoxic zone to less than 5,000 km2 by 2035, with
an interim target of 20% nutrient load reduction by 2025. This was revised from the 2001
action plan that called for the described reductions by 2015 (U.S. EPA. 2001). Each state
in the task force has a nutrient reduction plan, and progress toward stated goals are
reported in the Mississippi River/Gulf of Mexico Watershed Nutrient Task Force Report
to Congress (U.S. EPA. 2015c) as required by the Harmful Algal Bloom and Hypoxia
Research and Control Amendments Act of 2014 (HABHRC Act. 2014). Other states not
included in the Mississippi River/Gulf of Mexico Watershed Nutrient Task Force have
nutrient reduction programs in place.
Hypoxic conditions are reported from other U.S. locations (Figure 10-5). For example, in
the shallow estuary of Long Island Sound, atmospheric deposition is considered to
comprise a significant fraction of total N loading [19%; (Whitall et al.. 2004)1. and DO
levels below 3 mg/L are common, with levels below 2 mg/L known to occur. During
some years, portions of the Long Island Sound bottom waters become anoxic (DO
<1 mg/L), (Latimer et al.. 2014). The maximum extent and duration of hypoxic events
(<2 mg/L DO) in Long Island Sound has varied considerably since the 1980s (U.S. EPA.
2016k). Between 1987 and 2014, the average annual maximum extent was 98 km2
(61 mi2). In 2014, the hypoxic area was 32 km2 (20 mi2), with the lowest DO levels
occurring in the western end of the Sound. The shortest hypoxic event occurred in 2014,
lasting 2 days. The longest hypoxic event was 71 days in 1989. Over the full period of
record (1987-2014), the average annual maximum duration was 31 days. Other estuaries
identified in the 2008 ISA where hypoxic conditions occur included Chesapeake Bay and
the Pamlico Estuary in North Carolina (U.S. EPA. 2008a). Murphy et al. (2011) analyzed
a 60-year record of Chesapeake Bay hypoxic zone volumes and found a slight but
significant decreasing trend in late summer hypoxia, which aligns with the decrease in N
loading due to management controls. This analysis also showed a correlation between N
loading and the duration of summer hypoxia events (Murphy et al.. 2011). In a modeling
study in Chesapeake Bay using analysis of monitoring data, nutrient loading was
identified as the main mechanism driving interannual hypoxia variabi 1 it\ (Li et al..
2016b). In upwelling regions such as along the West Coast, hypoxic events driven by
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upwelling can be advected into estuaries from global ocean circulation rather than
anthropogic nutrient additions (Brown and Power. 2011; Brown and Ozretich. 2009).
10.2.5. Submerged Aquatic Vegetation
SAV, rooted vascular plants that grow to the surface but do not emerge from the water, is
important to the quality of coastal ecosystems because it provides habitat for a variety of
aquatic organisms, serves as nursery grounds for estuarine invertebrates and fish, absorbs
excess nutrients, and traps sediments (U.S. EPA. 2008a; Handlev et al.. 2007; Lefcheck
ct al.). Recently, the presence of seagrass beds was linked to decreased bacterial
pathogens of humans, fishes, and invertebrates in the water column and a lower incidence
of disease in adjacent coral reefs (Lamb et al.. 2017). The loss of SAV can, therefore,
have a cascade of effects on other ecosystem characteristics and ecosystem services.
Water clarity is important for photosynthesis of SAV (Handlev et al.. 2007). Water
quality changes associated with excess N nutrient inputs, such as excessive algal growth
and hypoxia can impact SAV extent. For example, declines of Johnson's sea grass
(Halophila johnsonii) which is currently listed as threatened under the U.S. Endangered
Species Act has been linked to eutrophication of coastal waters, specifically low DO and
algal blooms that alter habitat by covering up substate (Hernandez et al.. 2016). Elevated
levels of N tend to increase epiphytes on the surface of SAV, in the absence of other
limiting factors, contributing to declines in seagrass biomass, shoot density, percentage
cover, production, and growth (G.Nelson. 2017; Nelson. 2017).
Seagrass loss is occurring globally with nutrient enrichment as a major driving factor
contributing to declines in SAV coverage (Latimer and Rcgo. 2010; Wavcott et al..
2009). SAV is included as a biological indicator for estuarine condition for U.S. coastal
waters in ASSETS-ECI in the NEEA and in U.S. EPA's Report on the Environment (U.S.
EPA. 2016k; Bricker et al.. 2007). At the time of the 2008 ISA, estimates of historical
losses of SAV and declines in habitat were available for some coastal regions of the U.S.;
although, there were few data documenting the long-term response of SAV to N loading.
In Waquoit Bay, MA, (Valiela et al.. 1992) reported a strong negative relationship
between modeled N loading and measured eelgrass (Zostera marina) area based on
measurements of eelgrass coverage from 1951 to 1990. In the NEEA report, only a small
fraction of the estuary systems evaluated reported high severity of SAV loss from the
1990s to 2004 (Bricker et al.. 2007). in part, because historical losses wiped out
seagrasses in many estuaries, and recent changes do not seem of high severity.
Since the 2008 ISA, additional studies are available on the relationship between N
loading and SAV abundance, including the development of thresholds of response to N
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loading in seagrasses (Table 10-4). Seagrass dieback in Snug Harbor Cape Cod, MA in
2010 was linked to N loading, which stimulated growth of epiphytes on the seagrasses
(Howarth et al.. 2014). Orth et al. (2010) observed a consistent negative correlation
between SAV abundance and N loading based on water quality data in Chesapeake Bay
from 1984-2006. In the Potomac River, a major tributary to Chesapeake Bay, a reduction
in total N from point and nonpoint sources was significantly correlated to increased SAV
abundance and diversity using field data from 1990 to 2007 (Ruhl and Rvbicki. 2010).
Similarly, the reduction of N and P input from point sources into Mattawoman Creek, a
tributary to Chesapeake Bay, led to large increases in SAV coverage and density
(Bovnton et al.. 2014). Benson et al. (2013) identified a tidal-averaged total N
concentration of <0.34 mg/L as a threshold for healthy eelgrass in a survey of
19 Massachusetts estuaries. Eelgrass coverage decreases markedly in shallow estuaries in
New England with N loading rates >100 kg N/ha/yr (based on nitrogen loading model
[NLM] estimates from wastewater, fertilizer, and atmospheric deposition inputs).
Loading rates above 50 kg/ha/yr are likely to impact habitat extent (Latimer and Rcgo.
2010). These ranges were found to be comparable to loading thresholds identified in
other East Coast estuaries (Table 10-4). In a modeling study using NLM applied to
small- to medium-sized estuaries of southern New England, direct atmospheric
deposition to the water surface made up an average of 37% of the N input, while indirect
deposition via the watershed averaged 16% of N loading. The percentage, however,
varied widely for individual estuaries (Latimer and Charpentier. 2010).
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Table 10-4 Nitrogen loading thresholds from multiple watershed sources versus
eelgrass loss.
Loading Threshold (kg/ha/yr)
Description
>20
100
•	Areal cover of eelgrass sharply reduced
•	Meadows disappeared rCape Cod estuaries, n = 10: (Bowen et
al.. 2007)1
-30
>60
•	Substantial eelgrass loss (80-96% of bed area)
•	Total disappearance TCape Cod estuaries, n = 7: (Hauxwell et al.,
2003)]
>64a
• Threshold based on nonparametric change-point analysis [95%
probability of change; Chesapeake Bay estuaries, n = 101; (Li et
al., 2007)1
>52
• Threshold based on nonparametric change-point analysis [95%
probability of chanae: New Enaland estuaries, n = 57: (Latimer
and Reqo. 2010)1
Consensus of Literature
Percentage of Eelgrass Area Loss (n = 57)
Mean Median 25th Percentile 75th Percentile
% %
<50
62 73 39 78
51-99
00
CO
00
CD
00
K)
CD
00
>100
93 100 95 100
aThis only includes point source inputs.
Source: Latimer and Reqo (2010).
The extent of SAV is stable or increasing in some coastal areas of the U.S. In Chesapeake
Bay, SAV coverage has increased from 41,000 acres (16,600 hectares) in 1978 to a peak
of 97,000 acres (39,250 hectares) in 2016 based on data collected by the Virginia Institute
of Marine Science KVIMS. 2016): Figure 10-61 as reported in U.S. EPA's Report on the
Environment (U.S. EPA. 2016k). SAV acreage has fluctuated in the bay since 2002,
covering an estimated 60,000 acres in 2013 then increasing in the most recent surveys
conducted in 2015 and 2016. This estimate is based on black-and-white, l:24,000-scale
aerial photographs of the Chesapeake Bay. Aerial monitoring first occurred in 1978 and
has occurred annually since 1984 (except during 1988). The SAV survey targets SAV
species, which as a group are sensitive to disturbance, particularly from eutrophication
and associated reductions in light availability. Pilots follow fixed flight routes to
comprehensively photograph all tidal shallow water areas of the bay and its tidal
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tributaries. Orth et al. (2010) reported a strong correlation between tributaries where
nutrient reductions have occurred and increases in seagrass abundance; however, further
reductions are necessary to meet SAV restoration targets. (Lcfchcck et al.) project a
catastrophic 95% loss of eelgrass Z. marina in Chesapeake Bay (and the associated
ecosystem services provided by this habitat type) in the next 30 years given the
conservative expectation of 2°C increase in temperature and continued trajectory of 40%
decline in water clarity due to the additive effects of these stressors especially on shallow
eelgrass beds. They report a 29% decline in eelgrass area in Chesapeake Bay since 1991
using high-resolution aerial imagery and water quality data. The authors note that
eelgrass abundance has increased in recent years but that recovery is limited to shallow,
nearshore areas and represents only a fraction of pre-1970s distribution of this species.
Despite extensive restoration efforts in the Chesapeake Bay watershed, water quality
indicators (chlorophyll a, DO, and Secchi depth) and biotic metrics of ecological
condition have shown little improvement during a 23-year period (Williams et al.. 2010)
The one exception to the pattern of no improvement in water quality was an observed
increase in the amount of SAV. Overall, water quality has decreased and chlorophyll a
levels have increased since 1986.
In Tampa Bay, FL, (see Appendix 16.4) data on seagrass (primarily Thalassia
testudinum) extent is available as far back as the 1950s. In the 1970s and early 1980s, the
bay exhibited signs of increasing eutrophication including loss of seagrasses. Seagrass
coverage has recovered to relatively predisturbed conditions (approx. 15,380 ha)
following N controls implemented in the mid 1980s (Greening et al.. 2014; Greening et
al.. 2011). Although the nutrient-load hypothesis of (Brauer et al.. 2012) suggests that
algae can outcompete or block out light for other macrophytes under high or variable
nutrient loads, this is not always the case. For example, in Yaquina Bay in Oregon,
distribution of eelgrass has remained stable despite high nutrient loads and algal blooms
(Kaldv. 2014). and seagrass populations in Puget Sound, WA have also remained
relatively stable for the last 40 years (Shelton et al.. 2017).
In other U.S. coastal areas, SAV coverage has declined. SAV coverage in the coastal
lagoon of Maryland and Virginia showed an increase with time until the early 2000s, and
then a decrease in the past decade, likely due to increasing anthropogenic nutrient inputs
that were accelerated by increased freshwater flow over the same time period (Glibert et
al.. 2014). In an analysis of 12 years of data for eelgrass areal coverage along the
Massachusetts coastline, there was an overall decrease in seagrass abundance although
the amount of change was highly variable between individual embayments (Costello and
Kenworthv. 2011).
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Exhibit 1. Extent of submerged aquatic vegetation (SAV) in the Chesapeake
Bay, 1978-2016
Tj
1 50
100
5
Cn 50
Estimated additional acreage
' Mapped acreage
1978 1982 19S6 1990 1994 1 99B 2002 200S 2010 2014
Year
There were partial Bay-wide surveys from 1 979 to 1 983, and no survey in 1 9S8.
Information on the statistical significance of the trend in this exhibit is not currently available. For
more information about uncertainty, variability, and statistical analysis, view the technical
documentation for this indicator.
Data source: Chesapeake Bay Program, 201 7
SAV = submerged aquatic vegetation.
Source: U.S. EPA (2016k).
Figure 10-6 Extent of submerged aquatic vegetation in the Chesapeake Bay
1978-2016.
SAV is often at a competitive disadvantage under N enriched conditions due to the fast
growth of opportunistic macroalgae that preferentially take up NH4+ and can block light
from seagrass beds (Abrcu et al.. 2011). Eelgrass from the Pacific Northwest exhibited
increased growth rates with increasing NH4+ concentrations, but growth rate was not
related to NO;, concentration (Kaldv. 2014). Temperature was found to significantly
affect the response of eelgrass to nutrients, but the effects were not synergistic. A
significant increase in leaf length and decreases in shoot density and aboveground and
belowground biomass in eelgrass was observed concurrent with increased shading by
algae in field studies of 12 estuaries in Atlantic Canada (Schmidt et al.. 2012). In the
same study, C and N storage (two ecosystem services provided by eelgrass) declined with
increasing eutrophication.
A modeling framework to link variable freshwater inputs with seagrass (Hctlodule
wrightii, Thalassia testudinum) biomass and timing of growth and chlorophyll a was
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applied to the Caloosahatchee River estuary in southwest Florida (Buzzclli ct al.. 2014).
The model predicted organic N in the upper estuary, and chlorophyll a in time and space
reasonably well; however, there was more variation in NH44" while NOx (nitrate-nitrite)
was proportional to freshwater inflow. Overall, the model reflected variations in seagrass
biomass, although timing and growth did not match the variability of field observations.
A spatial model of coastal Australia was used to relate N loading from land to the extent
of SAV (Fcrnandcs et al.. 2015). The largest N plume was associated with discharges
from an industrialized estuary and a wastewater treatment plants. The location and size of
the N plumes changed with seasonal influences. The results of spatial model analysis
comparing the plume to seagrass distribution obtained from video surveillance showed
that dense seagrass meadows only occurred in areas that were unaffected by N plumes,
regardless of the seasonal influences on the plumes.
10.2.6. Indices of Estuarine Condition
Biological and chemical indicators have been used by the states to develop numeric
nutrient criteria for estuaries (Appendix 7.2.10). Indicators may also be combined into an
overall condition rating to measure ecosystem function, structure and processes in a
standardized approach (U.S. EPA. 2016g; Boria et al.. 2012; U.S. EPA. 2012b; Devlin et
al.. 2011; Bricker et al.. 2007). Several assessment frameworks for eutrophic condition
have been developed in the U.S. (e.g., U.S. EPA's NCCA and the NOAA NEEA
ASSETS-ECI) and other countries (e.g., Trophic State Index [TRIX], Institutfrangais de
recherche pour Vexploitation de la mer [IFREMER], transitional water quality index
[TWQI]); however, the applicability of a specific framework to areas outside of the
region where they were originally developed may be limited (McLaughlin et al.. 2014;
Boria etal.. 2012). Within the same estuary, results of assessment of eutrophic condition
may vary depending on the framework used for evaluation and the associated chemical
and biological indicators (McLaughlin et al.. 2014; Garmendia et al.. 2012; Devlin et al..
2011). In a comparison of methods to assess nutrient enrichment impacts, Devlin et al.
(2011) suggested that indices incorporating annual data, frequency of occurrence, spatial
coverage, secondary biological indicators, and a multicategory rating scale are more
robust and representative. Two of these methods that have been applied to U.S. waters
are ASSETS-NCI (Bricker et al.. 2007) and NCCA (U.S. EPA. 2016g. 2012b).
ASSETS-NCI is only applied to nutrients and focuses on response indicators rather than
chemical indicators, while the NCCA can be used for assessing other stressors to coastal
areas and integrates both chemical and biological data.
The ASSETS ECI in the NEEA (Bricker et al.. 2007) was used to estimate the likelihood
that an estuary is experiencing or will experience eutrophication based on five ecological
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indicators: chlorophyll a, macroalgae, DO, nuisance/toxic algal blooms, and SAV
[Figure 10-3; (Bricker et al.. 2007)1. ASSETS uses the frequency and spatial extent of
algal blooms combined with the 90th percentile of annual values for chlorophyll a
(Table 10-2). Estuaries are divided into salinity zones, and ratings are combined as an
area-weighted sum. Results from the NEEA were included in the 2008 ISA, and the
biological indicators are the same in both reports for nutrient effects in estuarine systems.
The NCCA reports represent collaboration between U.S. EPA, NOAA, U.S. Fish and
Wildlife Service, and coastal state agencies. The most recent sampling period was 2010
(U.S. EPA. 2016g). NCCAs use chlorophyll a DO, and three additional indicators (DIN,
DIP, water clarity) to determine a water quality index (U.S. EPA. 2016g. 2012b). The
NCCA include data on water quality, sediment quality, benthic community composition,
and fish tissue contaminants to determine the overall condition of the nation's coastal
waters. In the most recent NCCA report, water quality was rated good in 36% of coastal
and Great Lakes nearshore waters, fair in 48%, and poor in 14%, based on measures of
the eutrophication parameters that make up the water quality index (P, N, water clarity,
chlorophyll a, and DO concentrations).
Additional indices applied to U.S. waters include biological indicators of estuarine
condition. (Fertig et al.. 2014) describe a Eutrophication Index applied to Barnegat
Bay-Little Egg Harbor Estuary, NJ using weighted indicators of water quality
(temperature, DO, TN, TP), light availability (chlorophyll a, total suspended solids,
Secchi depth, macroalgae percentage cover, percentage surface light, epiphyte biomass),
and seagrass (Zostera spp.) response (aboveground biomass, belowground biomass,
density, percentage cover and length). The biological condition gradient conceptual
framework (Davies and Jackson. 2006) was applied to Greenwich Bay, RI to assess
estuarine habitat overtime (Shumchenia etal.. 2015). Biological indicators included
seagrass extent, benthic community, primary productivity, and shellfish. In a regional
survey of 23 estuaries in southern California, 78% of estuaries were rated "moderate" or
"worse" based on macroalgae, 39% based on phytoplankton, and 63% based on DO using
the European Union Water Framework Directive (McLaughlin et al.. 2014). With the
ASSETS framework, 53% of surveyed areas were identified as impaired. The variability
in categorizing estuarine condition was influenced by spatial and temporal scales as well
as which indicators and thresholds were selected. Additional indices used in other
countries for describing estuarine condition are reported in Zaldivar et al. (2008). Boriaet
al. (2008). Boria et al. (2012). Devlin etal. (2011). Garmendia et al. (2012). and
Andersen et al. (2014).
A modeling indicator for ecosystem response to N uptake, which includes a measure of
the loss of species richness, was used to create a marine eutrophication Ecosystem
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Damage indicator (meED) (Cosmc et al.. 2017). The indicator is based on the loss of
species richness caused by hypoxia, which is in turn caused by eutrophication. The meED
ecosystem damage indicator itself indicates the potential impact of eutrophication in the
receiving habitat based on species density estimates. Cosme et al. (2017) support the
meED for inclusion in Life Cycle Impact Assessments (LCIA) when characterizing N
emissions. Their paper describes the calculation of the meED indicator for 66 large
marine ecosystems and maps risk of ecosystem damage due to eutrophication based on
this indicator.
10.3. Effects on Biodiversity
Increased N loading to coastal areas can lead to shifts in community composition,
reduced biodiversity, and mortality of biota. Biodiversity is important for ecosystem
stability and function, including provision of ecosystem services (Chapter 1.2.2.4).
Evidence for impacts to biodiversity include paleontological evidence (Appendix 10.3.1).
altered phytoplankton community composition (Appendix 10.3.2). responses of
phytoplankton to reduced versus oxidized N (Appendix 10.3.3). bacteria/archaea
diversity (Appendix 10.3.4). benthic diversity (Appendix 10.3.5). and fish diversity
(Appendix 10.3.6). The form of N supplied can significantly affect phytoplankton
community composition in estuarine and marine environments (Glibert et al.. 2016; Paerl
et al.. 2000; Stolte et al.. 1994). In hypoxic areas, mortality of benthic biota and
avoidance of I0W-O2 conditions by mobile organisms lead to changes in biodiversity and
loss of biomass (Diaz and Rosenberg. 2008). Energy transfer through the food web can
also be altered by a decrease in predators and an increase in microbes from oxygenated
areas to anoxic zones.
10.3.1. Paleontological Diversity
Sediment records from Chesapeake Bay showed alterations in producers and consumers
correlated to land use change in the watershed (Sowers and Brush. 2014; Brush. 2009). In
this estuary, diatom community structure has shown a steady decrease in overall diversity
since 1760 as estimated by sediment core analysis (Cooper and Brush. 1993). Diatom
populations began to shift from benthic to planktonic in the early 1900s corresponding to
the increased N flux and higher sediment inputs [which decreased light penetration;
(Brush. 2009)1. Additional sediment cores from two distinct locations in the estuary show
a shift to an estuarine food web that is predominately planktonic (Sowers and Brush.
2014). An increase in the abundance of the formaminifera Ammobaculites spp. and a
decrease in the abundance of the polychaete Nereis spp. were observed along with the
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change in primary producers. Paleontological evidence from England shows significant
changes in benthic species composition and ecosystem functioning between periods of
extremely low O2 (Caswell and Frid. 2013). The evidence suggests that normal benthic
functions are maintained during early hypoxia, but these functions collapse once
thresholds of severely low O2 are reached.
10.3.2. Phytoplankton Diversity
As reported in the 2008 ISA, excess N can contribute to changes in phytoplankton
species composition (U.S. EPA. 2008a'). High loadings of N and P can also increase the
potential for Si limitation, with associated changes in diatoms. Such changes to the
phytoplankton community and functional groups (e.g., diatoms, dinoflagellates,
cryptomonads, cyanobacteria, chlorophytes) can also affect higher trophic levels because
these organisms are the base of the estuarine food web. For example, if the nutrient mix
favors species that are not readily grazed (e.g., cyanobacteria, dinoflagellates), trophic
transfer will be poor, and relatively large amounts of unconsumed algal biomass will
settle to the bottom, which could stimulate decomposition, O2 consumption, and the
potential for hypoxia (Paerl et al.. 2003). Changes in phytoplankton species abundance
and diversity have been further documented through in situ bioassay experiments, such as
the results reported by (Paerl et al.. 2003) for the Neuse River estuary, North Carolina.
Effects were species specific and varied dramatically depending on whether, and in what
form, N was added. The findings illustrate the potential impacts of N additions on
phytoplankton.
Several studies published since the 2008 ISA use shifts in diatoms as a measure of N
enrichment effects on species diversity. In microcosm studies from Raritan Bay, centric
and chain-forming diatoms resulted from enriched NO3 concentrations, differing from
the pennate diatoms and green flagellates that accompanied ambient NO3 concentrations
(Rothenberger and Calomeni. 2016). The Si:N ratio was identified as an important factor
governing the phytoplankton community dynamics in the bay. Phytoplankton community
structure was altered by eutrophication in the Skidaway River estuary, Georgia, as
evidenced by a shift away from diatoms and towards nonsilicate nanoplankton (Verity
and Borkman. 2010). There was a strong relationship between diatom taxon distribution
and TN concentrations in Charlotte Harbor, FL (Nodine and Gaiser. 2014). Differences
were noted in the phytoplankton community structure during the time periods before,
during, and after eutrophication in the Black Sea (Mikaelvan et al.. 2013). Diatoms and
dinoflagellates especially were found to be dependent on high N and N:P ratios. While
nutrient ratios are known to influence phytoplankton community structure, Davidson et
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al. (2012) pointed out that these nutrient ratios are important only when at least one
nutrient is low enough to limit growth.
Shifts in phytoplankton community structure are known to occur in estuaries due to N
enrichment, but climatic changes may at times outweigh the impacts of eutrophication
I (Pacrl et al.. 2010); Appendix 10.1.4.11. Thus, physical changes caused by climate
change, such as water temperature, stratification, circulation, and hydrologic variability,
must be taken into account when modeling phytoplankton community responses to N
enrichment (Pacrl et al.. 2014). Likewise, Glibert et al. (2014) noted changes in
phytoplankton community composition along with increased anthropogenic N input and a
switch to long-term wet conditions in the coastal lagoons area of Maryland and Virginia.
A 3-year data set from Raritan Bay indicates that both climatic conditions and nutrient
concentrations affected phytoplankton composition from 2010-2012 (Rothcnbcrgcr et al..
2014). The abundance of flagellates in the bay over time reflects a shift in dominance
away from diatoms that is characteristic of eutrophic systems. Location and surrounding
land cover, as well as the type of N input, are important factors in the response of the
phytoplankton community to N addition (Reed et al.. 2016). Phytoplankton community
composition in four tidally influenced sites along a gradient from highly developed to
undeveloped (natural) in coastal SC were influenced by surrounding land cover and the
form of N, with cyanobacteria, dinoflagellates, and other potentially HAB-forming
species most often found at the more developed sites.
10.3.3. Diversity of Phytoplankton Under Different Forms of Nitrogen (Reduced
vs. Oxidized)
Reduced forms of atmospheric N are increasing relative to oxidized N in the U.S.
(Appendix 10.1.2) and play an increasingly important role in eutrophication and HAB
dynamics in coastal areas. Specific phytoplankton functional groups have a preference for
NH4+ over NOs (Appendix 10.2.2). perhaps leading to selective stimulation of primary
production, especially in light-limited estuarine and coastal waters (Glibert et al.. 2016;
Paerl et al.. 2000; Stolte et al.. 1994). Increasing loads of NH3+/NH4+ have been linked to
the expansion of HABs (Glibert et al.. 2016; Esparza et al.. 2014; Altman and Paerl.
2012; Blomqvist et al.. 1994). For example, in San Francisco Bay, a comparison of N
isotopes in cells of the cyanobacterial HAB species Microcystis aeruginosa with N in
rivers flowing into the bay indicated the primary source of N that supported the bloom
formation was likely NH4+ (Lehman et al.. 2015). Phytoplankton response to the form of
N supplied (Appendix 10.2.2. and Table 10-3) could lead to altered phytoplankton
growth and community composition and have cascading effects on trophic structure and
biogeochemical cycling (Glibert et al.. 2016; Paerl et al.. 2000). Cyanobacteria, and many
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chlorophytes and dinoflagellates may be better adapted to the use of NH/, while diatoms
generally thrive in environments with oxidized forms of N such as NO3 . Figure 10-7
summarizes the dominant functional phytoplankton groups in primarily reduced N and in
primarily oxidized N conditions.
Field studies indicate that HAB-forming dinoflagellates are often associated with
enrichment of reduced N. In Maryland and Virginia coastal bays, Glibert et al. (2014)
observed a regional shift in phytoplankton community composition to species that do
well in reduced N conditions. Phytoplankton community dynamics varied with the form
of N in nutrient enrichment experiments using water from the New River estuary, NC
(Altman and Paerl. 2012). Photopigment analysis used to identify and quantify taxonomic
groups revealed that the addition of riverine dissolved organic nitrogen (DON) promoted
dinoflagellates, chloropytes, and myxoxanthophyll (cyanobacteria) while zeaxanthin
(cyanobacteria) was most frequently detected with inorganic N. In the Neuse River
estuary, NC where NH44" concentrations have increased over time, the abundance of
aphidophytes (including the potentially toxic dinoflagellate species H. akashiwo),
haptophytes, chlorophytes, and the bloom-forming dinoflagellate Heterocapsa rotundata
have also increased (Rothenberger et al.. 2009). Incubation experiments carried out in the
Neuse River estuary in spring, early summer, and late summer to test how the growth and
community composition of phytoplankton responded to N availability and specific form
of N (Cira et al.. 2016). Phytoplankton community composition varied by season, and
different taxa were found to be able to uptake both urea and NO3 when conditions were
limiting, although the results varied by taxa. Seasonal experiments suggested that N was
not the only factor controlling phytoplankton growth in the spring, as the experiment
conducted in March showed that only growth rates of fucoxanthin-containing (brown
algae) species were stimulated by N addition. In field studies along a gradient from
highly developed to undeveloped (natural) study sites in the southeastern U.S.,
phytoplankton communities at the more developed sites showed higher biomass and
growth rates with N (particularly urea) additions (Reed et al.. 2016). In addition,
cyanobacteria, dinoflagellates, and other potentially HAB-forming species were more
often found at the more developed sites.
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diatoms
crypto-
phytes
dirio
flagellates
pico-
tyanobacteria
S	I
5=|

C. u	•Ł*
M_	J2
—j O	'."1J
5 urn


Source: Gilbert et al. (2016).
Figure 10-7 Summary conceptual schematic illustrating the effect of changes
in the proportion of Nm+ and NO3" in the loads of N provided to a
natural system. When NtVis the dominant form, and when waters
are warmer, flagellates, cyanobacteria, and chlorophytes among
other classes may proliferate, leading to overall productivity
dominated by the small size class of algae (e.g., <5 |jm). In
contrast, when NOs" is the dominant form provided, especially
under cooler water conditions, diatoms more likely dominate, and
their overall production will be more likely dominated by cells of a
larger size class (e.g., >5 pm). Moreover, chlorophyll a yield and
total production may be higher than under the NH4+enrichment
condition.
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In mesocosm studies in the P-limited Florida Bay ecosystem, chl a concentrations
(phytoplankton biomass) increased with N + P treatment but did not increase with only N
enrichment. However, N only enrichment (especially in the form of NH44") did change the
phytoplankton assemblage in the direction of more picocyanobacteria (Shangguan ct al..
2017). A shift of phytoplankton biomass toward diatoms making up a larger portion of
the total biomass was observed when N in the form of NO3 was added, either alone or in
combination with P. Using chlorophytes and diatoms from Suisun Bay, CA grown under
controlled laboratory conditions Berg et al. (2017) found that the four diatoms species
tested grew faster on NH44" than NOs suggesting that diatoms were not under a
competitive disadvantage under high NH44". Only the chlorophyte R. planktonicus grew
significantly faster on NO3 . None of the diatoms tested in controlled lab conditions were
sensitive to NH44" at concentrations that would be found in the local environment. The
results show that the responses are highly species specific in tolerance of NH4+.
Not all studies have found variation in algal response with the form of N. Richardson et
al. (2001) examined the effects of different forms of N application (NO3 . NH4+, urea) on
the structure and function of estuarine phytoplankton communities in mesocosm
experiments in the Neuse River estuary, NC. Even though NH4+ is more readily taken up
by phytoplankton in this estuary than is NO3 (Twomev et al.. 2005). the results of the
Richardson et al. (2001) study suggested that phytoplankton community structure was
determined more by the hydrodynamics of the system than by the form of N available for
growth. Twomev et al. (2005) measured Neuse River estuary phytoplankton uptake rates
of NH4+, NO3 . and urea. NH44" was the dominant form of N taken up, contributing about
half of the total N uptake throughout the estuary. Uptake varied spatially; in particular,
N03 uptake declined from 33% of the total uptake in the upper estuary to 11 and 16% in
the middle and lower estuary, respectively. Urea uptake contributed 45 and 37% of the
total N uptake in the middle and lower estuary, showing the importance of regenerated N
for fueling phytoplankton productivity in the mesohaline sections of the estuary.
Therefore, N budgets based only on inorganic forms may seriously underestimate the
total phytoplankton uptake (Twomev et al.. 2005).
10.3.4. Diversity of Bacteria, Archaea, and Microzooplankton
Since the 2008 ISA, a few studies have explored ammonia-oxidizing bacteria (AOB) and
ammonia-oxidizing archaea (AOA) community responses in relation to N in coastal
waters. These organisms play a key role in N cycling and the AOA were described
relatively recently (Konneke et al.. 2005). Bouskill et al. (2012) examined the distribution
and abundance of AOA and AOB across a variety of aquatic environments, including two
sites at different salinity levels in Chesapeake Bay. The AOB were dominate in the
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freshwater end of the continuum, whereas AOA were more common in open ocean areas.
The abundance of both groups were correlated with the concentration of NH4+ in the
water, especially relative abundance of AOB. A similar pattern was observed in the
Sacramento-San Joaquin Delta where AOB were more abundant in the NH4+-rich
Sacramento River compared to the San Joaquin River and Suisun Bay where AOA were
prevalent (Damashck et al.. 2015). Community structure of benthic ammonia oxidizers
differed across the region and appeared to be related to nutrient inputs. In the Puget
Sound estuary, WA, AOA were generally more abundant, however, AOB increased
relative to AOA during periods of high NH4+ (Urakawa et al.. 2014). Community
structure of plantonic ciliates was found to be significantly related to the spatial
distribution of NO;, -N concentrations across a gradient of sites in Jiaozhou Bay, China,
suggesting the potential applicability of plantonic ciliate functional groups as a water
quality indicator (Xu et al.. 2016a).
10.3.5. Benthic Diversity
Invertebrates associated with bottom substrates of coastal systems are useful indicators of
biological change for impacted waters. Several indices based on benthic invertebrates,
macroinvertebrates, or macrophytes have been used to assess biological condition in
response to the European Water Framework Directive (Andersen et al.. 2014; Zaldivar et
al.. 2008). A few of these indices have been tested in U.S. waters. Boriaetal. (2008)
compared the Benthic Index of Biotic Integrity (B-IBI), the AZTI Marine Biotic Index
(AMBI) and its multivariate extension the M-AMBI in Chesapeake Bay. There was
relative agreement between methodologies, and differences were related to spatial
variability and habitat type. Comparable regional benthic indices for the
Northeast/Acadian, northeast Virginian, southeast Virginian, southeast Carolinian, and
Gulf Louisianan coasts were developed for the U.S. EPA's coastal assessments (NCCA)
(U.S. EPA. 2016g).
N enrichment has been shown to have significant but complex effects on
suspension-feeding macroinvertebrates in estuaries. Shellfish species can respond
positively to increased N loading and high phytoplankton biomass levels due to increased
quantity and quality of food particles (Carmichael et al.. 2004). A recent study in Long
Island's Peconic Estuary suggested that eutrophication affects shellfish through changes
in the quality of food and not the quantity (Wall et al.. 2013). Select species (oysters [C.
virginica] and clams [Mercenaria mercenaria]) in eutrophic areas experienced enhanced
growth rates that were strongly correlated with high densities of autotrophic
nanoflagellates and centric diatoms. Other species (scallops [Argopecten irradians| and
slipper limpets) suffered negative effects and grew at the slowest rate at the most
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eutrophic sites. Bay scallop growth was negatively correlated with densities of
dinoflagellates, which were more abundant at the most eutrophic site (p < 0.001). In
rocky shore environments in Ireland, similar to those of the U.S. northeastern Atlantic
coast, N enriched sites near industrial and sewage outfalls have been shown to have
reduced total abundance and number of molluscan taxa and altered community
composition compared with control sites (Atalah and Crowe. 2012). Nitrite and NH3
levels explained the highest percentages of variation in the molluscan assemblage
structure (13 and 12.5%, respectively), owing in large part to the absence of three rare
species at contaminated sites. The observed shift in community structure caused by
different levels of molluscan tolerance to N enriched conditions has been suggested for
use as a biological indicator of eutrophication (Atalah and Crowe. 2012). The role of
benthic invertebrates in coastal N cycling and the use of shellfish for coastal N
remediation is discussed in Appendix 7.2.6.11.
10.3.6. Fish Diversity
A few studies have recently reported effects on fish biodiversity in nutrient-impacted
estuaries. Comparison of trophic organization of fish in estuarine reaches of nine rivers in
Victoria, Australia showed that inorganic N loading was an important factor explaining
trophic diversity of fish assemblages. There was greater trophic diversity of fish
assemblages with mid to high inorganic N loading suggesting that DIN influences not
only estuary primary productivity but is transmitted upward through the food chain
(Warn etal.. 2016). In Prince Edward Island estuaries, adult and young-of-the-year
abundance of pollution-tolerant mummichogs (Fundulus heteroclitus) increased in
estuaries with greater intensity of N loading from agricultural land use (Finlcv ct al..
2013). In this study, eutrophication was associated with changes in habitat variables,
which were linked to mummichog abundance. In an estuary in the southern Gulf of St.
Lawrence in Canada, loss of eelgrass was linked to declines in fish diversity, but the loss
did not change the positions of organisms within the food chain (Schein et al.. 2013;
Schein et al.. 2012).
10.3.7. Trophic Interactions
Altered trophic transfer following nutrient-associated changes in phytoplankton
community composition were reported in the 2008 ISA. Phytoplankton that are not
readily grazed, such as cyanobacteria and dinoflagellates, are not transferred to higher
trophic levels as efficiently as diatoms [more readily grazed by zooplankton which are
then consumed by fish, (Pacrl et al.. 2003)1. Since the 2008 ISA, studies have further
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characterized the effects of nutrient enrichment on trophic interactions. In the Skidaway
River estuary, GA, increasing nutrient concentrations and changes in nutrient ratios led to
increases in eukaryotic organisms (heterotrophic and mixotrophic flagellates,
dinoflagellates, ciliates, and metazoan zooplankton) but not diatoms, potentially altering
food web predator-prey relationships (Verity and Borkman. 2010). In an experimental
study in Chesapeake Bay, Reynolds et al. (2014) demonstrated that a reduction of
crustacean grazers controlling algal growth on seagrasses resulted in 65% reduced
seagrass biomass and nearly sixfold increased epiphytic algae. In experimental plots
where grazers were removed, aboveground eelgrass biomass was reduced following
fertilization. In the presence of grazers, eelgrass biomass increased with fertilization.
When predators were excluded, mesograzers were able to limit ephiphyte growth even
when nutrients were added. Based on these findings, the researchers suggested that it is
unlikely that reducing nutrient pollution alone would restore seagrass meadows where
alterations to food webs have already reduced populations of algae-feeding mesograzers.
Similar results were reported in McSkimming et al. (2015) from Posidonia angustifolia
meadows in Australia where mesograzers responded to nutrient addition by increasing
grazing per capita, resulting in greater consumption of epiphytic algae.
A review of nutrient addition studies from the North Atlantic suggests that, overall, the
addition of nutrients caused the biomass of opportunistic macroalgae to approximately
double. Higher numbers of midlevel predators has the same effect because they lead to a
smaller population of grazers eating the macroalgae (Ostman et al.. 2016). Analysis of
review data indicated that the effect of midlevel predators on green macroalgae biomass
increased with eutrophication. Recovery of sea otter (Enhydra lutris) populations in
California appear to control algal growth on eelgrass by increased predation of crabs
(Cancer spp. and Pugettia producta) which, in turn, decrease predation of mesograzers
such as sea slugs (Phyllaplysia taylori) that feed on the algae (Hughes et al.. 2013).
Newer literature has provided evidence for complex interactions between eutrophication
and other stressors and subsequent effects on trophic interactions. For example, Burnell et
al. (2013) used mesocosms to demonstrate the effect of eutrophic conditions on
plant-herbivore grazing interactions between the sea urchin Amblypneustes pallidas and
the seagrass Amphibolis antarctica affected by warming and acidification. Nutrient
enrichment offset increased grazing associated with acidification and warming; however,
it did not fully counter the additive effects of these two stressors.
10.3.8. Models Linking Indicators to Nitrogen Enrichment
Process-oriented models such as nutrient-phytoplankton-zooplankton (NPZ) models are
used to predict the response of organisms such as HABs to various known and/or
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predicted nutrient conditions (Swancv et al.. 2008). A nutrient-driven phytoplankton
model developed by Scavia and Liu (2009) was expanded by Evans and Scavia (2013) to
test the sensitivity of the response of chlorophyll a levels and DO levels to N enrichment.
Results indicated that separate processes control chlorophyll a and DO sensitivity
(estuary flushing time and relative mixing depth, respectively), and that these sensitivities
vary among estuaries (Figure 10-8).
Using the European regional Seas Ecosystem Model for an area of the Japan Sea, Lee and
Yoo (2016) showed that the phytoplankton community there will shift to smaller
phytoplankton, and the food web structure will likely be altered if atmospheric deposition
continues to intensify in the region as predicted. For the study period 2001-2012, the
model estimated that the atmospheric N deposition in the Ulleung Basin would increase
the annual net primary production by an average of 4.58% (range from 3.77-10.58%).
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"—•—Blue Hill B
— O— Patuxent R
. —*— Penobscot B
—O— Potomac R
—X— St, Marys R and Cumberland S
6
u 10
Chi = chlorophyll; DO = dissolved oxygen; TN = total nitrogen.
The x-axis shows river total nitrogen load relative to the 1992 SPARROW total nitrogen load for each estuary. In the figure legend,
R = river, B = bay, and S = sound.
Source: Evans and Scavia (2013).
Figure 10-8 Forecasting curves for effects on total nitrogen loadings on
(A) chlorophyll and (B) dissolved oxygen (mean and 95%
confidence interval) for selected estuaries demonstrating the full
range of sensitivity to relative total nitrogen loading.
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10.4. Animal Behavior and Disease
In addition to changes in biological indicators (i.e., chlorophyll a, HABs, macroalgal
abundance, SAV) and altered species diversity there is increasing evidence for a role of N
in behavior and disease in coastal biota. These indirect responses may affect the fitness of
organisms inhabiting nutrient-enriched waters.
10.4.1. Behavior
Coastal biota exhibit behavioral responses to hypoxic conditions, including acclimation
or avoidancc(Townhi11 et al.. 2017; Levin et al.. 2009). Some fish have the ability for
aquatic surface respiration, a behavior in which the fish swims to the surface and is able
to take advantage of the relatively higher levels of DO in the water there. Dixon et al.
(2017) assessed four common species of shallow estuarine finfish for their reactions to
episodic hypoxic conditions and altered pH levels. Severe hypoxia induced surface
respiration behavior of three of the four fish species, but all were unaffected by diel
variations in pH.
Turbidity caused by excess algae in the water, a secondary effect of eutrophication, has
been shown in laboratory studies to potentially alter some fish behaviors, with
implications for mate selection and offspring survival rates. It is not known whether some
of these behavior changes will be adaptive or if they will have positive or negative effects
on the fish populations. For instance, when the strength of sexual selection on several
traits is relaxed, the relative importance of survival selection may increase (Candolin.
2009). The color contrast created by increased turbidity may lead to a heightened feeding
response in some fish larvae such as California yellowtail (Seriola lalandi), which
exhibited elevated growth and survival rates compared to larvae raised in clear
(oligotrophic) water (Stuart and Drawbridge. 2011).
Several studies of the three-spined sticklebacks (Gasterosteus aculeatus) indicate that
observed alterations of reproductive behaviors are linked to eutrophic conditions
(Figure 10-9). Laboratory manipulations suggest that turbidity caused by eutrophication
may affect mate selection and breeding success in this euryhaline fish species (Heuschele
et al.. 2012; Heuschele and Candolin. 2010). When given the choice between dense or
sparse algae, male sticklebacks preferred to nest in dense algae (which simulated
eutrophic conditions). The nests of those males were more likely to be parasitized by
other males; however, the males nesting in dense algae also acquired more eggs, and
thus, had a higher probability of reproduction (Heuschele and Candolin. 2010). This may
be because reduced visibility leads to reduced mate selectivity searching by females,
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although it is unknown whether these behavior changes would have positive or negative
impacts on the population (Hcuschclc et al.. 2012).
Eutrophication
-	filamentous algae
-	water turbidity
Reproductive
success
I
Sexual selection
1
Population
demography
1
Survival selection
Source: Candoliri (2009).
Figure 10-9 The pathway of effects of eutrophication on different reproductive
behaviors and selection forces in Gasterosteus aculeatus.
3	Simulated turbidity has also been shown to impair visual mate choice in an eastern
4	Atlantic species of marine pipefish, Syngnathns typhis (Sundin et al.. 2010). A follow-up
5	study in the same species allowed other factors to occur, namely mating competition and
6	mate encounter rates which counteracted the effects of lack of visual cues and led to
7	increased sexual selection in turbid waters (Sundin et al.. 2017). In the same species, the
8	latency period between courting and copulation was prolonged in I0W-O2 conditions,
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although other measures of reproductive behavior (total time spent courting, probability
of mating, dancing, male pouch-flap behavior) were unaffected (Sundin et al.. 2015).
Other types of fish mating behaviors are also affected by eutrophication. Male
sticklebacks engage in an opportunistic mating behavior called "sneaking" which is less
successful under more turbid (eutrophic) conditions (Vlicgcr and Candolin. 2009).
Eutrophication has so far been found to have conflicting impacts on nest building
behavior in male sticklebacks, with varying implications for the survival rate of
stickleback offspring. Male sticklebacks built smaller nests with wider openings during
eutrophic (algal bloom) conditions (Wong et al.. 2012). These males also took longer to
complete the nest than did males that built during nonbloom conditions. However, under
normal water conditions in a laboratory setting, male sticklebacks that had been collected
from eutrophic waters built their nests faster than those collected from noneutrophic
waters, although the nests did not differ in structure or composition between the two
groups (Tuomaincn and Candolin. 2013V In European sand gobies (Pomatoschistus
minutus), another species that exhibits paternal care of eggs, algal turbidity altered male
behaviors (Jarvenpaa and Lindstrom. 201IV Males spent more time away from the nest,
and exhibited less fanning of the eggs with their pectoral and tail fins. Goby egg survival
rates were actually found to be higher under turbid conditions. Although the mechanism
for this outcome is unknown, it was proposed that males may have increased other forms
of care in the presence of eutrophic conditions, leading to higher egg survival.
Under normal conditions, it is beneficial for sticklebacks to live in larger groups where
they gain protection from predators, are able to forage more efficiently, and are able to
find a mate more easily than in smaller groups. However, in eutrophic (turbid) water,
sticklebacks did not show a preference for joining a larger shoal as they did under normal
conditions (Fischer and Frommen. 2013). This lack of exhibited shoal preference under
turbid conditions could eventually reduce individual stickleback fitness. In addition,
sticklebacks changed shoals less frequently in turbid water than observed under normal
conditions, leading to a decrease in the transfer of important social information, such as
the location of foraging routes. Therefore, increased turbidity may reduce social learning
opportunities for sticklebacks and may, in turn, impair their foraging efficiency (Fischer
and Frommen. 2013).
10.4.2. Disease
Diel-cycling hypoxia in estuaries is a natural phenomenon that can be exacerbated by
increased biomass and productivity associated with anthropogenic nutrient loadings
(Tyler et al.. 2009). Tyler et al. (2009) observed that upper areas of four Delaware
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estuaries experienced increased incidence of severe hypoxia (DO <2 mg/L) where
nutrient loading was greater. The duration and spatial extent of diel-cycling hypoxia in
the estuaries were primarily associated with water temperature, previous day's insolation,
hours of morning ebb tide, and daily stream flow. Oysters (C. virginica) exposed to
periods of diel-cycling hypoxia were demonstrated to have increased incidence and
progression of Perkinsus marinus infection (Dermo) in field and laboratory experiments
(Brcitburg et al.. 2015). Field experiments were conducted in Chesapeake Bay. The
authors suggest that the likely mechanism is negative effects of hypoxia on host immune
response.
Seagrass meadows were recently shown to decrease pathogens harmful to both human
health and marine organisms. In field studies adjacent to islands in Indonesia affected by
wastewater pollution, bacterial pathogens were reduced by up to 50% in seagrass
meadows and coral reefs adjacent to seagrass beds had reduced disease incidence (Lamb
et al.. 2017). Several recent studies have suggested that seagrasses may be more
susceptible to wasting disease caused by the marine slide mold (genus Labyrinthula)
under conditions of elevated NOs loading (Hughes et al.. 2017; Kaldv etal.. 2017).
Nutrient enrichment is one of several factors linked to increased disease susceptibility
and bleaching in corals (D'Angclo and Wiedenmann. 2014; Vega Thurber et al.. 2014;
Wiedenmann et al.. 2013). Coral bleaching occurs when the symbiotic relationship
between the coral host and microalgae (Zooxanthellae spp.) breaks down. Although most
research on nutrient loading on corals includes both N and P several studies have isolated
effects of N which impacts corals via distinct pathways from P (Shantz and Burkcpilc.
2014). Wiedenmann et al. (2013) reported that the coral-algae symbiosis can be
interrupted by elevated inorganic N concentration rather than both N and P. In a
metanalysis of nutrient impacts on corals, N reduced coral calcification 11% on average,
while increasing photosynthetic rate (Shantz and Burkepile. 2014). Increased DIN
decreases the temperature threshold at which coral bleaching occurs (Wooldridge. 2009).
The threatened status of staghorn coral (Acropora cervicornis) and elkhorn coral
(Acroporapalmata) under the U.S. Endangered Species Act has been linked to indirect N
pollution effects, specifically low DO and algal blooms that alter habitat, and other
non-nutrient stressors (Hernandez et al.. 2016).
10.5. Nutrient Enhanced Coastal Acidification
Since the 2008 ISA, several studies have suggested that the increased respiration caused
by N enrichment may exacerbate coastal ocean acidification through alteration of the
carbon cycle (Appendix 7.2.4). Dissolution of atmospheric anthropogenic carbon dioxide
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(CO2) into the ocean has caused increasing acidification of coastal waters, resulting in
long-term decreases in pH (Wallace et al.. 2014; Orr et al.. 2005). N enrichment is
expected to worsen this acidification because degradation of excess organic matter then
produces CO2 in the water column, which in turn can make the water more acidic
I Figure 10-10; (Wallace et al.. 2014; Sunda and Cai. 2012; Cai et al.. 2011c; Howarth et
al.. 2011)1. An important additional source of organic matter leading to overall declines
in pH is potentially allochthonous organic matter inputs which have been increasing in
many coastal watersheds from changing land use (Wilson et al.. 2016; Wetz and
Yoskowitz. 2013). Both N-driven eutrophication and anthropogenically enhanced
allochthonous organic matter loading operate simultaneously.
In additional to microbial degradation of organic matter, respiration of living algae and
seagrasses during the night can also drive acidification. In an N-enriched seagrass
ecosystem in Cape Cod, a very pronounced diel pattern in pH was observed with highest
acidity at dawn (Howarth et al.. 2014). The pH varied from below 7.0 at dawn to near
seawater levels of 8.1 by sunset. Acidification also can be enhanced indirectly as shown
in Chesapeake Bay through the creation of anoxic waters and oxidation of hydrogen
sulfide (Cai et al.. 2017b).
Ocean acidification, which can be exacerbated by elevated N input, is projected to alter
marine habitat, have a wide range of effects at the population and community level and
affect food web processes (Mostofa et al.. 2016; Andersson et al.. 2015; Gavlord et al..
2015; Sunda and Cai. 2012; Donev et al.. 2009). A major finding of The West Coast
Ocean Acidification and Hypoxia Science Panel was that while the dominant cause of
ocean acidification in the region is global CO2 emissions, local discharge of nutrients and
organic C is a factor exacerbating ocean acidification and hypoxia (Chan et al.. 2016).
Nutrient additions can trigger algal blooms, which upon decomposition, further reduce
oxygen levels in the water column and thus also lower the pH.
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Increasing
atmospheric
CO,
Loading
Algal blooms
from nutrient
inputs ,
Adverse
biological
impacts
/ Increasing \
[COz]/lower pH
as regulated by
^temperature and
salinity S
Microbial \
respiration of
organic
v matter /
Lower 02/
hypoxia
C02 = carbon dioxide; N = nitrogen; 02 = oxygen.
Modified from: Sunda and Cai (2012).
Figure 10-10 Pathway of nutrient-enhanced coastal acidification from nitrogen
loading to biological effects. Both microbial respiration of organic
matter and increasing atmospheric carbon dioxide lower pH of
coastal waters.
1	Acidification of coastal waters may cause varying degrees of harm to marine organisms
2	that produce calcium carbonate shells or skeletons, including oysters, clams, sea urchins,
3	shallow water corals, and calcareous plankton (Mostofa et al.. 2016; Gledhill et al.. 2015;
4	Pfister et al... 2014; Kroekcr et al.. 2013; Sunda and Cai. 2012). The acidifying process
5	decreases the saturation state of the two minerals (aragonite and calcite) that most
6	bivalves use to form their shells (Barton et al.. 2015; Barton et al.. 2012). Lower pH of
7	seawater can also potentially impair physiological energetics and spawning capacity of
8	shellfish (Xu et al.. 2016b). Commercially important species from the New England
9	coastal region have shown biological responses to changes in the carbonate system
10	associated with ocean and coastal acidification (Gledhill et al.. 2015). Already, there are
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declines in oyster production on the U.S. West Coast due to the inability of oysters to
create shells due to acidification (Chan et al.. 2016; Barton et al.. 2015; Hettinger ct al..
2012). However, other factors can confound the response of these sensitive organisms.
More research is needed to accurately determine the interactions and combined impact of
ocean acidification and N enrichment and other stressors on U.S. ecosystems like coral
reefs and estuaries with sensitive shellfish populations (Kroekeretal.. 2013).
In a recent modeling study of risk of food web and fisheries to ocean acidification in the
California current nearshore, invertebrates and associated fishery revenues are expected
to be impacted to a greater extent than pelagic species (Marshall et al.. 2017). Although
this modeling study did not take N inputs or other measures into account, invertebrates
that live in or on the benthic layer (crabs, shrimps, bivalves) will experience the strongest
effects of ocean acidification, followed by those species that consume benthic
invertebrates such as demersal fish and Dungeness crabs (Metacarcinus magister).
However, the model also showed that different species of groundfish reacted differently
to the changing pH, so some species and associated fishery dynamics may be impacted to
a greater degree than others.
Research on costressors associated with conditions of coastal acidification and
eutrophication suggest that interactions between elevated CO2, decreasing pH, and
nutrient inputs are complex. In macroalage Ulva pertusa grown under conditions of low
pH and high NH44", both growth rate and NH4+ uptake were significantly higher than in
treatments with higher pH and lower nutrient enrichment (Kang and Chung. 2017).
Young and Gobler (2016) tested the effects of elevated concentrations of CO2 alone, and
in combination with elevated nutrient (N + P) levels, on the growth rates of two common
species of opportunistic macroalgae, the rhodophyte Gracilaria and the chlorophyte
Ulva. Results showed that higher levels of pCOi significantly enhanced the growth rates
of both types of macroalgae, and the combination of enriched N + P and pCOi did appear
capable of synergistically promoting the growth of Ulva. Given that eutrophication can
yield elevated levels of pCOi. this study suggests that the overgrowth of macroalgae in
eutrophic estuaries can be directly promoted by acidification.
Eelgrass productivity and growth are predicted to increase under conditions of elevated
CO2 because photosynthesis in these plants is thought to be currently limited by CO2
(Koch et al.. 2013; Alexandre et al.. 2012; Palacios and Zimmerman. 2007). However,
seagrass productivity responses to combined stressors ofpC02 and NO3 were found to
be variable between different species (Ow et al.. 2016). In mesocosm experiments with
the seagrass Zostera noltii, photosynthesis increased under enriched CO2, but NO3 and
NH4+ uptake and growth rate were not significantly affected, indicating that seagrass
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production may be limited by N availability under increased CO2 (Alexandre ct al..
2012V
Combined effects of NO3 enrichment, acidification, and temperature resulted in the
highest reproductive success for the green tide-forming algae Ulva rigida, pointing to
potentially more severe green tides under future scenarios than under present conditions,
especially in areas where eutrophication is a concern (Gao et al.. 2017). A decline in
calcification of the coral Marginopora rossi was observed at a pH of 7.6 alone or in
combination with nutrients (Rcvmond et al.. 2013). In NO3 enriched seawater, the
number of zooxanthellae cells in the host coral increased but with no observable benefits
on coral growth. A threshold for ocean acidification at pH 7.6 alone or in combination
with eutrophication will lead to a decline inM rossi calcification.
Models show that while the impact of each acidification pathway (N enrichment and
atmospheric CO2 dissolution) may be moderate, the combined effect of the two may be
much larger than would be expected by just adding the effects of each pathway together
(Sunda and Cai. 2012; Cai et al.. 2011c). These models, which incorporate the expected
future higher atmospheric CO2 levels and current levels of eutrophication, show dramatic
increases in acidification of coastal waters below the pycnocline. Model predictions
agreed well with pH data from hypoxic zones in the northern Gulf of Mexico and both
the modeled and the measured decreases in pH are well within the range shown to impact
marine fauna (Sunda and Cai. 2012). Data from the northern Gulf of Mexico revealed a
significant positive correlation between subsurface water pH and O2 concentration,
linking acidification to low O2 levels from organic matter decomposition (Cai et al..
2011c). Results from a coupled physical-biogeochemical model in the Gulf of Mexico
point to eutrophication driven by river nutrient input as an important factor in
acidification and hypoxia generation (Laurent et al.. 2017). Findings from the model
showed that acidified waters are predicted in a thin layer close to the bottom of the Gulf,
similar to the vertical distribution of the seasonal hypoxia zone.
The increase in atmospheric CO2 dissolution is also expected to alter the N cycle in the
ocean, with implications for the entire ecosystem. Acidification will result in decreased
rates of nitrification, which combined with expected increasing N deposition, is expected
to cause the NH44" concentration of the water to rise (Lefebvre et al.. 2012). A study of the
coccolithophore Emiliania huxleyi, which plays a major role in the global carbon cycle by
regulating the exchange of CO2 across the ocean-atmosphere interface, found that higher
levels ofNH4+ can affect the morphology and calcification of the coccolithophore. The
combined effect of higher NH4+ levels and greater acidification could reduce calcification
rates, suggesting a greater alteration to the ocean's carbon cycle due to the combined
effect of increased NH4 /NO3 and acidification (Lefebvre et al.. 2012).
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10.6.
Summary of Thresholds and Levels of Deposition at Which
Effects Are Manifested
Studies linking changes in estuary nutrient status to atmospheric deposition have been
limited, although some states are addressing atmospheric inputs as part of their
development of Total Maximum Daily Loads (TMDLs) (Linker et al.. 2013). Since the
2008 ISA, additional thresholds of response to N have been identified for biological
indicators. In general, the identification and application of indicators and thresholds to
assess eutrophication in coastal areas is more highly developed in Europe than in the U.S.
due to the implementation of the EU's Water Framework Directive (Zaldivar et al..
2008). Here we limit discussion to indicators and thresholds that have been applied to
U.S. water bodies.
For chlorophyll a, both NEEA ASSETS and U.S. EPA NCCA have a similar range for
categorizing eutrophic conditions; however, ASSETS uses the 90th percentile chlorophyll
concentration of annual data and U.S. EPA NCCA uses growing season (June-October)
values [(Borjaetal.. 2012); Table 10-21. For ASSETS, chlorophyll concentrations of
0-5 |ig/L is a water body with low risk of eutrophication, 5-20 |ig/L is moderate,
>20 |ig/L is high with >60 |ig/L indicating hypereutrophic conditions. NCCA identified
chlorophyll concentration of >20 |ig/L as an indicator of an estuary in poor condition.
Chlorophyll concentrations between 5-20 |ig/L are classified as fair and chlorophyll
concentration 0-5 |ig/L was good for sites located in the Northeast, Southeast, Gulf, West
Coast, and Alaska (U.S. EPA. 2012b). Values were lower for sites in Hawaii, Puerto
Rico, U.S. Virgin Islands, American Samoa, and Florida Bay (0.5 (ig/L good, 0.5-1 (ig/L
fair, >1 |ig/L. poor).
For DO, ASSETS uses the 10th percentile of annual data. DO concentrations of 0 mg/L
are anoxic, 0-2 mg/L are indicative of hypoxic conditions, and 2-5 mg/L are biologically
stressful conditions (Devlin et al.. 2011; Bricker et al.. 2007). Spatial coverage (range of
0-10% [low] to >50% [high]) and frequency of occurrence (persistent, periodic,
episodic) are also included in determining reference thresholds for DO with ASSETS.
For U.S. EPA NCCA, the cutpoint used for poor DO condition is <2 mg/L in bottom
waters (U.S. EPA. 2016g. 2012b). Because many states use higher concentrations, the
NCCA considers concentrations between 2 and 5 mg/L as fair, and >5 mg/L as good.
Under the U.S. Clean Water Act, states are in the process of developing numeric nutrient
criteria to better define levels of N and P that affect estuaries and coastal marine waters
(Appendix 7.2.10). The numeric values include both causative (N and P) and response
(chlorophyll a, biocriteria) variables to assess eutrophic conditions. For progress toward
state development of numeric criteria for N see http://cfpub.epa.gov/wqsits/nnc-
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development/. Atmospheric deposition as a source of N may represent a potential target
for remediation such as TMDL, nutrient budgets, and allocations.
The Chesapeake Bay 2010 TMDL load reductions include atmospheric deposition. This
is the first time atmospheric N loads to tidal waters have been included in a TMDL for
reduction (U.S. EPA. 2010). For Chesapeake Bay, a numerical chlorophyll criteria
ranging from 1.4 to 15 mg/m3 (|ig/L) based on seasonal mean across salinity zones was
proposed as a restoration goal with 90th percentile compliance limits ranging from 4.3 to
45 mg/m3 [fig/L; (Harding et al.. 2014)1. These values were based on relationships of
chlorophyll a concentrations to DO, water clarity, and presence of toxic algae.
Chlorophyll a concentration ranges of 7.2 to 11 mg/m3 (|ig/L: May-August mean) in the
bay and 9 to 14 mg/m3 (|ig/L: annual mean) in the tributaries would be protective of low
DO conditions. For SAV growth, a mean chlorophyll a concentration of 7.9-12 mg/m3
(fig/L) with a threshold of 19-28 mg/m3 (|ig/L) during the growing season would allow
for sufficient light conditions. For decreased risk for toxic cyanobacteria, a mean summer
chlorophyll a concentration of 15 mg/m3 and a threshold of 25 mg/m3 (|ig/L) was
recommended.
For seagrasses in New England estuaries, a threshold of N loading (N export of
wastewater, fertilizer, and atmospheric deposition from the watershed combined with
direct atmospheric deposition to waters) >100 kg N/ha/yr was identified. Eelgrass is
essentially absent in areas with these levels of N loading, and levels above 50 kg N/ha/yr
are likely to impact habitat extent ITable 10-4; (Latimer and Rego. 2010)1. These values
were based on literature threshold values from Bowen and Valiela (2001). Hauxwell et al.
(2003). Li et al. (2007). and Latimer and Rego (2010). Greaver et al. (2011) identified the
range of 50-100 kg/N/ha/yr total N loading as the empirical critical load for loss of
eelgrass based on (Latimer and Rego. 2010). These threshold levels for seagrasses were
developed in shallow, poorly flushed systems that may not be universally representative
of all estuaries.
10.7. Summary and Causal Determinations
Causal determinations between N loading in coastal areas and biological effects of N
enrichment (Appendix 10.7.1) and nutrient-enhanced coastal acidification
(Appendix 10.7.2) are presented in this summary along with supporting evidence for
these relationships.
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10.7.1.
Nitrogen Enrichment
In the 2008 ISA, the body of evidence was sufficient to infer a causal relationship
between N deposition and the alteration of species richness, species composition, and
biodiversity in estuarine systems (U.S. EPA. 2008a'). The strongest evidence for a causal
relationship was from changes in biological indicators of nutrient enrichment
(chlorophyll a, macroalgal abundance, HABs, DO, SAV) and N was recognized as the
major cause of harm to the majority of estuaries in the U.S. (Brickcr et al.. 2008; NRC.
2000). Phytoplankton are the base of the coastal food web, and increases in primary
producer biomass and altered community composition associated with increased N can
lead to a cascade of direct and indirect effects at higher trophic levels. For this ISA, new
information is consistent with the 2008 ISA, and the causal determination has been
updated to reflective more specific categories of effects. The body of evidence is
sufficient to infer a causal relationship between N deposition and changes in biota,
including total primary production, altered growth, total algal community biomass,
species richness, community composition, and biodiversity due to N enrichment in
estuarine environments.
Since the 2008 ISA, additional evidence has shown that reduced forms of atmospheric N
play an increasingly important role in estuarine and coastal eutrophication and HAB
dynamics. In many parts of the U.S., especially the Southeast, Midwest and Mid-Atlantic
deposition of reduced N has increased relative to oxidized N in last few decades. The
form of N delivered to some coastal areas of the U.S. is shifting from primarily NOs to
an increase in reduced forms of N. The increase in highly bioreactive reduced N from
deposition and other sources is often a preferred form of N for specific phytoplankton
functional groups (e.g., cyanobacteria, dinoflagellates) including harmful species.
(Glibert et al.. 2016; Paerl et al.. 2000). Atmospheric inputs to estuaries are
heterogeneous across the U.S., ranging from <10 to approximately 70% of the N inputs
(Table 7-9). so deposition may play a significant role in altering nutrient dynamics in
some coastal systems (i.e., Mid-Atlantic, Southeast) that support an increasing human
population and spatial expansion of industrial scale animal operations (Li et al.. 2016d).
Chlorophyll a is a broadly applied indicator of phytoplankton biomass and used as a
proxy for assessing effects of estuarine nutrient enrichment. It can signal an early stage of
water quality degradation related to nutrient loading and is incorporated into water
quality monitoring programs and national-scale assessments, including U.S. EPA's
NCCA and the NEEA. In general, 0-5 |ig/L chlorophyll concentration is considered to be
good, chlorophyll concentrations between 5-20 |ig/L are classified as fair, and >20 |ig/L
indicates an estuary in poor condition. Phytoplankton sampling and sediment core
analysis have shown changes in phytoplankton community structure in estuaries with
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elevated N inputs. These shifts at the base of the food web to species that are not as
readily grazed (e.g., cyanobacteria, dinoflagellates) have a cascade of effects that include
poor trophic transfer and an increase in unconsumed algal biomass, which could
stimulate decomposition, O2 consumption, and the potential for hypoxia (Paerl ct al..
2003). Although availability of N is an important factor for productivity in estuaries,
many newer studies have emphasized the role of physical and hydrologic factors in
increase of chlorophyll a and other indicators of N enrichment (Hart et al.. 2015; Turner
et al.. 2015; Wilkerson et al.. 2015; Glibert et al.. 2014; Kennison and Fong. 2014;
Rothenberger et al.. 2014; Paerl et al.. 2010; Yang et al.. 2008).
Incidence of HAB outbreaks continues to increase in both freshwater and coastal areas, a
problem that has been recognized for several decades (U.S. EPA. 2008a; Bricker et al..
2007; Paerl et al.. 2002; Paerl and Whitall. 1999; Paerl. 1997). Of the 81 estuary systems
for which data were available in the NEEA reported in the 2008 ISA, 26 exhibited a
moderate or high symptom expression for nuisance or toxic algae (Bricker et al.. 2007).
HAB bloom conditions and effects of HAB toxins on coastal biota have been further
characterized since the 2008 ISA (Wood et al.. 2014; Miller et al.. 2010). The form of N
affects phytoplankton growth and toxin production of some HAB species. CyanoHAB
presence has been documented in estuaries along the U.S. Mid-Atlantic and Southeast
coasts as well as Gulf coast estuaries (Preece et al.. 2017).
In addition to phytoplankton, seaweed growth is also stimulated by increased N inputs.
Macroalgal blooms can smother benthic organisms and corals and contribute to the loss
of important SAV by blocking the penetration of sunlight into the water column. Studies
published since the 2008 ISA provide further evidence that macroalgae respond to the
form of N, with some species showing greater assimilation and growth rates with NH44"
than with NO3 (Wang et al.. 2014a; Ale etal.. 2011).
Since the 2008 ISA, seagrass coverage is improving or stable in some estuaries like
Tampa Bay and Chesapeake Bay, while many areas continue to see declines in seagrass
extent. Loss of SAV habitat can lead to a cascade of ecological effects because many
organisms are dependent upon seagrasses for cover, breeding, and as nursery grounds.
SAV is often at a competitive disadvantage under N enriched conditions due to the fast
growth of opportunistic macroalgae that preferentially take up NH4+ and can block light
from seagrass beds. Increased epiphyte loads on the surface of macrophytes from nutrient
enrichment may reduce growth of SAV. Since the 2008 ISA, additional studies are
available on the relationship between N loading and SAV abundance (Bovnton et al..
2014; Orth et al.. 2010; Ruhl and Rvbicki. 2010). and several studies have identified
specific regional thresholds lYBenson et al.. 2013; Latimer and Rego. 2010); Table 10-41.
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Greaver et al. (2011) identified the range of 50-100 kg/N/ha/yr total N loading as the
empirical critical load for loss of eelgrass based on (Latimer and Rego. 2010).
Increased algal biomass associated with nutrient over-enrichment leads to increased
decomposition of organic matter, which decreases DO. Oxygen depletion largely occurs
only in bottom waters under stratified conditions, not throughout the entire water column.
A variety of ecological impacts are associated with low DO (Figure 10-4). and this
indicator has been used in national assessments of coastal condition including U.S. EPA's
NCCA and the NEEA. DO concentrations of 0 mg/L are anoxic, 0-2 mg/L are indicative
of hypoxic conditions, and 2-5 mg/L are biologically stressful conditions (Devlin et al..
2011; Bricker et al.. 2007). For example, many fishes are absent at DO levels below
2 mg/L, and mortality of tolerant organisms can occur at approximately 0.5 mg/L.
Hypoxia has recently been shown to affect reproductive parameters in fish which may
lead to effects at the population level. In laboratory conditions, increased turbidity
associated with eutrophic conditions has been shown to alter fish reproductive behaviors
(Candolin. 2009). Macroinvertebrate community structure is also affected by the duration
and severity of hypoxia. A shift to benthic organisms with shorter life spans and smaller
body size has been observed in coastal areas that experience severe seasonal hypoxia
(Diaz and Rosenberg. 2008). Reduced species density and diversity in the northern Gulf
of Mexico are linked to the duration of hypoxic events (Baustian and Rabalais. 2009).
Post-2007 literature includes additional information on the extent and severity of
eutrophication and hypoxia in sensitive regions. Diaz et al. (2013) assessed global
patterns in hypoxia from the 1960s to 2011. Areas of eutrophication-related hypoxia are
found on the U.S. East and West Coasts and the Gulf of Mexico (Figure 10-5). Overall,
the extent of hypoxia is increasing globally with some areas showing signs of
improvement. In the 2008 ISA, the largest zone of hypoxic coastal water in the U.S. was
the northern Gulf of Mexico on the Louisiana-Texas continental shelf (U.S. EPA. 2008a).
This area, which forms annually between May and September, continues to be the largest
in the U.S. and the second largest in the world, averaging about 16,500 km2 (10,250 mi2)
in size (Dale et al.. 2010; Jewett et al.. 2010). In the summer of 2017, the hypoxic zone in
the Gulf was the largest ever measured at 14,123 km2 (8,776 mi2) (U.S. EPA. 2017e).
Atmospheric deposition from watersheds in the Mississippi/Atchafalaya River basins
contributes approximately 16 to 26% of the total N load in the Gulf of Mexico
(Robertson and Saad. 2013; Alexander et al.. 2008). Long Island Sound also experiences
periods of anoxia in some years. In other U.S. coastal systems, the incidence of hypoxia
is increasing; however, the severity of DO impacts are relatively limited temporally and
spatially (Jewett et al.. 2010; U.S. EPA. 2008a; Bricker et al.. 2007). In the Pacific
Northwest, coastal upwelling can be a large source of nutrient loads, and advection of
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upwelled water can introduce into estuaries hypoxic water that is not related to
anthropogenic sources.
The indicators described above (chlorophyll a, HABs, macroalgal abundance, DO, SAV,
and benthic diversity) have been incorporated into indices that describe eutrophic
conditions in coastal areas. In the 2008 ISA, the ASSETS ECI developed for the NEEA
was used as an assessment framework for eutrophic conditions in coastal U.S. estuaries
(Bricker et al.. 2007). The NCCA incorporates indicators that include chlorophyll a and
DO to assess U.S. waters (U.S. EPA. 2016g. 2012b). Additional indices of overall
condition of estuarine functioning that incorporate biological indicators of eutrophication
have since been developed both in the U.S. and other countries (Boria et al.. 2012; Devlin
et al.. 2011; Boria et al.. 2008). Comparisons of these frameworks have led to the
identification of more robust and representative methods to measure estuarine response,
including incorporation of annual data, frequency of occurrence, spatial coverage,
secondary biological indicators, and a multicategory rating scale (Devlin etal.. 2011).
In addition to having effects on biota in estuarine environments, N enrichment is one of
several factors linked to increased disease susceptibility, bleaching, and reduced
calcification rate in corals (Appendix 10.4.2). Near-coastal coral reefs in the U.S. occur
off southern Florida, Texas, Hawaii and U.S. territories in the Caribbean and Pacific. The
threatened status of staghorn coral (Acropora cervicornis) and elkhorn coral (Acropora
palmata) under the U.S. Endangered Species Act has been linked to indirect N pollution
effects, specifically low DO and algal blooms that alter habitat, and to other non-nutrient
stressors (Hernandez et al.. 2016).
10.7.2. Nutrient-Enhanced Coastal Acidification
Since the 2008 ISA, N enrichment has been recognized as a possible contributing factor
to increasing acidification of marine environments. Dissolution of atmospheric
anthropogenic CO2 into the ocean has led to long-term decreases in pH (Wallace et al..
2014; Orr et al.. 2005). With increasing N inputs to coastal waters, CO2 in the water
column is produced from degradation of excess organic matter from changing land use,
as well as respiration of living algae and seagrasses, which in turn can make the water
more acidic (Wilson et al.. 2016; Howarth et al.. 2014; Wallace et al.. 2014; Wetz and
Yoskowitz. 2013; Sunda and Cai. 2012; Cai et al.. 2011c; Howarth et al.. 2011).
Nutrient-enhanced coastal acidification has been documented in systems with strong
thermal stratification with spatial or temporal decoupling of production and respiration
processes. Models show that while the impact of each acidification pathway (N
enrichment and atmospheric CO2 dissolution) may be moderate, the combined effect of
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the two may be much larger than would be expected by just adding the effects of each
pathway together (Sunda and Cai. 2012; Cai etal.. 2011c). The body of evidence is
suggestive of, but not sufficient to infer a causal relationship between N deposition
and changes in biota, including altered physiology, species richness, community
composition, and biodiversity due to nutrient enhanced coastal acidification.
Ocean acidification, which can be exacerbated by elevated N input, is projected to alter
marine habitat, have a wide range of effects at the population and community level and
affect food web processes (Mostofa et al.. 2016; Andersson et al.. 2015; Gavlord et al..
2015; Sunda and Cai. 2012; Donev et al.. 2009). Organisms that produce calcium
carbonate shells, such as calcareous plankton, oysters, clams, sea urchins, and coral are
impacted by increasing acidification of waters (Barton et al.. 2015; Kroeker et al.. 2013;
Barton et al.. 2012; Hettinger et al.. 2012). Decreased saturation rates of aragonite and
calcite, two minerals needed in shell formation, are observed in acidic conditions. These
physiological responses to acidifying conditions in coastal waters may lead to effects at
the population level. Changes in the carbonate system, including decreased pH, have
been shown to elicit biological responses in commercially important species from the
New England coast and there are documented declines in oyster production on the U.S.
West Coast linked to ocean acidification. Research on costressors associated with
conditions of coastal acidification and eutrophication suggest that interactions between
elevated CO2, decreasing pH, and nutrient inputs are complex.
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APPENDIX 11. NITROGEN ENRICHMENT EFFECTS
IN WETLANDS
This appendix describes the effects of N deposition on wetland ecosystems. The
introduction contains an overview of the relationship between N deposition and wetland
endpoints, as well as the wetland classification system used in this appendix
(Appendix 11.1). Regional sensitivity in wetlands is related to position within the
watershed (Appendix 11.2). N deposition causes changes to biogeochemical pools and
processes in wetlands, specifically to N (Appendix 11.3.1) and C cycling
(Appendix 11.3.2). Nitrogen eutrophication affects wetland primary producers via
alteration of aboveground plant biomass (Appendix 11.4). plant stoichiometry and
physiology (Appendix 11.5). plant architecture (Appendix 11.6). and plant demography
(Appendix 11.7). Nitrogen eutrophication also alters wetland biodiversity via changes to
plant community composition (Appendix 11.8.1) and to consumer communities
(Appendix 11.8.2. There are a number of critical loads published for wetland ecosystems
(Appendix 11.9).
11.1. Introduction
The 1993 Oxides of Nitrogen Air Quality Criteria document (hereafter referred to as the
1993 AQCD) and the 2008 ISA for Oxides of Nitrogen and Sulfur—Ecological Criteria
(hereafter referred to as the 2008 ISA) evaluated the effects of nitrogen deposition on
wetland ecosystems. The 1993 AQCD found that the three main ecological effects of N
deposition on wetland ecosystems were reduced biodiversity, modified microbial
processes, and increased primary production. The 2008 ISA supported and extended the
conclusions in the 1993 AQCD, especially with regard to the effects of N deposition on
species diversity, and also found evidence for alterations of ecosystem nitrogen and
carbon cycling. The 2008 ISA found that the evidence was sufficient to infer a causal
relationship between N deposition to wetland ecosystems and the alteration of
biogeochemical cycling, as well as the alteration of species richness, species composition,
and biodiversity. New evidence published between 2008-2015 from observational
studies, experimental N additions in the field and in mesocosms, and reanalysis of large
data sets supports and extends the conclusions of the 2008 ISA. The body of evidence is
sufficient to infer a causal relationship between N deposition and the alteration of
biogeochemical cycling in wetlands. In addition to new information on species
biodiversity, there is recently published evidence of N deposition effects on endpoints not
covered in the 2008 ISA, including alterations to plant physiology and plant architecture.
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The body of evidence is sufficient to infer a causal relationship between N deposition
and the alteration of growth and productivity, species physiology, species richness,
community composition, and biodiversity in wetlands.
Wetland vegetative communities are adapted to high levels of natural organic acidity, so
S and N deposition are unlikely to cause any acidification-related effects at levels of
acidic deposition commonly found in the U.S. (U.S. EPA. 2008a. in Annex B).
Sensitivity of wetlands to N deposition as a nutrient are well documented in the 2008 ISA
(Appendix 11.2). Hydrologic pathways are often the same pathways of N input;
therefore, they are useful for discussing the N sources and sensitivity to atmospheric N
deposition. In general, wetland types (Table 11-1) that receive most of their total N input
from the atmosphere are most sensitive to atmospheric deposition. Nearly all new N
comes from atmospheric deposition in ombrotrophic bogs because these wetlands only
receive water inputs via precipitation. They develop where precipitation exceeds
evapotranspiration and where there is some impediment to drainage of the surplus water
(Mitsch and Gosselink. 1986). Freshwater fens, marshes, and swamps are characterized
by ground and surface water inputs that are often on the same order of magnitude as
precipitation (Koerselman. 1989). Lastly, estuarine and coastal wetlands receive water
from precipitation, ground/surface water, and marine/estuarine sources. Therefore, bogs
(and presumably vernal pools) are among the most vulnerable wetlands to the
nutrient-enrichment effects of N deposition (Krupa. 2003).
Table 11-1 Wetland classification used in the Integrated Science Assessment.
Wetland Term Definition (adapted from, Wakelev. 2002: Mitsch and Gosselink. 2000: Cowardin et al„ 1979)
Soil-based classification
Peat	Substrate consisting of partially decomposed plant litter. In bryophyte-dominated wetlands (bogs
and fens), this layer is composed of dormant or dead moss. In marshes and swamps, peat consists
of dead litter, including plant leaves, stems, and roots, as well as undifferentiated soil organic
matter and sediments.
Peatland
Any wetland that accumulates or has historically accumulated organic C stores in the substrate.

Net primary productivity and accumulation of fixed C exceed decomposition rates. In Europe, the

synonymous term "mire" is used.
Hydrology-based classification
Permanent
Wetland where a zone of the soil profile or peat profile remains saturated with water. The deeper
wetland
soil profile is typically saturated, anoxic, and often has a high percentage of organic matter.
Intermittent
Wetland where soils are saturated with water for short periods of time, annually or every few years;
wetland
soils tend to have higher mineral contents than do permanent wetland soils.
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Table 11-1 (Continued): Wetland classification used in the Integrated Science
Assessment.
Wetland Term Definition (adapted from, Wakelev. 2002: Mitsch and Gosselink. 2000: Cowardin et al.. 1979)
Intertidal	Wetlands where soils or sediments are inundated with water at high tide, and are exposed to air at
wetland	low tide. This includes salt marshes, mangroves, and unvegetated mudflats.
Soil-, hydrology-, and vegetation-based classification
Bog or peat Defined by vegetation. A wetland with high moss cover and a depth profile with high organic
bog	content (peat), sometimes comprised of compressed senesced mosses. Ericaceous shrubs and
certain graminoids are also particularly adapted for bog conditions.
Ombrotrophic A peatland dominated by acidophilic mosses, particularly Sphagnum mosses. Water is derived
or oligotrophic entirely from precipitation, with high [DOC], low pH, and low nutrient concentrations.
bog
Mesotrophic, A peatland that supports herbaceous and woody plant species with a wetland surface composed of
poor, or	moss species. Water sources include precipitation, surface water, and groundwater. Water pH
intermediate	(acidic to circumneutral) and nutrient concentrations are intermediate between ombrotrophic bog
fen	and minerotrophic fen.
Minerotrophic,	A peatland dominated by herbaceous graminoid species typical of marshes, with a mat composed
rich, eutrophic,	of moss species. Water sources include precipitation, surface water, and groundwater. Water has
or calcareous	higher pH (circumneutral to alkaline), as well as higher concentrations of calcium ions and other
fen	nutrients, than other bog types.
Swamp	Defined by vegetation. A wetland dominated by woody plants, particularly tree and shrub species
with physiological adaptations to occasional or constant soil inundation.
Mangrove An intertidal swamp. A subtropical or tropical swamp dominated by halophytic trees and shrubs.
Tidal inundation by ocean or estuarine waters is frequent. In the U.S., dominant species are trees
in the Avicennia and Rhizophora genera. Mangroves are important nursery habitats for marine
animal species.
Marsh	Defined by vegetation. A wetland dominated by herbaceous plants with physiological adaptations
to occasional or constant soil inundation.
Freshwater Marsh dominated by herbaceous plant community composed of species with physiological
marsh	adaptations to soil inundation. Includes marshes in lakes, rivers, and nonsaline regions of
estuaries. Salt- and sulfide-intolerant marsh plant species are common. Soils may be peat or
mineral sediments.
Pothole	A type of freshwater marsh. A shallow pond with herbaceous and submerged aquatic vegetation,
common in the prairie ecosystems of the Dakotas and central Canadian plains. A pothole is often
not connected by surface water flow to its watershed; water sources include precipitation and
groundwater.
Tidal marsh An estuarine marsh. Used in this document to describe estuarine marshes in which water level
fluctuates in response to ocean tides but marsh water has low or no salinity (i.e., water sources are
freshwater). Depending on physical location in the estuary, occasional saltwater intrusion may alter
biogeochemistry and plant community. A mix of saline-tolerant and saline-intolerant plant species
is common.
Salt or coastal Halophytic herbaceous species are dominant. Tidal inundation by ocean water structures the plant
marsh	community, which typically forms zones of single species monocultures or low-diversity
communities of plants with similar tolerances for salt and water stress. Plant roots trap alluvial
sediments to build and maintain the elevation of the marsh surface, or platform. Salt marshes are
important nursery habitats for marine animal species.
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Table 11-1 (Continued): Wetland classification used in the Integrated Science
Assessment.
Wetland Term Definition (adapted from, Wakelev. 2002: Mitsch and Gosselink. 2000: Cowardin et al.. 1979)
Intermittent Defined by hydrology and vegetation. Wetland where soils are saturated, with soil processes
typical of aquatic sediments, for short periods annually or every few years; inundation occurs with
sufficient frequency to prevent persistence of flooding-intolerant plant species.
Vernal pool An area with mineral soils which is flooded by rainwater or rising water tables and then isolated
from surface water flow as it slowly dries up. These seasonally inundated areas are important
nursery habitat for a number of invertebrate and vertebrate species.
Riparian	An area with mineral soils, and a high water table due to close proximity to a river, stream, or lake,
wetland	Periodic inundation occurs when water levels rise. Vegetation is a mix of terrestrial plants and
plants with physiological adaptations to occasional soil inundation.
C = carbon; DOC = dissolved organic carbon.
11.2. Regional Sensitivity
There is no new information on regional sensitivity of wetlands to N deposition, and
models of sources of N to wetland ecosystems are not yet available. In the 2008 ISA,
wetlands were described in order of sensitivity to N deposition. Bogs and fens are among
the most sensitive wetland ecosystems to N deposition. In the U.S., peat-forming bogs
and fens are most common in glaciated areas, especially in portions of the Northeast and
Upper Midwest (U.S. EPA. 1993). N deposition was projected to drastically change
species composition based on experimental results in European fens (Pauli et al.. 2002;
Aerts and de Caluwe. 1999). N input and output rates of freshwater or riparian marshes
and swamps are intermediate between bogs and coastal marshes. N deposition increased
primary productivity 30% and increased methane emissions in freshwater marshes.
Atmospheric N inputs contribute to eutrophication problems in coastal marshes at many
locations through direct deposition to the marsh, indirect deposition to the watershed, or
direct deposition to estuaries or coastal waters followed by tidal delivery of aqueous N
loads to wetlands (see Table 7-9 for estimates of N deposition to estuaries). However,
marine inputs of N are typically higher than direct atmospheric input to coastal wetlands.
The National Wetlands Condition Assessment (NWCA) was conducted by U.S. EPA in
2011 to characterize the condition of North American wetlands under multiple
anthropogenic stressors and to provide a baseline for future assessments. The final report
did not provide soil N or surface water ammonia, nitrate-nitrite, or total N, although these
data were collected as part of the survey and are available (U.S. EPA. 2016i). Ecological
endpoints that the NWCA identified as correlated with total N loading are the Normative
Plant Stressor Indicator (NPSI; see Appendix 11.8.1.3 for an explanation of the link
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1	between N loading and non-native plants) and microscystin concentrations (see
2	Appendix 9). although the dataset includes a number of other metrics that may respond to
3	N deposition. Nationally, 39% of wetland area was estimated to experience moderate to
4	very high stress based on the NPSI, with stresses particularly high in wetlands of the
5	Interior Plains and West (see Figure 11-1). Microcystis a toxin produced by
6	cyanobacteria in response to available N, was detected in low concentrations in surface
7	water at 12% of national wetland sites, with particularly high detection rates (34%) in the
8	Interior Plains (U.S. EPA. 2016i).
National
Coastal
Plains
Eastern Mtn. &
Upper Midw.
hfc
CJV
&SK
'\r \>

Interior
Plains
West
Nonnative Plants
Percent Area
Nonnative Plants
Area
20 40 60 80 100 0
Percent Area
i i Lnwi I Moderate
High I
20,000,000
Area
Very High
40,000,000
Source: U.S. EPA (2016iV
Figure 11-1 Estimated extent of wetlands stressed by nonnative plants as
determined by Nonnative Plant Stressor Indicator, at a national or
regional level.
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11.2.1. Climate Modification of Ecosystem Response to Nitrogen
Changes in mean annual temperature and frequency and magnitude of precipitation will
affect the responses of all wetlands to N loading; changes in sea level will affect the
response of salt marsh, mangroves, and freshwater tidal wetlands. Temperature effects on
wetlands have been demonstrated in European bog ecosystems, where increased
temperatures increased cover of woody species and decreased Sphagnum moss cover just
as N deposition does (Hcdw all et al.. 2017). In metaregressions of Sphagnum moss
response to N addition and temperature, the two factors synergistically depressed
Sphagnum production, with an 1°C increase in summer temperature having an impact on
Sphagnum equivalent to an additional 40 kg N/ha/yr (Limpens et al.. 2011). Increasing
temperatures may strengthen N effects upon wetland ecosystems.
Hydrologic regimes are important controls on wetland cycling and productivity, so
changes in the magnitude and frequency of precipitation can have strong effects on
ecosystem N retention and C storage. The same metaregression of Sphagnum mosses
referenced above found that increased precipitation also increased Sphagnum sensitivity
to N addition-induced decreases in production (Limpcns et al.. 2011). Experimental
mesocosms modelling changes in precipitation to salt marshes found that precipitation
delivered in infrequent, heavy storm events decreased N retention and plant productivity,
even though storm events delivered a higher N deposition load to the marsh (Hanson et
al.. 2016; Oczkowski et al.. 2016). Shifts in precipitation towards less frequent, more
intense rain events may strengthen N deposition-induced decreases in salt marsh N
retention while weakening N deposition-induced increases in plant productivity.
Sea level rise will affect tidal wetlands through salt water intrusion into tidal freshwater
wetlands (Barcndrcgt and Swarth. 2013) and through increasing inundation of salt
marshes, both of which have the potential to change community composition and
decrease wetland extent. Experimental sea level rise of 10 cm increased N stimulation of
belowground and aboveground plant biomass in Kirkpatrick Marsh (see (Langlcv et al..
2013) in Appendix 11.3.2.1.2 and Appendix 11.4.1). but the response was
species-specific, with only one species responding to the N deposition. Sea level rise
effects on marsh response to N deposition are not well understood and may vary by
species.
11.3. Soil Biogeochemistry
The 2008 ISA described the relationship between N deposition and the alteration of
biogeochemical cycling in wetlands. N deposition alters N cycling by altering microbial
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communities, the rates at which they transform N between pools, ecosystem retention or
release of transformed N, and the emissions of the greenhouse gas nitrous oxide
(Appendix 11.3.1). N deposition alters C cycling by altering soil processes such as
decomposition and respiration (resulting in emissions of the greenhouse gases carbon
dioxide and methane), as well as by altering soil pools such as belowground biomass and
soil organic matter (Appendix 11.3.2). There is new evidence since 2008 of N deposition
effects upon biogeochemical cycling across wetlands, as well as in salt marshes,
mangroves, bogs, and riparian wetlands. The body of evidence is sufficient to infer a
causal relationship between N deposition and the alteration of biogeochemical
cycling in wetlands.
11.3.1. Nitrogen Pools and Processes
Water table depth tends to fluctuate more in wetlands than terrestrial ecosystems. The
variable depth creates dynamic anoxic-oxic boundaries and many varied niches for
different microbial communities. As a result, wetlands are important disproportionately to
their spatial area on the landscape (i.e., hotspots) for the microbial transformation of
nitrogen between oxidized and reduced forms. These nitrogen transformations can
improve water quality of downstream streams, lakes, and estuaries by removing N in the
soil or water of an ecosystem into the atmosphere. However, these transformations can
result in emissions of the potent greenhouse gas nitrous oxide. Alteration to N cycling or
N cycling microbial endpoints indicate changes to the important wetland function of
improving water quality.
N deposition contributes to total N load in wetlands. The chemical indicators that are
typically measured include NOs and NH4+ leaching, DON leaching, N mineralization,
and denitrification rates, including N2O emissions produced by incomplete
denitrification. N dynamics in wetland ecosystems are variable in time and among types
of wetlands and environmental factors, especially water availability (Howarth ct al..
1996a). A wetland can act as a source, sink, or transformer of atmospherically deposited
N (Devito et al.. 1989). and these functions can vary with season and hydrological
conditions. Vegetation type, physiography, local hydrology, and climate all play
significant roles in determining source/sink N dynamics in wetlands (Mitchell et al..
1996; Arhcimcr and Wittgrcn. 1994; Koerselman et al.. 1993; Devito et al.. 1989).
N mineralization (microbial transformation of organic N to inorganic forms of N) has
been shown to increase with N addition, and this can cause an increase in wetland N
export to adjacent surface water (Groffman. 1994). In general, ecosystem leaching losses
of NO3 from wetlands to downstream aquatic systems are small, although recent
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research suggests that N addition to peatlands increases export of N in both inorganic and
organic forms (Edokpa et al.. 2015). Bogs and fens in the Marcell Experimental Forest
that receive 2.8-4.7 kg N/ha/yr as atmospheric deposition export 0.26 (bogs) and 1.34 kg
N/ha/yr (fens), respectively (Hill et al.. 2016). Wetlands tend to have low N export to
surface water because anoxic zones within wetlands are favorable for microbial
denitrification of N from NOs to gaseous N forms. Elevated N inputs to wetlands will
often increase the rate of denitrification (Duffy and Kahara. 2011; Cooper. 1990;
Broderick et al.. 1988; Dierberg and Brezonik. 1983). because N additions to aquatic
environments with high organic C increase denitrifier abundance, activity, and
denitrification rates (Kim et al.. 2015b). This mitigates environmental effects associated
with increased N supply to soils and drainage waters; however, increased denitrification
may also increase the emissions of greenhouse gases (e.g., N2O) to the atmosphere.
Denitrification appears to be negligible in wetland environments that are typically
nutrient (including N) poor, such as some bogs and fens (Morris. 1991). Incomplete
denitrification and emission of N2O tend to be higher in freshwater tidal wetlands than in
saline wetlands like mangroves or salt marshes (Welti et al.. 2017). In salt marshes,
dissimilatory nitrate reduction to ammonium (DNRA) is an important process that cycles
N while retaining N within marsh sediments (see also Appendix 7.2.6.5). N addition
decreases DNRA while increasing nitrification and denitrification rates (Peng et al..
2016). decreasing N retention in salt marshes, and increasing N2O emissions (Chmura et
al.. 2016; Moseman-Valtierra et al.. 2011).
New studies summarized below in salt marsh, mangrove, peat bog, and riparian wetlands
have evaluated the effects of N loading/N addition on endpoints related to N cycling. In
addition, there is a new study on wetland N removal, synthesizing information across
wetland types.
11.3.1.1. Across Wetlands
N removal, defined as the sum of denitrification, plant uptake, and burial in sediments,
was evaluated by analyzing data from 109 wetlands distributed globally, including
agricultural wetlands, freshwater marshes, freshwater swamps, and coastal marshes
(Jordan et al.. 2011). Total N loads to these wetlands ranged from 0.2-90,480 kg N/ha/yr;
however, the sources of reactive N were not identified by the authors. Across this global
data set, wetland N removal as measured in outflow is proportional to wetland N load. N
removal efficiency was 26% higher in nontidal wetlands than in tidal wetlands.
The 2008 ISA found evidence that N deposition alters the emission of nitrous oxide
(N2O) and methane (CH4), which contribute to global warming. A meta-analysis of
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19 N addition observations, ranging from 15.4 to 300 kg N/ha/yr, found that N
enrichment increased wetland N2O emissions by 207% (Liu and Greaver. 2009).
11.3.1.2. Salt Marsh
New studies have evaluated the effects of N loading/N addition on tidal export, N
mineralization, nitrification, and microbial community structure in salt marshes. Tidal
export of N from salt marshes was studied in a long-term (>30 years) fertilization
experiment in Massachusetts (Brin et al.. 2010) in which tidal export of N increased with
increasing N load. Tidal export of ammonia increased linearly with increasing annual N
load, and tidal export of nitrate increased exponentially with increasing annual load (see
Table 11-2 for equations). Likewise, a linear relationship was documented between
increasing N and decreasing potential N mineralization after 7 months of N addition to
three salt marshes in California [see Table 11-2 for equation; (Vivanco et al.. 2015)1. In
this same study, there was no relationship between N addition and net nitrification rates;
however, the nitrification rates differed among the marsh sites.
In Kirkpatrick Marsh, MD, N addition decreased N retention in marshes dominated by
native plants and by the invasive lineage of Phragmites australis. N addition of 250 kg
N/ha/yr decreased ecosystem N retention by 55%, driven largely by a 57% decrease in N
retention belowground, in roots, rhizomes, and soil (Pastore et al.. 2016). N addition also
altered the relative allocation of N to pools in marsh plants. N addition increased the total
pool of N stored in aboveground biomass through stimulatory effects on biomass
production (see Appendix 11.4.1). but decreased the total pool of N stored in
belowground biomass by 6%. N addition decreased the pool of N in belowground
biomass but also shifted the distribution of root N higher in the soil profile than in control
plots (to 15-25 cm depth from 40-50 cm depth), making it more vulnerable to leaching
and to microbial transformation: N addition increased pore water NH4 concentrations
120% and N2O flux out of soils by 220% (Pastore et al.. 2016). The invasive species P.
australis established roots at depths below the native plant community (see
Appendix 11.3.2.1.2). which increased mineralization deep in the peat profile to increase
pore water [NH4+] 9-72% at 40-70 cm (Mozdzer et al.. 2016). Like many terrestrial
invasive species, P. australis may respond to N loading by creating a positive feedback
loop of increasing soil N availability and expanding invasion. Increasing reactive N may
change the diversity and composition of microbial communities responsible for
denitrification and nitrification. Change in the microbial communities could change the
rate of the chemical reactions that microbes perform in the wetland. Increasing N loads
alter the abundance of denitrifying bacteria.
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In the long-term marsh fertilization experiment in a salt marsh in Massachusetts, nirS
DNA sequences were amplified to assess denitrifying bacteria (Bowcn et al.. 2013). The
plots sampled were fertilized with sewage sludge, which contained other nutrients,
metals, and organic material in addition to N, and N addition rates were 221, 655, or
1,966 kg N/ha/yr. With the addition of N, denitrifying bacteria communities became
more dissimilar, as widely distributed bacteria declined, and abundance and richness of
unique denitrifiers increased (Bowcn et al.. 2013). Lage etal. (2010) sampled sediments
associated with Spartina patens in a New England salt marsh. This study used the a mo A
DNA sequence in sediments to sample the ammonia-oxidizing bacteria, which produce
nitrites from ammonia in the first step of nitrification. The majority of species affected by
N addition were within a clade of Nitrosospira-like sequences found in other salt
marshes. N addition decreased evenness of distribution and altered composition,
specifically the relative abundance of taxa within the marine and estuarine
Nitrosospira-like clade, suggesting that there are fine-scale genetic differences to the
bacteria's ability to use nitrogen.
An incubation experiment of soils collected from different zones of salt marsh in
Yancheng Nature Reserve added ammonium or nitrate in equal amounts of N to assess
what effects N chemical species had on microbial cycling of N. N addition increased net
N mineralization in all marsh zones, with oxidized N increasing net N mineralization
16-29%, and reduced N increasing net mineralization 58-69% (Zhang et al.. 2016c). N
forms had similar rates upon net nitrification rates, as NOs addition increased
nitrification rates 34-54%, and NH4 addition increased nitrification rates 65-94%. N
addition also increased the temperature sensitivity of N mineralization (Q10) in low
marshes (Zhang et al.. 2016c). suggesting a synergistic stimulation between increased N
loading and temperature upon N release from salt marshes to the marine environment.
11.3.1.3. Mangrove
In mangrove ecosystems, N addition suppressed N fixation and increased denitrification
(Whigham et al.. 2009). Laboratory incubations showed that the denitrification rate was
15 times higher in soils that had received 100 kg N/ha/yr than in control soils (N load on
N deposition in control soils not reported). However, field measurements of N2O
emissions were 5.6 times higher in fertilized than in control soils, indicating incomplete
denitrification in fertilized soils. Nitrogen fixation rates were suppressed by 88% in
fertilized plots (Whigham et al.. 2009).
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11.3.1.4. Freshwater Tidal Marsh
The Madisonville Nutrient Plots on the Tchefuncte River, LA, have received N addition
of 0, 50, 200, or 1,200 kg N/ha/yr for 11 years. There was no measurable effect of N
addition upon total N, microbial N, soluble N, denitrification rates, or C cycling in the top
10 cm of soil (Stcinmullcr et al.. 2016). although there were effects upon aboveground
productivity, plant stoichiometry, and plant community (see Appendix 11.4.2.
Appendix 11.5.3. and Appendix 11.8.1.2).
11.3.1.5. Riparian or Intermittent Wetlands
New studies on riparian wetlands include N addition relationships to denitrification rates
and bacterial community composition. In a riparian wetland in Durham, NC, wetland
species' potential denitrification activity increased linearly with increasing soil total
inorganic N [see Table 11-2 for equation; (McGill et al.. 2010)1. In freshwater riparian
wetlands along the James River, VA, soils were collected and incubated with added
nitrogen. Nitrogen addition increased the abundance of denitrifying bacteria—quantified
as copies of the nirS sequence—by 541% when labile organic matter was also added
(Morrissev et al.. 2013). When N addition occurred simultaneously with the addition of
recalcitrant organic matter, denitrifying bacterial abundance decreased 96% with
increased N. Community composition of denitrifying bacteria shifted in response to N
addition, as did the composition of bacteria capable of dissimilatory nitrate reduction to
ammonia [quantified as copies of nrfA sequence; (Morrissev et al.. 2013)1.
In riparian forests, increased N loads altered the symbiotic association between A In us
incana ssp. tenuifolia (grey alder) and Frankia (actinorhizal, N-fixing bacteria). N
addition of 100 kg N/ha/yr decreased the density of Frankia-hosting root nodules by 62%
compared with nodules from unamended control plots. Nodule respiration rates were
28% lower under N fertilization, indicating lower rates of Frankia metabolism, and N
fixation in nodules was 31% lower than in controls (Ruess et al.. 2013).
A recent study of an alpine meadow and intermittent wetland on the Tibetan Plateau
found that adding 30 kg N/ha/yr changed the wetland from a net sink to a net source of
N2O emissions to the atmosphere, suggesting N addition stimulated microbial
denitrification (Wang et al.. 2017c).
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11.3.1.6.
Bog and Fen
In peat bog ecosystems, N addition decreases ecosystem retention of N and increases N
exports in surface water from hydrologically connected bogs. In Mer Bleue Bog, Ontario,
N addition decreased the N retention efficiency of the ecosystem, as measured by the
recovery of an 15N tracer. When N was added at the rate of 16 kg N/ha/yr from
2000-2007, a pulse of added 15N was recovered at lower rates from fertilized plots than
from unamended control plots (N deposition was 8 kg N/ha/yr), indicating a 71%
reduction in average retention efficiency across biomass pools (Xing etal.. 2011). A
recent study (samples taken in 2014) in this bog found increases in nitrate and ammonium
concentrations in response to N addition in peat, with N addition increasing |NOs |
71-169% and [NH/] 90-269%, suggesting that the living Sphagnum moss is no longer
retaining N but leaching it into the peat below (Pinsonncault et al.. 2016).
A European study of N pools in bogs found that N deposition changes the retention of N
within vascular plants and peat soil. Under N treatment equivalent to deposition of 2 kg
N/ha/yr, 15N recovery was 66% for mosses, 3.8% for shrubs, and 1.8% for graminoids.
Deposition of 47 kg/ha/yr changed 15N recovery to 12% for shrubs as well as increasing
the total N stored in shrubs by 360% compared to mesocosms receiving 2 kg N/ha/yr. 15N
recovery also declined in peat under N deposition, with 15N recovery of 60% under 2 kg
N and 15N recovery of 36% under 47 kg N (Zaiac and Blodau. 2016). Increases in plant N
uptake along with decreasing efficiency of plant N retention (see Appendix 11.5.5) can
lead to increases in dissolved organic N exports from bogs. A study in a NPK-fertilized
and limed (addition of CaCCh) peatland catchment in the U.K. found high levels of N
export (14 kg N/ha watershed/yr) from the wetland to the Kinder River, with over half of
the exported N in the form of dissolved organic nitrogen (Edokpa et al.. 2015).
The Whim bog experiment in Scotland added N as either ammonium or nitrate to
simulate wet deposition, which with ambient N deposition of 8 kg N/ha/yr resulted in
treatment N loads of 16, 32, and 64 kg N/ha/yr. Increasing NOs loading increased
concentrations of dissolved organic nitrogen (DON) in pore water, while increasing NH4+
loading increased cation leaching from moss into pore water, and increased DIN in pore
water at the highest N addition level (Chiwa et al.. 2016). N addition impaired the ability
of the Sphagnum moss mat to absorb and retain N, beginning at 32 kg N/ha/yr for
oxidized N and 64 kg N/ha/yr for reduced N addition. This study illustrates the different
effects of reduced and oxidized N upon bog ecosystems (see also, Appendix 11.5.5).
However, a recent analysis by (Wieder et al.. 2016) suggested that North American and
European Sphagnum species have different tolerances for and responses to N loading, so
critical loads for North American bogs should be inferred with caution from European
bogs.
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1	Recent studies have documented high rates of nitrogen fixation by microbial diazotrophs
2	living on and in dead Sphagnum cells (Vile 2014 and Larmola 2014 in (van Den Elzen et
3	al.. 2017). Microbial N fixation rates were observed at N deposition of 25 kg N/ha/yr, and
4	added another 6 kg N/ha/yr to moss and peat (van Den Elzen et al.. 2017).
11.3.1.7. Summary Table
Table 11-2 New studies on nitrogen addition effects on nitrogen cycling in
wetlands.
Type of
Additions or Load
Biological and



Ecosystem
(kg N/ha/yr)
Chemical Effects
Study Site
Study Species
Reference
Agricultural
0.02-90,480 kg
Removal: wetland N
Global
Reanalysis of
Jordan et al.
wetlands;
N/ha/yr as N load,
removal (Nremoval) is

data from
(2011)
intertidal
mean load for each
proportional to N load

109 global

marshes;
wetland class:
(Nload).

wetlands

freshwater
agricultural: 426,
log(Nremovai) — 0.943 x log



bogs and
intertidal: 211,
(Nioad) - 0.033. N



marshes;
freshwater bogs
removal efficiency is 26%



freshwater
and marshes: 890,
higher in nontidal than



swamps;
freshwater
tidal wetlands.



other
swamps: 69, other:




wetlands
280 kg N/ha/yr
Deposition = not
reported




Wetlands
N addition
N addition increases N2O
Global
Meta-analysis of
Liu and Greaver
(natural and
experiments, N
emissions 207% across

data collected
(2009)
agricultural)
addition of 15.4 to
300 kg N/ha/yr
Deposition = not
wetlands (n = 19).

from North
America, South
America,


reported


Europe, and




Asia

Salt marsh
Addition = 180 kg
Tidal export of N
Great
Spartina
Brin et al. (2010)

N/ha/yr (as
increases with increasing
Sippewissett
alterni flora,


Milorganite, NPK
N addition (Nadd as kg
Marsh, MA
Spartina patens,


10-6-4), 520 kg
N/ha/yr).

and Distichlis


N/ha/yr (as urea or
For NH4+ export

spicata


as Milorganite),
(NHxexport, in kg




1,560 kg N/ha/yr
N/season): NHxexport




(as Milorganite)
= 0.00083 x Nadd +0.432.
For NO3" export
(NOxexport, in kg
N/season): NOxexport =
0 122 x 6-1"10018Nadd



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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Salt marsh
Addition = 0,100,
200, 400, 800,
1,600,
3,200 kg N/ha/yr as
urea
N deposition =
3-5 kg N/ha/yr as
reported in
Tonnesen et al.
(2007)
Mineralization: potential
net N mineralization
(Nmin) decreases in a
linear response to added
N (Nadd, in g N/m2/yr):
Nmin — —0.0015 x Nadd
0.022.
Nitrification: no significant
relationship.
Morro Bay
National
Estuary,
Carpinteria Salt
Marsh
Reserve,
Tijuana River
Reserve
Estuary, CA
Salicornia
depressa
(Salicornia
virginica) stands
Vivanco et al.
(2015)
Salt marsh
Addition: 50 mg
N/kg soil, as
(NH4)2S04 or
NaNOs
Ambient
deposition = not
specified
N addition increased net
N mineralization in all
three marsh zones. NO3"
addition increased net N
mineralization 16-29%,
and NH4+ addition
increased net N
mineralization by
58-69%. The most
predictive soil parameter
for net N mineralization
was labile C:labile N
(r = -0.85)
N addition increased net
nitrification in all three
marsh zones. NO3"
addition increased net
nitrification 34-54%, and
NH4+addition increased
net nitrification by
65-94%. The most
predictive soil parameter
for net nitrification was
labile C:labile N
(r = -0.81).
N addition increased Q10
values (stimulatory effect
of temperature upon net
N mineralization) in P.
australis and S.
alterniflora marsh zones.
Aerobic soil
incubations
constructed
from core
samples from
Yancheng
National Nature
Reserve, China
Monospecific
stands across
marsh elevation
gradient:
Suaeda salsa
(native) in high
marsh,
Phragmites
australis
(native) in low
marsh, Spartina
alterniflora
(invasive, native
to North
America) in
previously
unvegetated
mud flats
Zhang et al.
(2016c)
June 2018
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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Biological and
Chemical Effects
Study Site Study Species
Reference
Total marsh retention
(AG biomass, BG
biomass, soil) of 15N label
was 55% lower in N
treatment, because
belowground (BG
biomass + soil) N
retention was 57% lower
in N treatment. Across
elevated CO2 treatments,
belowground N retention
was 51% lower in N
treatment.
N addition increased N
mass in AG biomass for
C4 grasses. N addition
increased N mass in total
plant biomass for the first
6 yr, but not for the last
2 yr of measurement.
N addition decreased N
mass of fine roots. N
addition decreased total
N BG by 6% by shifting N
mass across the soil
profile: N addition
increased BG N mass at
depths of 15-25 cm, but
decreased BG N mass at
depths of 40-50 cm.
N addition increased N2O
flux from plots by 220%
and pore water NH4+
concentrations increased
120%.
Kirkpatrick
Marsh, Rhode
River,
Edgewater, MD
C3
Schoenoplectus
americanus; C4
Spartina patens
and Distichlis
spicata
Pastore et al.
(2016)
N addition increased
pore water
concentrations of NH4+
by 9 and 72% at depths
of 40 and 80 cm,
respectively.
P. australis increased
SOM decomposition in
recalcitrant peat 1.8-1.9
times over rates in
unvegetated peat.
Mesocosms of
invasive (from
seeds) or
native (from
rhzome
fragments)
plants in
Kirkpatrick
Marsh, Rhode
River,
Edgewater, MD
Invasive
Phragmites
australis; native
community of
Spartina patens,
Schoenoplectus
americanus
Mozdzer et al.
(2016)
Type of Additions or Load
Ecosystem (kg N/ha/yr)
Intertidal Addition = 250 kg
marsh	N/ha/yr (as NH4CI)
since 2006
Ambient
deposition = not
reported
Average N loads to
Chesapeake Bay
are 140 kg N/ha/yr
15N tracer applied
2006, measure AG
2005-2013,
measured BG 2014
Intertidal Addition = 250 kg
marsh	N/ha/yr (as NH4CI)
Ambient
deposition = not
reported
Salt marsh Addition = 1,630 kg
N/ha/yr as NH4NO3
Deposition = not
reported
Bacterial community: N Scarborough
addition decreases	Marsh, ME
evenness of
p-proteobacteria Laqe et al. (2010)
containing
amoA gene in
sediments
associated with
Spartina patens
ammonia-oxidizing
bacterial community and
changes community
composition (p = 0.017).
June 2018
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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Salt marsh
Addition = sewage
sludge. Low
fert = 221 kg
N/ha/yr, high
fert = 655 kg
N/ha/yr, very high
fert = 1,966 kg
N/ha/yr
Deposition = not
reported
Bacterial community:
ubiquitous nirS
sequences declined with
increasing N addition
(higher abundance and
richness of unique
denitrifying species in
higher N treatments).
Great
Sippewissett
Marsh,
Falmouth, MA
Denitrifying
bacteria in
sediment
(community)
Bowen et al.
(2013)
Mangrove
Addition = 100 kg
N/ha/yr
Deposition = not
reported
Denitrification:
denitrification rate (lab
incubation) was 15x
faster in N addition
sediments.
Indian River
Lagoon, FL
(Impoundment
SLC-24)
Avicennia
germinans and
associated
sediments
Whiaham et al.
(2009)
Tidal marsh
Addition = 0, 50,
200, and 1,200 kg
N/ha/yr (as
slow-release
methylene urea,
which releases
NH4+), in
combination with 0
or 131 kg P/ha/yr,
to mimic
Mississippi River
diversion N and P
loading rates
Ambient
deposition = not
specified
11 yr of N addition had
no significant effect on
total P, total N, microbial
N, potentially
mineralizable N, or
potential denitrification
rates in the top 10 cm of
marsh soil.
Madisonville
nutrient plots
(microtidal
pulses of
10 cm),
Tchefuncte
River, Lake
Pontchartrain
Estuary, LA
Sagittaria
I and folia,
Polygonum
punctatum,
Eleocharis fallas
Steinmuller et al.
(2016)
Riparian
wetland
Addition = n/a
Deposition = not
reported
Denitrification: potential Durham, NC
1, 4, or 8 f
denitrification activity in
species from
spring increases with
Carex crinita,
total inorganic N in soil
Carex lurida,
(y = 6.96x+ 20.52).
Scirpus

cyperinus,

Juncus effusus,

Panicum

virgatum,

Chasmanthium

latifolium,

Eupatorium

fistulosum,

Veronia

noveboracensis,

Asclepias

incarnata, and

Lobelia

cardinalis.
McGill et al. (2010)
June 2018
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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Type of
Additions or Load
Biological and



Ecosystem
(kg N/ha/yr)
Chemical Effects
Study Site
Study Species
Reference
Riparian
Addition = 100 kg
Frankia nodule densities
Bonanza
Alnus incana
Ruess et al.
floodplain
N/ha/yr
decreased 62%. Nodule
Forest LTER,
ssp. tenuifolia
(2013)
successional
Deposition = not
N fixation declined 31%
AK
and associated

forest
reported
and nodule respiration

Frankia strains


declined 28%.



Riparian
Addition = soil
Nitrate addition increased
James River,
Soil microbial
Morrissev et al.
wetland
incubations with
DNF (copies nirS)
Charles City
communities
(2013)

0.5, 2, 4 mg N/g
abundance 541 % when
County, VA
involved in


wet sediment as
labile OM was present

denitrification


KNOs
and decreased DNF

(DNF, gene


Deposition = not
abundance 96% when

nirS) or


reported
recalcitrant OM was

dissimilatory


present. Community
composition of DNF and
DNRA shifted in
response to nitrate
addition.

nitrate reduction
to NH4+ (DNRA,
gene nrfA)

Seasonal
Addition: 30 kg
N addition significantly
Mesocosms at
Carex
Wana et al.
wetland,
N/ha/yr
increased N2O
Luanhaizi
pamirensis,
(2017c)
alpine
Ambient
emissions, changing
wetlands,
Carex

meadow
deposition = 8.7 to
plots from net sinks to
Tibetan
atrofusca,


13.8 kg N/ha/yr
net sources of N2O to the
Plateau
Hippuris


atmosphere. N addition
reduced the global
warming potential (sum
of CO2, CH4, and N2O
fluxes) of the plots from a
net source to a net sink.

vulgaris,
Triglochin
palustris, and
Heleocharis
spp.

Ombrotrophic
Addition = 16 kg
Retention efficiency: N
Mer Bleue Bog,
Bog plant
Xina et al. (2011)
peat bog
N/ha/yr as NH4NO3
addition decreased the
Ontario,
community:


N deposition = 8 kg
retention efficiency of
Canada
dwarf shrub


N/ha/yr as
ecosystem N pools (15N
(measured
species and


quantified by
tracer) by 71%.
2007)
mosses:


Turunen et al.


Sphagnum


(2004)


magellanicum,
Sphagnum
capillifolium,
and Polytrichum
strictum

June 2018
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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Ombrotrophic
peatland
Addition = 16 kg
N/ha/yr (5N, or 5x
background
deposition), 32 kg
N/ha/yr (1 ON), or
64 kg N/ha/yr
(20N), all as
NH4NO3
Also, 64 kg N/ha/yr
as NH4+ only
(20N-NH4), and
64 kg N/ha/yr as
NO3" only
(2ON-NO3)
Ambient
deposition = 3 to
5 kg N/ha/yr, -50%
as NO3"
Extractable N in peat
increases with N
addition: [NO3"]
increases 71% (5N),
125% (10N), and 169%
(20N); and [NH4+]
increases 90% (5N),
197% (10N), and 268%
(20N).
Mer Bleue bog,
Ottawa,
Canada
Nonliving peat
in top 10 cm of
bog
Pinsonneault et al.
(2016)
Ombrotrophic
peatland
Ambient
deposition = 2 kg
N/ha/yr at DS;
12 kg N/ha/yr at
WM; 47 kg N/ha/yr
at FS
Mesocosms were
established with
NH4NO3 solutions
that mimicked
relative N
deposition levels
15N tracer was
added in 48
applications over
6 mo at a rate of
23 kg N/ha/yr
At DS, 15N recovery was
60% in peat (depth
5-40 cm), 66% in living
Sphagnum, 3.8% in
shrubs, 1.8% in
graminoids, and 0.01%
for DIN (NO3" and NH4+).
At 47 kg N/ha/yr, 15N
recovery was 36% in
peat, 12% in shrubs, and
0.4% for DIN, but
otherwise similar to
recovery for pools at DS.
Data from LV and CF
mesocosms are not
considered because
mesocosm N addition
levels (CF: 300%
increase over DS N
solution, LV: 700%
increase over DS) did not
reflect ambient N
deposition differences
(CF: 750% increase over
DS deposition, LV: 300%
increase over DS).
Mesocosms
constructed
using peat bog
cores from
sites in
northern and
western
Europe—
Degero
Stormyr,
Sweden (DS);
Fenn's,
Whixall, and
Bettisfield
Mosses NNR,
UK (WM);
Frolichshaier
Sattelmoor,
Germany(FS)
(Peat cores
also collected
at Little
Vildmose,
Denmark [LV];
and Cors
Fochno, Wales,
UK [CF])
Sphagnum
capillifolium, S.
fallax, S.
magellanicum,
S. papillosum,
S. pulchrum, S.
rubellum,
Andromeda
polifolia,
Calluna
vulgaris, Erica
tetralix, Rubus
chamaemorus,
Vaccinium
oxycoccos,
Eriophorum
vaginatum,
Eriophorum
angustifolium
Zaiac and Blodau
(2016)
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Table 11-2 (Continued): New studies on nitrogen addition effects on nitrogen
cycling in wetlands.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Peatland
catchment
9.6% of the
catchment area
fertilized with NPK
at 19.5 kg N/ha in
summer
35% of the
catchment area is
limed, 1,000 kg
CaC03/ha in
summer
28 kg N/ha/yr
(Helliwell et al..
2007)
Positive correlation
(r= 0.44) between
annual fluxes of DIN and
DON in stormflow.
Negative correlations
between fall DON and
DIN during baseflow
(r= -0.73) and stormflow
(r= - 0.92).
Annual total dissolved
nitrogen flux in river was
14 kg N/ha/yr, of which
54% is DON.
Kinder River
catchment,
south
Pennines, U.K.
Water samples
from the Kinder
River outlet
Edokpa et al.
(2015)
Ombrotrophic Addition = 8, 24,
peatland and 56 kg N/ha/yr
as either NH4+ or
NO3" since 2002
Ambient deposition
total N = 8 kg
N/ha/yr, 3 kg
N/ha/yr as wet
NOx, 3 kg N/ha/yr
as wet NHx, and
2 kg N/ha/yr as dry
NHx
In the oxidized N addition Whim bog,
plots, increasing uptake Edinburgh,
of N by Sphagnum	Scotland
increased pore water
DON and pore water
anion deficit.
In the reduced N plots,
increasing uptake of N by
Sphagnum increased
pore water cation
(K+ + Mg2+ + Ca2+ + H+).
Pore water DIN
increased only in the
highest addition of NhV
(56 kg N/ha/yr).
Sphagnum
moss
(Sphagnum
capillifolium, a
hummock-
forming
species)
Heathland
community:
Calluna
vulgaris,
Eriophorum
vaginatum,
Hypnum
jutlandicum,
Pleurozium
schreberi, and
Cladonia
portentosa
Chiwa et al. (2016)
DNF = denitrification; DNRA = dissimilatory nitrate reduction to ammonium; fert = fertilizer; ha = hectare; kg = kilogram;
KNO3 = potassium nitrate; LTER = Long-Term Ecological Research; N = nitrogen; NH4+ = ammonium; NH4NO3 = ammonium nitrate;
NPK = nitrogen, phosphorus, potassium; S = sulfur; yr = year.
11.3.2. Soil Carbon Cycling
1	Wetlands can be hotspots for emissions of carbon dioxide and methane. High water levels
2	in wetland soils prevent the rapid decomposition of dead roots and fallen stems by
3	aerobic bacteria or fungi. Over long time periods (decades, centuries, or millennia), these
4	conditions result in large stores of carbon belowground in wetland systems. Methanogens
5	are able to decompose some of these anaerobic, saturated carbon stores to produce
6	methane, and in severe droughts, a larger microbial community can produce large pulses
7	of carbon dioxide by decomposing freshly aerated older organic carbon. Endpoints of
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methane and carbon dioxide emissions have important implications for the wetland
function of long-term carbon storage and regulation of atmospheric composition.
A meta-analysis that included wetlands with other nonforest ecosystems indicated no
effect of N deposition on overall net ecosystem exchange of carbon (see
Appendix 11.3.2.2.1). In other words, any gain in carbon capture by photosynthesis was
offset by ecosystem respiration and C leaching. There were not enough studies to
evaluate wetlands as a separate category.
11.3.2.1. Belowground Decomposition, Respiration, and Biomass
Respiration includes the release of carbon dioxide that can be measured in the air. Soil
respiration may be the result of root or microbial activity. The process of decomposition
often releases carbon dioxide from organic matter. Decomposers feed on dead organic
matter, breaking it down into smaller compounds such as carbon dioxide or methane,
water, and nutrients. Nitrogen may change the rate of respiration and decomposition.
Low decomposition in wetlands results in the buildup of dead plant material in the soil. In
relatively closed systems, such as rain-fed bogs, a nutrient-poor, high-organic-acid
ecosystem is produced. In open systems where flowing surface water supplies additional
nutrients and sediments, the growth of new root systems on top of undecomposed roots
can trap sediment and increase the height of the soil surface. This accretion of new
wetland soil depends on both plant growth and maintenance of old carbon stored in soil
organic matter. Wetland existence depends on the sum of these processes: if the sum of
new plant growth does not exceed the decomposition of the soil organic matter, the
wetland may subside until water levels rise above the soil surface. In this case, the
drowned wetland area converts to an aquatic system. Wetland accretion is particularly
important in estuarine and marine systems, where wave action and tides regularly wash
sediment and dead plant material out of the wetland into aquatic systems, where the
material nourishes aquatic food webs. Belowground endpoints, such as bulk density, soil
organic matter, and root and rhizome production, are all indicators of wetland health and
continued existence (Table 11-3).
11.3.2.1.1. Across Wetlands
A new study of 90 wetlands around the Gulf Coast from Florida to Texas found that
higher N in the soil correlated with lower soil bulk density | see Table l l -3 for equation;
(Ncstlcrodc et al.. 2014)1. Wetlands included both tidal and nontidal marshes and
swamps. Reactive N or atmospheric N deposition loads were not quantified. This result
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suggests that higher N in wetlands correlates with a reduction in wetland soil stability,
making wetlands more susceptible to erosion.
Salt Marsh
A new study confirms N loading increases carbon dioxide production from salt marsh
soils. The new study uses an older model developed by Wigand et al. (2003). the
Nitrogen Loading Model (NLM) for wetlands and estuaries of Narragansett Bay, RI, to
determine annual N loads for different watersheds (sum of Nr from runoff, deposition,
and municipal waste, 10.3-6,727 kg N/ha/yr) within the bay. Sampling of cordgrass
marshes [Spartina alterniflora and Spartinapatens; (Wigand et al.. 2009)1 within these
watersheds showed that soil respiration of CO2 increased linearly and at similar rates with
increasing annual N load in stands of S. alterniflora and S. patens. Decreases in sediment
soil %C and %N and in belowground biomass were also observed as soil respiration
increased (Wigand et al.. 2009). Two additional experiments evaluated how N addition
affects soil respiration (Anisfeld and Hill. 2012) and general CO2 (see. Wigand et al..
2015). In both cases, N addition increased these rates; however, the amount of N was
over 1,000 kg/ha/yr, too high a level to evaluate the effects of N deposition.
In the Great Sippewissett Marsh fertilization experiment, fertilization with 520 kg N
decreased long-term carbon storage in substrates by 31% (Turner et al.. 2009). The
stability of the marsh substrate (measured as the physical resistance to force) decreased
60% in urea-only fertilized plots. In contrast, N addition to S. alterniflora marshes
located in Cocodrie, LA did not affect the physical resistance of the surface marsh
substrate, and no changes at any profile were detected with the addition of 1,200 kg
N/ha/yr (Turner. 2011). This study found negative effects upon marsh physical resistance
at high rates of N addition (2,300-18,600 kg N/ha/yr) not relevant to evaluating the
effects of N deposition.
The stability of marsh peat soils depends in part upon coarse roots and rhizomes, which
marsh plants produce for physical support and as perennial storage organs. A long-term
study at Goat Island, SC found stimulatory effects of N addition upon Spartina
alterniflora coarse roots and rhizomes, and organic matter in peat (Wigand et al.. 2015;
Davev et al.. 2011). but the experimentally added N load was 4,200 kg N/ha/yr, too high
to infer N deposition effects. Each year, marsh plants produce a new set of fine roots to
acquire nutrients from the sediment and pore water. In Kirkpatrick Marsh, vegetation
consisted of Schoenoplectus americanus, Spartina patens, and Distichlis spicata. N
addition (250 kg N/ha/yr) decreased fine root production by 42 and 84% compared to
control plots in the 3rd and 4th years of the experiment (Langlcv and Mcgonigal. 2012.
2010). When S. americanus and S. patens were planted in mesocosms for a sea level by
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nitrogen fertilization factorial experiment, results were different. There was no
belowground response in mesocosms at current sea levels to nitrogen addition (Langlev
ct al.. 2013). However, in a mesocosm mimicking a 10 cm rise in sea level, N addition
increased belowground biomass by 130%. In a mesocosm mimicking a productive marsh
elevated 15 cm above current sea level, N addition increased fine roots in mescosms by
20-40% (Langlev et al.. 2013). The long term study at Goat Island found inhibitory
effects of N addition upon fine root mass (Davcv etal.. 2011). but the experimentally
added N load was too high to infer N deposition effects.
North American salt marshes are currently being invaded by the European lineage of
Phragmites australis, which is identified as a noxious weed or banned from sale or
transport by six states (USDA. 2015b). A Kirkpatrick Marsh study compared establishing
P. australis mesoscosms to mesocosms of the native plant community described by
Langlev et al. (2013). and found that 250 kg N/ha/yr increased belowground biomass of
the invasive species (Mozdzcr et al.. 2016). The invasive species had a greater rooting
depth than the native community, and N addition increased root biomass 23-69% at
depths of 10 to 40 cm, giving P. australis a competitive advantage over the native
species. P. australis roots primed microbial activity and increased decomposition of
buried peat by 1.8-1.9 times rates in uncolonized peat (Mozdzcr et al.. 2016). creating a
positive feedback loop between increasing N and invasive species expansion. A large
scale eutrophication experiment conducted at Plum Island Estuary, MA added nitrogen
and phosphorus to tidal inflows to raise tidal nitrate concentrations to 15 times the
background N load (Deegan et al.. 2012). Although the annual nitrogen load cannot be
calculated, the results of this experiment illustrate the mechanisms by which N addition
destabilizes salt marshes. Belowground biomass decreased 33% compared to control
marshes, and this in turn altered drainage, resulting in 4% higher water content in creek
banks in enriched marshes. Creek bank with decreased stabilizing root mats and
increased water content were less stable, with increasing numbers and lengths of creek
bed fractures over time (see Table 11-3 in Appendix 11.3.2.1.6 for equation). Creek
banks were so destabilized by N addition that the number of slumps, or creek edges
sliding into creek beds, was 113% higher in enriched than control marsh, which increased
the creek channel width:depth ratio by 28%. The unvegetated area of exposed mud was
200% higher in the enriched marsh as a consequence of the marsh destabilization
(Deegan et al.. 2012).
Freshwater Tidal Marsh
In marshes, coarse roots and rhizomes are produced by plants to serve as physical
supports and as storage organs for nutrients and carbohydrates. In a freshwater tidal
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marsh dominated by Zizaniopsis miliacea in the Altamaha River Estuary, GA, addition of
500 kg N/ha/yr decreased the biomass of live rhizomes by 71%, and the mass of
macro-organic matter (living + dead roots) decreased 33% (Ketetal.. 2011). An
N-loading experiment added 50, 200, or 1,200 kg N/ha/yr to a freshwater tidal marsh on
the Tchefuncte River, LA (Graham and Mendelssohn. 2016). Researchers used two
different sampling methods to assess belowground biomass. Addition of 200 or 1,200 kg
N/ha/yr reduced biomass of established roots by 51%. Addition of 1,200 kg N/ha/yr also
increased growth of new roots into empty soil 106% over control (Graham and
Mendelssohn. 2016). This suggests that high and very high levels of N loading to tidal
freshwater marshes may destabilize existing marshes, while very high N loading may
accelerate establishment or expansion of new marshes. However, N addition in this
experiment had no measurable effect upon organic matter content, bulk density, total C,
total P, or N cycling in the top 10 cm of soil (see Appendix 11.3.1.4) (Steinmuller etal..
2016V
Intermittent Wetland
An N addition study was established on the Tibetan Plateau in an alpine meadow that is
also an intermittent wetland. N addition of 30 kg N/ha/yr increased belowground biomass
26%. N addition also stimulated CO2 uptake 25% (Wang et al.. 2017c). which offset the
concurrent stimulation of N2O emissions (see Appendix 11.3.1.5).
Bog and Fen
In the 2008 ISA, soil respiration in bogs had been studied in European countries under a
natural gradient of atmospheric N deposition from 2 to 20 kg N/ha/yr. Studies found
enhanced decomposition rates for material accumulated under higher atmospheric N
supplies resulted in higher carbon dioxide emissions. There are three new studies on an
Ontario bog and two new studies from European systems.
The Mer Bleue Bog in Ontario is the site of two concurrent fertilization experiments. In
one experiment, treatments were initiated in 2000 with nitrogen addition of 16 kg
N/ha/yr. The second experiment was initiated in 2005 in separate plots, where
nitrogen-only treatments of 32 kg N/ha/yr and 64 kg N/ha/yr were applied. Each
experiment had separate control plots. N addition at the rate of 16 kg N/ha/yr decreased
the concentrations of CO2 in peat substrate in the 8th year of N addition (2007),
decreasing CO2 by 19% at a 5-cm depth and 14% at a 25-cm depth (Wendel et al.. 201IV
N addition at the rate of 64 kg N/ha/yr did not significantly alter ecosystem respiration
when assessed 7 years after fertilization started (2011), but due to changes in
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5
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7
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38
photosynthesis and declines in cover (see Appendix 11.5.5 and Appendix 11.8.1.5). net
ecosystem exchange in the growing season was 46% lower than in unamended bog plots
(Larmolaet al.. 2013).
In the Mer Bleue ombrotrophic peat bog, 16 kg N/ha/yr added from 2000 to 2007
increased peat biomass 17% in the top 10 cm of the substrate. In addition, median
temperature at a 5-cm depth in the peat was 19% lower in plots that received
16 kg N/ha/yr than in control plots (Wendel et al.. 2011). which the authors posited was
due to increased shading by increasingly productive shrub species. In contrast, soil
temperatures were elevated deeper in the peat substrate; at 20-cm depth, average daily
temperature was 0.6°C higher in plots with 16 kg N/ha/yr added than in control plots
(Wcndcl etal.. 2011). More recently, peat samples from Mer Bleue suggest that N
addition of 16 kg N/ha/yr, along with the resultant increase in shrub production, has
increased soluble phenolics 17% in peat, with peat C:K ratio increasing 45%
(Pinsonncault et al.. 2016). indicating a shift towards more recalcitrant C in the peat as
well as K limitation of the bog (see Appendix 11.5.5). In terms of the microbial
community responsible for decomposition and respiration in the peat, N addition in all
amounts increased P and K limitation, as indicated by increasing soluble C:P and C:K
ratios and increasing activity of P and K acquiring enzymes. However, N addition also
increased microbial breakdown of carbon, as indicated by 62-97% increases in the
activity of (3-D-glucosaminidase (Pinsonncault et al.. 2016). A recently initiated addition
experiment in the same bog added 64 kg N/ha/yr as either NH4+ or NOs and found that
reduced N resulted in 15% higher soluble phenolics and suppressed enzyme activities
compared to oxidized N (Pinsonneault et al.. 2016). This result suggests that oxidized N
is more likely to enhance microbial decomposition of peat than an equivalent amount of
reduced N.
A recent study of mesocosms collected from European bogs under a natural gradient of N
deposition from 2 to 24 kg N/ha/yr showed that increasing N deposition correlated with
increased respiration (and increased gross primary productivity, and increased net C
uptake; see Appendix 11.4.5) up to 15 kg N/ha/yr, but that all C fluxes decreased by 40%
between 15 and 24 kg N/ha/yr (Estop-Aragones et al.. 2016). Importantly, N stimulation
of respiration continued under drought conditions, although there was no relationship
during drought between N deposition and GPP or net carbon uptake. Inference from this
study is limited by sample size; there was only one mesocsom per bog (Estop-Aragones
et al.. 2016). In a study of decomposition rates conducted in Eastern Europe, N deposition
altered decomposition rates in bogs. When a standard cellulose substrate was placed in all
bogs, decomposition mass loss was higher at the high N deposition (20-25 kg N/ha/yr)
Jizera Mountains site, 170% higher when placed in areas dominated by Sphagnum
rubrum, and 510% higher mass loss in areas dominated by S. magellanicum, than in the
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1	low N deposition (12.5 kg N/ha/yr) Jeseniky Mountains site (Jirousck et al.. 2015). When
2	S. rubrum samples were collected from both sites and decomposed in a common site,
3	moss collected from the high N site decomposed more slowly, with 30% less mass lost
4	than from moss collected at the low N site, indicating that N deposition increased the
5	recalcitrance of moss to decomposition (Jirousck etal.. 2015). In contrast, at the Whim
6	Bog N addition experiment, there was no effect of added N as either ammonium or nitrate
7	(treatment N loads of 16, 32, and 64 kg N/ha/yr) upon decomposition of senesced
8	Sphagnum moss (Manninen et al.. 2016). although there were effects upon living moss
9	and water quality (see Appendix 11.5.5 and Appendix 11.3.1.5).
11.3.2.1.6. Summary Table
Table 11-3 Loading effects upon belowground carbon cycling.
Additions or
Type of	Load (kg	Biological and Chemical
Ecosystem	N/ha/yr)	Effects	Study Site Study Species Reference
Estuarine N dep = not	Soil total N (Nson, as %N) 90 Gulf Coast Marsh	Nestlerode et
marsh,	reported	increased as soil bulk	wetlands, Texas community, soil al. (2014)
estuarine 3 ^p = nof	density (SBD, as g/cc) to Florida	and pore water
shrub swamp, rpnnrteH	decreased,	chemistry
palustrine	In(Nsoii) = -1.9233 *
marsh,	SBD+ 0.4165.
palustrine
swamp
Salt marsh Watershed total	In S. alterniflora stands, soil Narragansett Bare sediments Wiqand et al.
N load (10.3,	respiration (y) increased Bay, Rl	in Spartina (2009)
11.5, 27.4, 37.4,	linearly with N loading (x)	patens and
2,729.9, 3,661,	(y = 0.0006x + 2.04). Soil	Spartina
S dep = not
reported
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Salt marsh
Addition:
1,050 kg N/ha/yr
(low) as NaNC>3,
2,100 kg N/ha/yr
(medium) as
NH4NO3 or
NaNOs, 4,200 kg
N/ha/yr (high) as
NH4NO3 in same
plots in over
several years
N dep = not
reported
S dep = not
reported
Medium N increased annual
sediment respiration
(g C/m2/yr) by 53%.
Hoadley Creek
Marsh, Guilford,
CT
Spartina
alterni flora
Anisfeld and
Hill (2012)
Salt marsh
520 kg N/ha/yr in
U (N as urea
treatment), UP
(urea + 208 kg
P/ha/yr as
phosphate), and
HF (milorganite,
10-6-4 NPK)
plots, as reported
by Brin et al.
(2010)
N dep = not
reported
S dep = not
reported
Fertilization decreased the
organic:inorganic ratio by
31% in older, deeper
(pre-1963) marsh substrate,
and decreased the physical
resistance of the marsh
substrate (shear vane
strength) by 60% in U plots,
36% in UP plots, and 35%
in HF plots.
Great
Sippewissett
Marsh, MA
Spartina
alterni flora,
Spartina
patens, and
Distichlis
spicata
Turner et al.
(2009)
Coastal salt
marsh
Addition:
4,200 kg N/ha/yr
as NH4NO3 or
(NH4)2S04
N dep = not
reported
S dep = not
reported
N addition increased coarse
root volume 63% in the top
10 cm, or 61% in the top
20 cm of sediment. Organic
matter increased 20% in the
top 21 cm of sediment. CO2
emission increased 41%.
Goat Island,
(measured
2008)
SC
Spartina
alterni flora and
associated
sediments
Wiaand et al.
(2015)
Salt marsh
4,200 kg N/ha/yr
as NH4NO3
N dep = not
reported
S dep = not
reported
N addition increased peat
mass fraction by 28% and
root/rhizome mass fraction
by 58% in wet shallow
(<10 cm) sediments. Fine
root mass decreased 39%
in shallow sediments.
Goat Island, SC
(measured
2008)
Spartina
alterni flora
Davev et al.
(2011)
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Estuarine salt
marsh
250 kg N/ha/yr
N dep = not
reported
S dep = not
reported
Total fine root production
decreased by 42 and 84%
in the 3rd and 4th yr.
Kirkpatrick
Marsh, MD
(measured in
3rd and 4th yr,
2008-2009)
Schoenoplectus Lanqlev and
americanus Meaoniaal
(C3), Spartina
patens (C4),
and Distichlis
spicata (C4)
(2012. 2010)
Estuarine salt
marsh
NH4CI solution
injected into
mesocosm peat
at root-level,
250 kgN/ha,
increases N by
40% above
annual average
background
concentration
Deposition not
reported
At increased marsh
elevation (15 cm above
current sea level), N
addition increased total fine
root mass by 20-40%.
At 10 cm sea level rise, N
addition increases BG
biomass by 130% under
ambient [CO2].
Kirkpatrick
Marsh, MD
Factorial
mesocosm
experiment with
varying sea
levels: 35 or
15 cm rise in
marsh
elevation;
current sea
level; 10, 20, or
30 cm rise in
sea level.
Mesocosms
planted with
2 rhizomes
Schoenoplectus
americanus and
10 stems
Spartina
patens;
harvested each
year
Lanqlev et al.
(2013)
Salt marsh
Addition = 250 kg
N/ha/yr (as
NH4CI)
Ambient
deposition = not
reported
N addition did not
significantly affect the
rooting depth of Phragmites
australis, which was
significantly deeper than
rooting depth of native
community.
N addition increased
belowground biomass of P.
australis 37-69% between
10 and 40 cm depth, with
peak increases at 10 cm
depth.
N addition increased pore
water concentrations of NH4
by 9 and 72% at depths of
40 and 80 cm, respectively.
P. australis increased SOM
decomposition in
recalcitrant peat 1.8-1.9x
over rates in unvegetated
peat.
Mesocosms of
invasive (from
seeds) or native
(from rhzome
fragments)
plants in
Kirkpatrick
Marsh, Rhode
River,
Edgewater, MD
Invasive
Phragmites
australis; native
community of
Spartina
patens,
Schoenoplectus
americanus
Mozdzer et al.
(2016)
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Salt marsh
Nitrate and
phosphate
dissolved in tidal
inflows to raise
aqueous NO3"
concentrations to
70-100 |jM and
raise PO43" to
5-7 |jM.
Background load
in tides: 5 |jM
NO3-, 1 |JM
PO43-.
N dep = not
reported
S dep = not
reported
Marsh stability (F, as creek
bank fractures) decreases
with cumulative N
enrichment (y, in years of N
enrichment):
F = 0.42y+ 0.79. Creek
fractures were 4.4* as long
in enriched as in control
marsh.
Water content in creek bank
increases 4%.
Number of bank slumps into
creek bed is 2.1 * higher in
enriched marsh. Channel
width:depth increased 28%.
Belowground biomass
decreases 33%.
Area of unvegetated
exposed mud increased
200%.
Plum Island
Estuary, MA
Primary tidal
creeks with
Spartina
alterniflora in
low marsh
along creeks,
and Spartina
patens in high
marsh platforms
Deeqan et al.
(2012)
Freshwater
marsh
500 kg N/ha/yr as Addition of 500 kg N/ha/yr Altamaha
NH4CI or urea
N dep = not
reported
S dep = not
reported
decreased the biomass of
live rhizomes by 71%, and
the mass of macro-organic
matter (living + dead roots)
decreased 33%.
Estuary, GA
(2007, 2008)
Zizaniopsis
miliacea,
Pontederia
cord at a, and
Sagittaria
land folia
Ket et al.
(2011)
Tidal marsh
Addition = 0, 50,
200, 1,200 kg
N/ha/yr as urea
Ambient
deposition = not
reported
Addition of 1,200 kg N/ha/yr
increased root colonization
of soil 106% over control
(0 kg added N) as
evaluated by living root
biomass within ingrowth
cores (p-value = 0.03).
Addition of 200 or 1,200 kg
N/ha/yr reduced biomass of
established roots by 51% as
evaluated by sampling
standing root biomass
(p-value = 0.02).
Tchefuncte
River, Lake
Pontchartrain,
LA
Sagittaria
land folia
Graham and
Mendelssohn
(2016)
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Tidal marsh
Addition = 0, 50,
200, and
1,200 kg N/ha/yr
(as slow-release
methylene urea,
which releases
NH4+), in
combination with
0 or 131 kg
P/ha/yr, to mimic
Mississippi River
diversion N and P
loading rates
Ambient
deposition = not
specified
11 yr of N addition had no
significant effect on organic
matter content, bulk density,
total C, total P in the top
10 cm of marsh soil.
Madisonville
nutrient plots
(microtidal
pulses of
10 cm),
Tchefuncte
River, Lake
Pontchartrain
Estuary, LA
Sagittaria
lancifolia,
Polygonum
punctatum,
Eleocharis
fall as
Steinmuller et
al. (2016)
Seasonal
wetland, alpine
meadow
Addition: 30 kg
N/ha/yr
Ambient
deposition = 8.7
to 13.8 kg N/ha/yr
N addition increased
belowground biomass by
26%.
N addition increased CO2
uptake by 25%. N addition
reduced the global warming
potential (sum of CO2, CH4,
and N2O fluxes) of the plots
from a net source to a net
sink.
Mesocosms at
Luanhaizi
wetlands,
Tibetan Plateau
Carex
pamirensis,
Carex
atrofusca,
Hippuris
vulgaris,
Triglochin
palustris, and
Heleocharis
spp.
Wang et al.
(2017c)
Ombrotrophic
peat bog
16 (low), 32
(medium), or 64
(high) kg N/ha/yr
as N (NH4NO3) or
NPK(NH4N03
and KHPO4)
S dep = not
reported
N dep = 8 kg
N/ha/yr
Medium and high NPK
increased ecosystem
respiration in 8th yr by 24
and 32% respectively, and
decreased surface
temperature by 11 and
13%, respectively. In high
NPK, the water table was
42% closer to peat surface.
Mer Bleue Bog,
Ontario, Canada
Bog plant
community:
dwarf shrub
species and
mosses
Sphagnum
magellanicum,
Sphagnum
capillifolium,
and Polytrichum
strictum
Juutinen et al.
(2010)
Ombrotrophic Addition: 16 kg
peat bog	N/ha/yr as
NH4NO3
N dep = 8 kg
N/ha/yr as
quantified by
Turunen et al.
(2004)
N addition lowered
substrate temperature at
5-cm depth, decreasing
median temperature by
19% and increased
substrate temperature at
20-cm depth, by an average
of 0.6°C.
N addition decreased CO2
concentrations in substrate
by 19% at 5-cm depth and
by 14% at 25-cm depth.
N addition increased peat
biomass 17%.
Mer Bleue Bog,
Ontario, Canada
(measured
2007)
Bog plant
community:
dwarf shrub
species and
mosses
Sphagnum
magellanicum,
Sphagnum
capillifolium,
and Polytrichum
strictum
Wendel et al.
(2011)
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.

Additions or




Type of
Load (kg
Biological and Chemical



Ecosystem
N/ha/yr)
Effects
Study Site
Study Species
Reference
Ombrotrophic
16 (low), 32
Under high N, growing
Mer Bleue Bog,
Shrub species
Larmola et al.
peat bog
(medium), or 64
season net ecosystem
Ontario, Canada
(Vaccinium
(2013)

(high) kg N/ha/yr
exchange declined by 46%.

myrtilloides,


as N (NH4NO3) or


Ledum


NPK(NH4N03


groenlandicum,


and KH2PO4)


Chamaedaphne





calyculata) and





mosses





(Sphagnum





magellanicum,





Sphagnum





capillifolium,





Polytrichum





strictum)

June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Ombrotrophic
peatland
Addition = 16 kg
N/ha/yr (5N, or
5x background
deposition), 32 kg
N/ha/yr (1 ON), or
64 kg N/ha/yr
(20N), all as
NH4NO3
Also, 64 kg
N/ha/yr as NhU"1"
only (20N-NH4),
and 64 kg N/ha/yr
as NO3" only
(2ON-NO3)
Ambient
deposition = 3 to
5 kg N/ha/yr,
-50% as NO3-
16 kg N/ha/yr (5N): N
addition in this amount
increased the peat C:K ratio
45% and increased
peat-soluble phenolics 17%.
In terms of microbial
activity, N addition
decreased soluble C:N
17%, and decreased NAG
activity by 56%. N addition
increased soluble C:P 33%
and phosphatase activity
7%. N addition increased
soluble C:K45%. N addition
increased BDG activity 62%
and phenol oxidase activity
12%.
32 kg N/ha/yr (10N) and
64 kg N/ha/yr (20N): N
addition in these amounts
altered nutrient ratios of
peat: 15-23% decrease in
C:N, 15% increase in C:P
(64 kg N only), and
72-109% increase in C:K
compared to control plots.
In terms of microbial
activity, N addition
decreased soluble C:N
30-42%,	and decreased
NAG activity 72-77%. N
addition increased soluble
C:P 36-39% and
phosphatase activity
41-49%. N addition
increased soluble C:K
32-51%. N addition
increased BDG activity
80-97% but decreased
phenol oxidase activity
31-63%.
In comparing the form of N
in addition (20N-NH4+ vs.
2ON-NO3"), peat soluble
phenolics were 15% higher
under NH4+ than under NO3"
. Activity of all enzymes was
lower under NHV compared
to NO3" addition: BDG
decreased 11%, NAG
decreased 32%,
phosphatase decreased
7%, and phenol oxidase
decreased 56%.
Mer Bleue bog,
Ottawa, Canada
Nonliving peat
in top 10 cm of
bog
Pinsonneault
etal. (2016)
June 2018
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Table 11-3 (Continued): Loading effects upon belowground carbon cycling.
Type of
Ecosystem
Additions or
Load (kg
N/ha/yr)
Biological and Chemical
Effects
Study Site Study Species Reference
Ombrotrophic
peatland
Ambient
deposition of
NOx, NHy, and
SOx:
1.95-24.07 kg
N/ha/yr and
3.18-36.90 kg
S/ha/yr
Deposition
estimates from
EMEP (European
Monitoring and
Evaluation
Programme)-
based IDEM
(Integrated
Deposition
Model) from
Pieterse et al.
(2007)
N and S deposition rates
are correlated across these
sites (r= 0.87).
Respiration (CresP, as g
C/m2/day) was positively
correlated with N deposition
(Ndep, as kg N/ha/yr) up to
15 kg N/ha/yr: CresP
= 0.448 + 0.067 x NdeP;
R2 = 0.61. The relationship
between N and respiration
was weaker during drought
(R2 = 0.38) and
post-drought (R2 = 0.50)
periods.
Respiration, GPP, and net
carbon uptake were all 40%
lower in the mesocosm from
the 24 kg N/ha/yr than from
the 15 kg N/ha/yr site.
Lab incubations
of 14 Sphagnum
mesocosms
collected in
saturated
hollows in
European bogs:
U.K. and Ireland
(n = 9), Poland
(n = 4) and
Slovakia (n = 1)
Mesocsoms
were maintained
at a high water
table (0-5 cm in
depth) for
40 days, then
were subject to
drought for
100 days, then
were re wetted to
high water table
(0-5 cm in
depth) for
200 days.
Sphagnum
fall ax
Estop-
Araaones et
al. (2016)
Ombrotrophic
bog
Wet + dry
deposition,
estimated by
Jirousek et al.
(2011)
Jireza: 20-25 kg
N/ha/yr (high N)
Jeseniky: 12.5 kg
N/ha/yr (low N)
Decomposition of a
standard substrate
(cellulose) is higher at high
N site, 170% higher mass
loss in S. rubellum zone
and 510% higher mass loss
in S. magellanicum zone.
In comparing decomposition
of S. rubellum grown at high
N or low N sites, S.
rubellum from high N mass
loss was 30% lower
(i.e., less decomposition).
Two sites:
Jizerka in Jireza
Mountains
(warm
suboceanic
climate) and
Vozka in
Jeseniky
Mountains (2°C
colder)
Sphagnum
fall ax,
Sphagnum
magellanicum,
and Sphagnum
rubellum/
russowii
Jirousek et al.
(2015)
Ombrotrophic
peatland
Addition = 8, 24,
and 56 kg N/ha/yr
as either NH4+ or
NOs"
Ambient
deposition = 8 kg
N/ha/yr
Decomposition rates of
Sphagnum were not
affected by elevated N
levels. Only the highest
level of N addition altered
pore water chemistry.
Whim bog,
Edinburgh,
Scotland
Sphagnum
moss
(Sphagnum
capillifolium)
Manninen et
al. (2016)
ANPP = aboveground net primary productivity; C = carbon; cm = centimeter; C02 = carbon dioxide; dep = deposition; g = gram;
ha = hectare; kg = kilogram; m = meter; N = nitrogen; NaN03 = sodium nitrate; NH4CI = ammonium chloride; NH4N03 = ammonium
nitrate; (NH4)2S04 = ammonium sulfate; NPK = nitrogen, phosphorus, potassium; S = sulfur; yr = year.
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11.3.2.2. Methane Emissions
Methane (CH4) is an important greenhouse gas that is over 20 times more effective at
trapping heat than carbon dioxide. The primary biological source of methane is microbial
(methanogens in the domain Archaea), as is the primary biological sink (methanotrophs
among the Bacteria and Archaea). N addition can stimulate methane flux from wetlands
by increasing methanogen activity when dissolved labile organic C is abundant (Kim et
al.. 2015b). or by decreasing methane oxidation by methanotrophs, as was recently
observed in a boreal peatland (Lozanovska et al.. 2016). Therefore, understanding the
controls on these microorganisms is important for predicting methane flux from
ecosystems. In terms of carbon emissions, the 2008 ISA reported that N deposition alters
CH4 flux in wetland and forested ecosystems. A meta-analysis of 25 N addition
observations (N addition 30 to 240 kg N/ha/yr) found that N addition increased CH4
emissions by 95% from wetlands and grasslands [see Table 11-4 for other nonsignificant
results; (Liu and Greaver. 2009)1.
There is new evidence that N loading increases methane production in soils. In the N
addition experiments replicated in three California marshes (Vivanco et al.. 2015; Irvine
et al.. 2012). field-measured methane flux from soils increased linearly with increasing N
load (0, 100, 200, 400, 800, 1,600, 3,200 kg N/ha/yr), such that CH4 flux from soils
increased by 1.23 |ig CH4/m2/day for each 10 kg N/ha/yr added R2 = 0.23 and P = 0.025
(Vivanco et al.. 2015; Irvine et al.. 2012). Methane flux increased from low or nearly
negative to net positive values above -100 kg N/ha/yr, showing that the effect of added N
on methanogenesis, the final step in anaerobic decomposition, offset any increase in
methanotrophy. Differential effects of N on both methanogenesis and methanotrophy
have been documented in other ecosystems (Irvine et al.. 2012; Liu and Greaver. 2009).
The linearity of the trend became less robust when the exposure time increased to
14 months.
Soils were collected from the Tijuana River Reserve near the sites used in Vivanco et al.
(2015) and incubated in a laboratory microcosm experiment with factorial C and N
additions (Irvine et al.. 2012). Adding C and N together increased methane production by
44% above controls (Irvine et al.. 2012). suggesting that N loading will increase
microbial methane production disproportionally when accompanied by pulses of labile C.
The authors concluded that temperate salt marshes generally emit low levels of methane,
but these values are also highly spatially and seasonally variable (Cheng et al.. 2010;
Magenheimer et al.. 1996; Bartlett et al.. 1985; King and Wiebe. 1978). Despite the high
variability observed, however, the field experiments suggest that increased N availability
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stimulated methanogenesis and increased methane emissions in southern California salt
marshes, as in other ecosystems (Liu and Greaver. 2009).
A study of a salt marsh in the Yellow River Delta, China, added N loads in different
chemical forms to the marsh to determine the effects of N form upon microbial
communities involved in C cycling. N was added as 50 kg N/ha/yr as NH44", 50 kg
N/ha/yr as NOs . or 100 kg N/ha/yr as NH4NO3. All three forms of N increased the
abundance of archaeal methanogens as well as the abundance of bacteria that form
syntrophies that stimulate methanogen metabolism (Xiao et al.. 2017). Ordination of
bacterial community composition suggested that all forms of N shifted the microbial
community in the same direction, but that reduced N had much stronger effects on
community composition than did oxidized N. N form also affected methane production
and fluxes: all N forms increased methane in pore water, but only NH44" and NH4NO3
additions increased methane emissions from the plots (Xiao et al.. 2017). Reduced forms
of N have stronger effects on microbial C cycling than do equivalent amounts of oxidized
Table 11-4 Nitrogen loading effects upon methane emissions.
N.
11.3.2.2.1. Summary Table
Type of Additions or Load
Ecosystem (kg N/ha/yr)
Biological and Chemical
Effects
Study
Study Site Species
Reference
Wetlands N addition
No effect across wetlands Global
(n = 6) of N addition on net
ecosystem exchange.
N addition increased CH4
emissions by 95% across
grass + wetland + anaerobic
agricultural systems (n = 25).
No effect across drained
wetlands of N addition (n = 6)
upon biological CH4 uptake.
Meta-	Liu and Greaver
analysis of (2009)
data
collected
from North
America,
South
America,
Europe, and
Asia
(natural experiments, N
and	addition of 10 to
agricultural) 562 kg N/ha/yr
Deposition = not
reported
Salt marsh 100,200,400,800,
Methane flux (y, as mg
CH4/m2/day) increased
linearly with N addition (A/add,
as g N/m2/yr):
y= 0.00123/Vadd- 0.0122.
Morro Bay, Marsh	Irvine et al. (2012)
Carpinteria dominated by
1,600, and 3,200 kg
N/ha/yr as granular
urea
Salt Marsh,	Salicornia
Tijuana	depressa
River	(formerly
Reserve,	Salicornia
CA	virginica)
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Table 11-4 (Continued) Nitrogen loading effects upon methane emissions.
Type of
Additions or Load
Biological and Chemical

Study

Ecosystem
(kg N/ha/yr)
Effects
Study Site
Species
Reference
Salt marsh
0, 100, 200, 400,
CH4 flux from soils increased
Morro Bay
Stands of
Vivanco et al.

800, 1,600,
by 1.23 |jg CH4/m2/day for
National
Salicornia
(2015)

3,200 kg N/ha/yr as
each 10 kg N/ha/yr added
Estuary,
depressa


urea

Carpinteria
(formerly


N dep = 3-5 kg

Salt Marsh
Salicornia


N/ha/yr as reported

Reserve,
virginica)


in Tonnesen et al.

Tijuana



(2007)

River




Reserve
Estuary,
CA


Salt marsh
Addition of 50 kg
reduced N/ha/yr as
NH4CI, 50 kg
oxidized N as
KNO3, or 100 kg
N/ha/yr as NH4NO3
Ambient
deposition = not
specified
Addition of any form of N
increased bacterial
abundance of Geobacillus and
Clostridium, and increased
archaeal abundance of
Methanocellaceae.
Reduced-N and NH4NO3
additions decreased bacterial
abundance of Flavobacterium,
Bacillus, Gillisia,
Marinobacter, and
Desulfosarcina.
Redundancy analysis found
positive correlations between
ammonium-N measured in
sediment, methane flux, pore
water methane
concentrations, Geobacillus
abundance, Clostridium
abundance, and
Methanocellaceae
abundance.
In dry season, only 50 kg
reduced N addition plots are
methane sources; all other
plots are methane sinks.
In wet season, 50 kg oxidized
N increased pore water Chk
50 kg reduced N and 100 kg N
increased pore water CH4 as
well as CH4 emissions from
plots.
Marsh of
Phragmites
australis
and
Suaeda
heteroptera
in the
Yellow
River
Delta,
China
Bacterial and
archaeal
communities
assessed by
bacterial
primers
Ba338fand
Ba806r, and
archaeal
primers
Ar515f and
Ar907r
Xiao etal. (2017)
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11.4.
Production and Aboveground Biomass
Aboveground biomass is a measure of how much carbon is fixed by wetland plants.
Changes in this endpoint can affect food webs within the wetland, including dependent
migratory species. Changes in aboveground biomass will also be important for food webs
in rivers, lakes, or estuaries connected by surface water flow, because plant litter from
wetlands is an important base for food webs downstream. In the 2008 ISA, evidence from
Canadian and European peatlands showed that N deposition had negative or mixed
effects on Sphagnum productivity, depending on history of deposition. In Canadian
ombrotrophic peatlands experiencing deposition of 2.7-8.1 kg N/ha/yr, peat
accumulation increased with N deposition, but accumulation rates had slowed by 2004,
indicating a degree of N saturation. Coastal wetlands responded to N enrichment with
increased primary production, shifting microbial and plant communities and altering pore
water chemistry, although many of the studies in coastal wetlands used N enrichment
levels more similar to those of wastewater than atmospheric deposition. The new studies
published from 2008-2015 include work on tidal marsh, mangroves, and ombrotrophic
bogs.
11.4.1. Salt Marsh
In the 2008 ISA, primary production of plant species in coastal wetlands typically
increased with N addition; however, most studies applied fertilizer treatments that were
several orders of magnitude larger than atmospheric deposition (Darby and Turner. 2008;
Tyler etal.. 2007; Wigand et al.. 2003; Mendelssohn. 1979). N fertilization experiments
in salt marsh ecosystems show biomass stimulation from 6 to 413% with application rates
ranging from 7 to 3,120 kg N/ha/yr (U.S. EPA. 1993).
A number of new studies have evaluated N effects on production and biomass in coastal
wetlands. In Elkhorn Slough, the largest coastal marsh in California, addition of 150 kg
N/ha/yr increased S. pacifica biomass in two of the six experimental sites (Goldman
Martone and Wasson. 2008); a 36% increase over unamended plots occurred in a site
with unrestricted tidal flow, and a 53% increase occurred in a marsh where
impoundments restricted regular tidal exchange. Darby and Turner (2008) reported on a
nitrogen addition experiment at Cocodrie, LA, in which five loads of nitrogen were
applied to S. alterniflora marshes. Aboveground biomass increased linearly with
increases in N load, although the regressions were not reported. Additions of 230 and
465 kg N/ha/yr did not significantly raise aboveground (AG) biomass above unamended
marsh biomass, but 930 kg N/ha/yr increased AG biomass 122%, 1,860 kg N/ha/yr
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increased AG biomass 124%, and 3,720 kg N/ha/yr increased AG biomass 141%. A
nitrogen addition experiment with seven different N loads (0, 100, 200, 400, 800, 1,600,
and 3,200 kg N/ha/yr) was replicated at three different California estuarine reserves
[Morro Bay, Carpinteria Salt Marsh, and Tijuana River Marsh Reserve; (Vivanco ct al..
2015)1. Aboveground biomass of existing S. depressa increased linearly in response to N
addition across all three marshes and N addition levels starting at 100 kg N/ha/yr (see
Table 11-5 for equation, R2 = 0.58). When harvested plots were resampled, the regrowth
of aboveground biomass increased in a saturating response to increasing N load, with no
significant increases in biomass in response to N above 1,600 kg N/ha/yr (Table 11-5 for
equation). Similar saturation of plant response occurred in N tissue (see Appendix 11.5).
In the eutrophication experiment at Plum Island, MA, tidal nitrate enriched to 15 times
background concentrations, along with P additions, increased aboveground shoot biomass
and specific mass of the vascular plant community by 16 and 9%, respectively (Deeganet
al.. 2012). When the tidal N enrichment was repeated in later years without P additions,
added loads of 620 kg N/ha/yr or 1,200 kg N/ha/yr increased low marsh Spartina
alterniflora shoot specific mass 3 and 12%, respectively (Johnson et al.. 2016a). In the
high marsh at this site, added N loads were smaller: at 140 kg N/ha/yr, S. alterniflora and
Distichlis spicata increased shoot-specific mass 14 and 8% (Johnson et al.. 2016a). There
were also changes in plant architecture in the same plots (see Appendix 11.6.1). but no
effects on plant community. The Kirkpatrick Marsh in Maryland is the site of a nitrogen
fertilization experiment in a tidal marsh community consisting of Schoenoplectus
americanus, Spartina patens, and Distichlis spicata. Over the 4 years of the experiment,
nitrogen addition of 250 kg N/ha/yr increased total aboveground biomass by 47-79%
(Langlcv and Mcgonigal. 2010; Langlev et al.. 2009). When the biomass response was
partitioned by carbon fixation mechanism (C3 or C4 plants), biomass of C3 plants
declined with N addition to 81% in the 3rd year and 55% in the 4th year of initial
fertilized S. americanus biomass. Biomass of C4 plants S. patens and D. spicata
increased 2-51 times the control over the 4 years of the experiment (see
Appendix 11.8.1). or increased 129% in the 4th year over initial fertilized C4 biomass. In
a related experiment at Kirkpatrick Marsh, mesocosms were constructed to mimic
marshes at different sea levels, planted with Schoenoplectus americanus and Spartina
patens, and subjected to the same nitrogen treatment (Langlev et al.. 2013). In this
experiment, sea level had the strongest impact on biomass. In mesocosms set at the
current sea level, N addition increased total AG biomass by 25-75%. The effects of N
addition were stronger at other sea levels, but responses at those marsh surface elevations
were species-specific. In a scenario of a 10-cm sea level rise, N addition increased total
AG biomass 85-570% above biomass of similarly inundated unfertilized mesocosms.
This strong response was driven almost entirely by the response of S. americanus, which
at a 10-cm sea level rise increased its biomass 150-533% in response to N addition. In
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mesocosms that mimicked a 15- or 35-cm increase in marsh elevation above sea level (as
can occur in a highly productive marsh where productivity + accretion >
decomposition + export), total AG biomass increased only 15-95%. This community
response was driven by the response of S. patens, which at a 15 - to 35-cm marsh
elevation increase responded to nitrogen addition with 55-145% increases in AG
biomass (Langlev et al.. 2013). These responses reflect the tolerance of these plants for
inundation by tides and suggest that N addition may increase the productivity only of
plants growing squarely within the range of their physiological tolerance.
Several other studies have evaluated N addition levels over 1,000 kg N/ha/yr; in all cases,
N addition increased biomass, but the high levels of experimental N addition
exponentially exceed N loading from deposition, limiting the scope of inference from this
literature (Anisfeld and Hill. 2012; Nelson and Zavaleta. 2012; Ryan and Bover. 2012).
11.4.2. Mangrove
Mangrove ecosystems are ecologically and economically important because they provide
habitat not just to a diversity of resident species but also serve as nurseries to the juvenile
lifestages of many marine fish species. As slow-growing tree species, mangroves are
rarely destructively sampled in order to estimate productivity and aboveground biomass.
The mangroves at Indian River Lagoon, FL are the site of several nitrogen addition
experiments. Whigham et al. (2009) added 100 kg N/ha/yr to plots of dwarf Avicennia
germinans, black mangrove, which increased productivity by increasing the number of
new branches 150% above control plots.
In a long-term (since 1997) N addition experiment that added approximately
11,200 kg N/yr to each tree, growth rate of Rhizophora mangle, quantified by measuring
shoot elongation rates increased 290% in the shoreline fringe zone and increased 1,340%
in the interior scrub zone (Feller et al.. 2009). At Merritt Island National Wildlife Refuge,
N addition of 1,400 kg N/ha/yr in the ecotone where mangroves and cordgrasses grow in
competition increased A germinans seedlings' production of new leaves by 42%, which
increased total leaf biomass by 72% over unamended trees (Simpson et al.. 2013) and
altered above- and below-ground stoichiometry (see Appendix 11.5.2).
11.4.3. Freshwater Tidal Marsh
In a freshwater tidal marsh on the Tchefuncte River, LA, Graham and Mendelssohn
(2010) conducted an N loading experiment that added 50, 200, or 1,200 kg N/ha/yr to a
plant community dominated by Sagittaria lancifolia, Eleocharis fallax, and Persicaria
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punctata (formerly Polygonum punctatum). The aboveground net primary productivity of
the community increased with increasing N in a negative quadratic function, with ANPP
not increasing above 200 kg N/ha/yr (Graham and Mendelssohn. 2010).
At the Altamaha River Estuary, GA, N addition experiments were conducted in tidal
freshwater marshes dominated by giant cutgrass Zizaniopsis miliacea. Addition of 500 kg
N/ha/yr increased aboveground biomass 1.4-3.8 times control biomass in the 2nd
through 5th years of fertilization (Ket et al.. 2011; Frost et al.. 2009). A nitrogen
experiment conducted in estuarine marshes and swamps on the Nanticoke River in the
Chesapeake Bay found that while an added load of 670 kg N/ha/yr did not significantly
change total herbaceous plant community aboveground biomass, it did increase biomass
of the invasive Typha spp. by 47% (Baldwin. 2013). There are several other new studies
on the effects ofN loading on aboveground biomass of freshwater tidal ecosystems;
however, the addition rates are greater than 500 kg N/ha/yr.
11.4.4. Intermittent Wetland
An N addition study was established in an alpine meadow, which is also an intermittent
wetland on the Tibetan Plateau. N addition of 30 kg N/ha/yr increased ANPP of the
graminoid and herbaceous plant community by 11.5% (Wang et al.. 2017c). which also
affected C and N fluxes of the system (see Appendix 11.3).
11.4.5. Bog and Fen
In Sphagnum-dominated ombrotrophic bogs, higher N deposition resulted in higher tissue
N concentrations and greater NPP (Aldous. 2002a). but lower bulk density. A study of
23	ombrotrophic peatlands in Canada with deposition levels ranging from 2.7 to 8.1 kg
N/ha/yr showed peat accumulation increasing linearly with N deposition; however, in
later years of the study, the rate had begun to slow, indicating limited capacity for N to
stimulate accumulation (Turunen et al.. 2004). A study of European bogs under a
deposition gradient of 2 to 24 kg N/ha/yr found that gross primary productivity and net
carbon uptake of Sphagnum fallax mesoscosms were positively correlated with increasing
N deposition up to 15 kg N/ha/yr, but that the mesocosm collected from a bog receiving
24	kg N/ha/yr had 40% lower GPP and net C uptake than the mesocosm receiving 15 kg
N/ha/yr (Estop-Aragones et al.. 2016). However, there was N stimulation of GPP and net
C uptake at deposition levels below 15 kg N/ha/yr when mesocsoms were subject to
drought conditions. Inference from this study is limited by sample size; there was only
one mesocsom per bog (Estop-Aragones et al.. 2016).
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In the Mer Bleue ombrotrophic peat bog, 8 years of N addition at the rate of
16 kg N/ha/yr increased biomass of mosses (Sphagnum magellanicum, Sphagnum
capillifolium, and Polytrichum strictum) by 30% of moss biomass in control plots (Xing
ctal.. 2011).
In oligotrophic bogs in Gogebic County, MI, N addition of 60 kg N/ha/yr increased total
plant community productivity by 82% over unamended plot productivity, resulting in
2.6 times as much biomass in fertilized plots as biomass in unamended plots (Iversenet
al.. 2010). The plant community trends reflected the responses of the dominant vascular
shrub Chamaedaphne calyculata, which increased its productivity by 87%, thus
increasing its biomass in N addition plots 105% over its biomass in control plots (Iversen
et al.. 2010V
In general, vascular plants are able to respond more rapidly to N deposition than are moss
species. A stable isotope N addition study in Bourtanger Moor, Germany, showed that
grass Lolium multiflorum and sedge Eriophorum vaginatum responded to increasing N
addition of approximately 30-55 kg N/ha/yr with linear increases in aboveground
biomass (Hurkuck et al.. 2015). Plant interception of N deposition also varied between
species and with size for L. multiflorum (see Appendix 5).
Relative changes in biomass can lead to plant community changes. Model results from
Logofet and Alexandrov (1984) suggest 7 kg N/ha/yr is the threshold for an oligotrophic
bog to become a mesotrophic fen dominated by trees, as found in the 1993 Oxides of
Nitrogen AQCD. In freshwater-rich fens in Gogebic County, MI, N addition of
60 kg N/ha/yr increased productivity of dominant grass Calamagrostis canadensis, which
increased the species' biomass to 600% over its biomass in unamended fen plots (Iversen
et al.. 2010).
In many bogs and fens, N addition increases biomass and then abundance of vascular
plants and N-tolerant moss species, while N-sensitive species decline. A study of purple
pitcher plant, Sarracenia purpurea, growing in bogs across a deposition gradient of
3.4-5.0 kg N/ha/yr in the Adirondack Mountains, NY, identified a threshold for plant
response to N (Crumley et al.. 2016). with plant growth depressed 8-21% at deposition
levels above 4.1 kg N/ha/yr, and with other negative effects on plant physiology (see
Appendix 11.5.5). In a European mesocosm study of four moss species, elevated N
concentrations typical of rainwater in the fen depress growth rates 18% compared to
mosses growing in N concentrations typical of the fen groundwater (Andersen et al..
2016). Additional experimental N loading further decreases growth rate of all four
species. Two rare moss species, with declining abundance over the last 100 years, show
high sensitivity to N, with more growth rate variation explained by N addition (18-23%)
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than for the growth rate in the more common moss species (5-7%) (Andersen ct al..
2016V
Recent research has addressed differential responses of bogs and fens to reduced or
oxidized forms of N addition. In an Irish fen dominated by peat moss Sphagnum
contortum and brown moss Scorpidium revolvens, 50 kg N/ha/yr was applied as NOs, in
one treatment and as NH4+ in another treatment (Paulisscn et al.. 2016). Reduced N
decreased S. revolvens biomass 67% compared to control plots, while oxidized N had no
measurable effect on S. revolvens biomass. S.contortum biomass was not affected by
reduced or oxidized N additions, although the plant's physiology did change in response
to both treatments (see Appendix 11.5.5).
11.4.6. Summary Table
Table 11-5 Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Coastal high
marsh
Addition: 150 kg
N/ha/yr as urea
S dep = not
reported
N dep = not
reported
At two sites with
unrestricted and restricted
tidal flow, fertilization
increased S. pacifica
biomass by 36 and 53%,
respectively.
Six sites at
Elkhorn Slough,
Watsonville, CA
Marsh community
dominated by
Sarcocornia
pacifica, also
contained
Jaumea carnosa,
Frankenia salina,
Spergularia
salina, Distichlis
spicata, and
Atriplex
caiifornica/
triangularis
Goldman
Martone and
Wasson
(2008)
Estuarine salt
marsh
Addition: 100, 200,
400, 800, 1,600,
3,200 kg N/ha/yr as
urea
N dep = 3-5 kg
N/ha/yr as reported
in Tonnesen et al.
(2007)
Aboveground biomass
(AGB) increases in a
linear response to N
addition (Nadd, as g
N/m2/yr), AGB = 0.001 *
Nadd + 1.117 (R2 = 0.58),
while biomass regrowth
(AGR) increases in a
saturating response to N,
AGR =
-1.16 x e("001Nadd' + 1.89.
Morro Bay
National
Estuary,
Carpinteria Salt
Marsh Reserve,
Tijuana River
Reserve
Estuary, CA
Salicornia
depressa
(Salicornia
virginica) stands
Vivanco et al.
(2015)
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Salt marsh
Nitrate and
phosphate
dissolved in tidal
inflows to raise
aqueous NO3"
concentrations to
70-100 |jM and
raise PO43" to
5-7 |jM.
Background load in
tides: 5 |jM NO3",
1 |jM PCM3".
N dep = not
reported
S dep = not
reported
Vascular plant shoot
biomass increased 16%,
shoot specific mass
increased 9%.
Plum Island
Estuary, MA
Primary tidal
creeks with
Spartina
alterniflora in low
marsh along
creeks, and
Spartina patens in
high marsh
platforms
Deeaan et al.
(2012)
Salt marsh
Addition of nitrate
dissolved in
incoming tide
2011-2012,
addition of N at
70-100 |jM NaNOs"
in tide for added
load of
620-1,200 kg
N/ha/yr in low
marsh and
70-140 kg N/ha/yr
in high marsh
Ambient
deposition = not
specified
Low marsh S. alterniflora
in enriched creeks
received an additional N
load of 1,200 kg N/ha/yr
in 2011. That year, N
addition increased
shoot-specific mass 12%
above S. alterniflora in
reference creek plots.
Low marsh S. alterniflora
in enriched creeks
received an additional N
load of 620 kg N/ha/yr in
2012. That year, N
addition increased
shoot-specific mass 3%.
High marsh S. alterniflora
and D. spicata in enriched
creeks received an
additional N load of
140 kg N/ha/yr in 2011. N
addition increased S.
alterniflora shoot-specific
mass 14%. In D. spicata,
N addition increased
shoot-specific mass 8%.
Plum Island
Sound Estuary,
MA
Spartina
alterniflora,
Distichlis spicata,
and Spartina
patens (high
marsh); Spartina
alterniflora (low
marsh)
Johnson et al.
(2016a)
Estuarine
marsh
Addition: 1,337 kg
N/ha/yr as urea.
S dep = not
reported
N dep = not
reported
In S. pacifica, biomass
increased 54-185%
across habitats.
China Camp
State Park, CA
Sarcocornia
pacifica dominant
(C3 succulent
shrub), Distichlis
spicata (C4
grass), and
Jaumea carnosa
(C3
semisucculent
forb)
Ryan and
Bover (2012)
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Coastal salt Addition: 3,000 kg AG biomass increased 28 Elkhorn Slough, Sarcocornia
Nelson and
marsh	N/ha/yr as NH4NO3
S dep = not
reported
N dep = not
reported
and 216% in successive
summers, and increased
shoot-to-root ratio 249%.
Monterey Bay,
CA
pacifica dominant, Zavaleta
Distichlis spicata, (2012)
Frankenia salina,
and Jaumea
carnosa
Coastal salt
marsh
230, 465, 930,
1,860, or 3,720 kg
N/ha/yr as
(NH4)2S04
S dep = not
reported
N dep = not
reported
AG biomass increased
linearly with N load
(regression not given).
Aboveground live
biomass increased 122,
124, and 141% in
response to 930, 1,860,
and 3,720 kg N.
LUMCON,
Cocodrie, LA
Spartina
alterniflora
Darby and
Turner (2008)
Coastal salt
marsh
1,050 kg N/ha/yr
(low) as NaNC>3,
2,100 kg N/ha/yr
(medium) as
NH4NO3 or NaNOs,
4,200 kg N/ha/yr
(high) as NH4NO3 in
same plots in over
several years
S dep = not
reported
N dep = not
reported
ANPP increased by 132%
in low, 130% in medium,
and 120% in high N
treatments.
Hoadley Creek
Marsh, Guilford,
CT
Spartina
alterniflora
Anisfeld and
Hill (2012)
Estuarine salt Addition: 250 kg
marsh	N/ha/yr
S dep = not
reported
N dep = not
reported
S. americanus
aboveground biomass
decreased by 19 and
45% in the 3rd and 4th yr,
respectively.
Biomass of S. patens and
D. spicata increased
129% in the 4th yr over
initial fertilized C4
biomass.
Kirkpatrick
Marsh, MD
(measured in
3rd and 4th yr,
2008-2009)
Schoenoplectus
americanus (C3),
Spartina patens
(C4), and
Distichlis spicata
(C4)
Lanqlev and
Meqoniqal
(2012. 2010)
Estuarine salt Addition: 250 kg
marsh	N/ha/yr
S dep = not
reported
N dep = not
reported
In the first two growing
seasons, fertilization
increased aboveground
biomass in the second
season by 57%.
Kirkpatrick
Marsh, MD
(measured in
1st and
2nd years,
2006-2007)
Schoenoplectus
americanus,
Spartina patens,
and Distichlis
spicata
Lanqlev et al.
(2009)
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Estuarine
marsh
NH4CI solution
injected into
mesocosm peat at
root level,
250 kgN/ha,
increases N by 40%
above annual
average
background
concentration
Deposition not
reported
At current sea level, N
addition increased total
AG biomass by 25-75%
in 1st yr.
With sea level rise 10 cm,
S. americanus AG
biomass increases
150-230% in 1 st yr,
110-533% in 2nd yr.
Total AG biomass at
10 cm increased 85-95%
in 1st yr, 100-570% in
2nd yr.
At increased marsh
elevation (15 and 35 cm
above sea level 1st yr,
15 cm 2nd yr), S. patens
AG biomass increases
55-145% in 1 st yr,
60-90% in 2nd yr. Total
AG biomass at these
elevations increased
15-95% in 1st yr,
30-35% in 2nd yr.
Kirkpatrick
Marsh, MD
Factorial
mesocosm
experiment with
varying sea
levels: 35- or
15-cm rise in
marsh elevation;
current sea level;
10-, 20-, or 30-cm
rise in sea level.
Mesocosms
planted with
2 rhizomes
Schoenoplectus
americanus and
10 stems Spartina
patens', harvested
each year.
Lanalev et al.
(2013)
Mangrove/
marsh
ecotone
Addition: 1,400 kg
N/ha/yr
S dep = not
reported
N dep = not
reported
Mangrove leaf production
increased by 42% and
leaf biomass increased by
72%.
Merritt Island Avicennia
National	germinans
Wildlife Refuge,
FL
(Impoundment
T9)
Simpson et al.
(2013)
Mangrove
Addition: 1,400 kg
N/ha/yr
S dep = not
reported
N dep = not
reported
Fertilization increases
growth rate (shoot
elongation) by 290% in
the fringe zone and by
1,340% in the scrub zone.
Indian River
Lagoon, FL
(impoundment
MI23)
Rhizophora
mangle
Feller et al.
(2009)
Mangrove
100 kg N/ha/yr
N addition increased the
number of new branches
150% in Avicennia.
Indian River
Lagoon, FL
(impoundment
SLC-24)
Avicennia
germinans and
associated
sediments
Whiqham et
al. (2009)
Estuarine
Addition: 670 kg
Typha spp. biomass
Nanticoke
Plant community
Baldwin
tidal marsh
N/ha/yr
increased by 95%.
River, MD and

(2013)

S dep = not

DE



reported





N dep = not





reported




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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects	Study Site Study Species Reference
Freshwater
50, 200, or 1,200 kg
Community ANPP
Tchefuncte
Oligohaline plant
Graham and
estuarine
N/ha/yr as
increased with medium N
River,
community
Mendelssohn
marsh
Nutralene
load (Nadd as kg N/ha/yr),
Madisonville,
dominated by
(2010)

methylene urea
ANPP = -0.00165 x Nadd2
LA
Sagittaria


S dep = not
reported
+ 2.5091 x Nadd + 1270.3.

lancifolia,




Eleocharis fallax,


N dep = not
reported


and Polygonum
punctatum

Freshwater
500 kg N/ha/yr as
N addition increased
Altamaha River,
Zizaniopsis
Frost et al.
tidal marsh
NH4CI or urea
aboveground biomass by
140% and plant height by
32% in Z. miliacea. N
addition decreased leaf
GA
miliacea
(2009)
[N] by 99% and leaf [P] by
25%.
Freshwater
marsh
500 kg N/ha/yr as
NH4CI or urea
N dep = not
reported
S dep = not
reported
In Z. miliacea,	Altamaha	Zizaniopsis
aboveground biomass	Estuary, GA miliacea,
increased by 2.9-3.8x the Pontederia
control, as leaf number cordata, and
increased 52% and plant Sagittaria
height increased 25-40%. lancifolia
Ket et al.
(2011)
Seasonal Addition: 30 kg N addition increased	Mesocosms at Carex pamirensis, Wang et al.
wetland, N/ha/yr	ANPP 11.5% across	Luanhaizi	Carex atrofusca, (2017c)
alpine	Ambient	average (3 cm above soil wetlands,	Hippuris vulgaris,
meadow deposition = 8 7 to surface) and lowered	Tibetan Plateau Triglochin
13 8 kg N/ha/yr (^0 cm below soil	palustris, and
surface) water table	Heleocharis spp.
levels.
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Ombrotrophic
peatland
Ambient deposition
of NOx, NHy, and
SOx: 1.95-24.07 kg
N/ha/yr and
3.18-36.90 kg
S/ha/yr
Deposition
estimates from
EMEP (European
Monitoring and
Evaluation
Programme)-based
IDEM (Integrated
Deposition Model)
from Pieterse et al.
(2007)
N and S deposition rates
are correlated across
these sites (r = 0.87).
Gross primary
productivity (GPP, as g
C/m2/day) increased with
N deposition (Ndep, as kg
N/ha/yr) up to 15 kg
N/ha/yr: GPP= 1.262
+ 0.219NdeP; R2 = 0.65.
There was no relationship
between N deposition and
GPP during drought and
post-drought periods.
Net carbon uptake by
Sphagnum (Cuptake, as g
C/m2/day) was positively
correlated with N
deposition (Ndep, as kg
N/ha/yr) up to levels of
15 kg/ha/yr:
Cuptake—0.76 + 0.155xNdep,
R2 = 0.52. There was no
relationship between N
deposition and GPP
during drought and
post-drought periods.
Respiration, GPP, and
net carbon uptake were
all 40% lower in the
mesocosm from the 24 kg
N/ha/yr than from the
15 kg N/ha/yr site.
Lab incubations
of
14 Sphagnum
mesocosm
collected in
saturated
hollows in
European bogs:
U.K. and
Ireland (n = 9),
Poland (n = 4),
and Slovakia
(n = 1).
Mesocsoms
were
maintained at a
high water table
(0-5 cm depth)
for 40 days,
then were
subject to
drought for
100 days, then
were re wetted
to high water
table (0-5 cm
depth) for
200 days.
Sphagnum fallax
Estop-
Araqones et
al. (2016)
Ombrotrophic 16 (low), 32
peatbog (medium), or64
(high) kg N/ha/yr as
N (NH4NO3) or NPK
(NH4NO3 and
KHPO4)
In high NPK, shrub
biomass increased, with
40% higher leaf biomass,
and 86% higher woody
biomass.
Mer Bleue Bog, Bog plant
Juutinen et al.
Ontario,	community: dwarf (2010)
Canada	shrub species
and mosses
Sphagnum
magellanicum,
Sphagnum
capillifolium, and
Polytrichum
strictum
Ombrotrophic
peat bog
Addition: 16 kg
N/ha/yr as NH4NO3
N dep = 8 kg
N/ha/yr as
quantified by
Turunen et al.
(2004)
N addition decreased
aboveground moss
biomass by 30%.
Mer Bleue Bog, Bog plant
Xing et al.
Ontario,	community: dwarf (2011)
Canada	shrub species
(measured and mosses
2007)	Sphagnum
magellanicum,
Sphagnum
capillifolium, and
Polytrichum
strictum
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Rich fen and 60 kg N/ha/yr as
ombrotrophic urea
'309	S dep = not
reported
N dep = not
reported
In Chamaedaphne
calyculata, fertilization
increased productivity by
87% and biomass by
105%.
In Calamagrostis
canadensis, biomass
increased 600% over
unamended fen plots.
The entire plant
community responded to
N addition with increases
in productivity (82%) and
biomass (160%).
Gogebic	Plant community Iversen et al.
County, Ml consisting of (2010)
Sphagnum spp.,
ericaceous
shrubs, dominant
vascular plants
Carex
oligosperma and
Chamaedaphne
calyculata
Rich fen Addition: two
aqueous
treatments:
groundwater
(0.18 mg N/L,
0.02 mg P/L, 90 mg
Ca2+/L,
pH = 8.0-8.6) and
rainwater
(0.58-0.98 mg N/L,
0.03 mg P/L, pH
6.5-7.0)
Three N addition
levels dissolved in
each aqueous
treatment: no
additional N (low), 1
mg N/L added
(medium), 3 mg N/L
added (high)
Ambient
deposition = not
specified, but
rainwater N is
220-440% higher
than groundwater N
The daily relative growth
rate (RGR) of all four
species decreases 18%
under rainwater rather
than under groundwater.
The switch to rainwater
accounted for 10% of
RGR variation in C.
cuspidata, 7% of RGR
variation in H. vernicosus,
and 32% of RGR
variation in P. squarrosa.
RGR of all four species
decreased by 12% under
medium N addition, and
decreased 24% under
high N addition. N
addition accounted for
4.6% of RGR variation in
B.	pseudotriquetrum,
6.7% of RGR variation in
C.	cuspidata, 18% of
RGR variation in H.
vernicosus, and 23% of
RGR variation in P.
squarrosa.
Mesocosms
constructed
from sand,
peat, and chalk,
with all four
species added
to each
mesocosm,
fully factorial
design: two
aqueous
treatments *
addition
levels x 3P
addition levels
3N
Two bryophyte Andersen et
species common al. (2016)
in rich fens in
Denmark:
Calliergonella
cuspidata and
Bryum
pseudotriquetrum,
and two
bryophyte species
that have been
declining in
abundance over
the past 100 yr:
Hamatocaulis
vernicosus and
Paludella
squarrosa
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Table 11-5 (Continued): Nitrogen loading effects upon production and biomass.
Type of Additions or Load Biological and Chemical
Ecosystem (kg N/ha/yr)	Effects
Study Site Study Species Reference
Ombrotrophic
bog
Ambient
deposition = 3.4-5.0
kg N/ha/yr across
the region, divided
for this study into
the following
ranges: 3.4-3.6,
3.8-4.1,	4.4-4.9,
4.9-5.0	kg N/ha/yr
Pitcher plant growth
increased with increasing
N deposition between 3.4
and 4.1 kg N/ha/yr. Plant
growth measured as plant
diameter, which
correlates with pitcher
opening width, pitcher
length, keel width, and
total pitcher width. All
traits positively correlated
with increasing N
deposition between 3.4
and 4.1 kg N/ha/yr.
Deposition above 4.1 kg
N/ha/yr decreased pitcher
plant growth by 8-21% as
measured by plant
diameter (compared to
plant diameter measured
between 3.8. and 4.1 kg
N/ha/yr). Pitcher plant
growth was lower at
higher deposition levels of
4.4-5.0.
Adirondack
Mountains, NY,
(11 sites)
Sarracenia
purpurea
Crumley et al.
(2016)
Calcareous,
rich fen
Addition = 50 kg
N/m2/yr as NO3" or
50 kg N/m2/yr as
NH4+.
Ambient
deposition = 7-10
kg N/m2/yr, based
on Aherne and
Farrell (2002)
NHx decreased
Scorpidium living biomass
67%.
Scragh Bog,
central Ireland
Scorpidium
revolvens and
Sphagnum
con tort urn
Paulissen et
al. (2016)
AG = aboveground; ANPP = aboveground net primary productivity; C02 = carbon dioxide; dep = deposition; ha = hectare;
kg = kilogram; N = nitrogen; NaN03 = sodium nitrate; NH4CI = ammonium chloride; NH4NO3 = ammonium nitrate;
(NH4)2S04 = ammonium sulfate; S = sulfur; yr = year.
11.5. Plant Stoichiometry and Physiology
1	N addition alters the stoichiometry of plant tissue. When inorganic N stores increase in
2	the soil matrix due to N deposition, one of the first plant responses may be to increase N
3	uptake and storage. As a result, the concentration of N in plant tissue (often reported as
4	[N] or %N) is one of the earliest indicators of changing bioavailability of N within an
5	ecosystem. Plants may store the additional N in proteins, use it for further growth if other
6	nutrients and abiotic conditions are not limiting, or curtail growth of belowground,
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nutrient-acquiring roots. Plant tissue [N] also has important implications for consumers,
as many insect and vertebrate consumers preferentially graze on high N tissues, and plant
[N] can also affect decomposition rates. In systems in which N availability increases
while other nutrient cycles are unaltered, such as in nutrient-poor bogs, nutrient
imbalances develop and plant [K] and [P] decline. This section will consider the effects
of N deposition upon plant tissue [N] as a sensitive indicator of plant response, as well as
the effects of N deposition upon other elements in plant tissue, which would indicate
more detrimental effects of increased N availability upon plant growth.
Plant stoichiometry theory considers the balance of multiple chemical elements in living
tissues. The stoichiometry of plant tissue is often connected to its physiological function.
Most new studies on physiology and stoichiometry are for bogs and freshwater marshes.
There are a few studies on salt water marshes and fens. The new literature is summarized
in the following sections.
11.5.1. Salt Marsh
There are several studies on the effects of N addition on salt marsh plant stoichiometry.
In three California salt marshes, N addition at seven different levels (0, 100, 200, 400,
800, 1,600, and 3,200 kg N/ha/yr) showed that the succulent forb Salicornia depressa
increased leaf %N in a saturating response, with no statistically significant increases in
response above 800 kg N/ha/yr (Vivanco et al.. 2015). There are several N addition
studies at addition levels greater than 500 kg N/ha/yr [(Baldwin. 2013; Nelson and
Zavaleta. 2012; Ryan and Bover. 2012); see Table 11-61.
11.5.2. Mangrove
The N addition levels in all new studies on mangroves are generally higher (greater than
500 kg N/ha/yr) than would be relevant to the effects of N deposition. As perennial
plants, mangroves respond to increasing N loads by altering tissue storage of N as well as
by increasing productivity (Appendix 11.4.3). In salt-marsh mangrove ecotone, Avicennia
germinans seedlings responded to addition of 1,400 kg N/ha with 62% higher leaf tissue
%N, as well as increased N in belowground biomass (Simpson et al.. 2013). In more
mature mangrove ecosystems receiving higher N loads (11,200 kg N/yr per tree), N
addition decreased the efficiency with which Rhizophora mangle resorbed N from
senescing leaves by 7%, which resulted in 39% higher %N in senesced leaves than in
unfertilized trees (Feller etal.. 2009).
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11.5.3. Freshwater Marsh
N addition is documented to alter plant stoichiometry and physiology in freshwater
marshes. This topic was not addressed in the 2008 ISA. Five new studies have been
published since 2008.
In freshwater marsh mesocosms planted with wetland obligate graminoid Bolboschoenus
maritimus [cosmopolitan bulrush, multiple synonyms including Schoenoplectus
maritimus, listed as Special Concern by Connecticut and Rhode Island, and as
endangered by Illinois, New Jersey, and New York; (USDA. 2015b)l N addition
increased N stored in plant tissues. Aboveground tissue %N increased 31-114% of %N
in control plants with increasing N loads (32, 65, 108 kg N/ha/yr) added in fall 2009, and
increased 33-67% with increasing N loads (59, 119, 198 kg N/ha/yr) added in the
summer and fall of 2010 (Duguma and Walton. 2014). N addition also increased tissue N
in belowground tissues. In fall 2009, addition of 32 kg N/ha/yr did not affect BG %N, but
the addition of 65 kg N/ha/yr increased BG %N by 26%, and the addition of 108 kg
N/ha/yr increased BG %N by 126 to 1.29% tissue N. In summer and fall 2010, the
addition of 119 and 198 kg N/ha/yr increased BG %N 47 and 37%, with a tissue N value
of 1.34 and 1.25% N, respectively (Duguma and Walton. 2014).
In tidal freshwater marshes of the Tchefuncte River, LA, experimentally added N loads
of 50, 200, 1,200 kg N/ha/yr did not alter the relative dominance of Sagittaria lancifolia,
but did alter N storage in plant tissues. Plant tissue N (mg N/g leaf) increased 0.02 mg for
every 10 kg N/ha/yr of N load. The change in plant N shifted the N:P ratio of plant tissue,
0.04 N: 1 P increase for every additional 10 kg N/ha/yr (Graham and Mendelssohn. 2010).
N addition also decreased the efficiency with which plants resorbed nutrients from
senescing leaves for use in future growth. N addition decreased N resorption efficiency
(%) by 0.1 % for every additional 10 kg N/ha/yr, and decreased P resorption efficiency
(%) by 0.09% for every additional 10 kg N/ha/yr (Graham and Mendelssohn. 2010).
In freshwater marsh mesocosms at the Smithsonian Environmental Research Center
(SERC), native and introduced haplotypes of graminoid Phragmites australis were grown
in a nitrogen addition experiment. The species P. australis includes haplotypes (genetic
lineages) native to North America, as well as invasive haplotypes capable of forming
monocultures that exclude native plant species. Nitrogen fertilization of 250 kg N/ha/yr
increased the construction cost (grams of glucose invested in synthesizing a gram of
biomass) of leaves for the native F haplotype oiP. australis by 5% (Caplan et al.. 2014).
Nitrogen addition did not significantly alter construction costs of the invasive M
haplotype [listed as a noxious weed, banned invasive, or plant pest by Alabama,
Connecticut, Massachusetts, South Carolina, Vermont, and Washington; (USDA. 2015b;
Caplan et al.. 2014)1. These results suggest that N enrichment favors the invasive
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haplotype over the native haplotype and may facilitate invasion; this is of concern
because P. australis tolerates brackish conditions and invades North American wetlands
all along the freshwater-freshwater tidal-salt marsh continuum.
In tidal freshwater marshes in the Altamaha River Estuary, GA, N addition of 500 kg
N/ha/yr in the 2nd year stimulated productivity (Appendix 11.4.2) to the extent of
diluting N and phosphorus (P) in plant tissues, decreasing leaf N (%) by 22%, and
decreasing leaf P (jxg/g) by 25% (Frost et al.. 2009). In the later years of the experiment,
N addition increased leaf N (%) above control plot leafN by 25-43% (Ketetal.. 2011).
and this increase altered leaf nutrient ratios, decreasing leaf C:N 18-28% and increasing
leaf N:P 25-61% over leaves from control plots. The increases in productivity combined
with alterations in leaf tissue nutrients altered total nutrient pools stored in aboveground
biomass, increasing total aboveground N by 236% and total aboveground P by 169%
(Ket et al.. 2011).
11.5.4. Riparian Wetland
In developing riparian forests at Bonanza Forest LTER, Alaska, N addition of 100 kg
N/ha/yr increased leafN 10% in tree species Alnus incana ssp. tenuifolia growing in late
successional riparian forest, but not in recently colonized or midsuccessional sand bars
(Ruess et al.. 2013). N addition also altered A incana ssp. tenuifolia's internal P cycling
efficiency, decreasing resorption of P from senescing leaves by 21% in midsuccessional
forest (Ruess et al.. 2013).
11.5.5. Bog and Fen
Carnivorous plants grow optimally under low-N conditions, and many are limited in their
distribution to bogs and fens. Recent research suggests that N deposition can disrupt their
N-acquisition strategy of digesting the invertebrates they trap. Purple pitcher plants
(.Sarracenia purpurea) were measured and sampled across a narrow N deposition
gradient (3.4-5.0 kg N/ha/yr) across bogs in the Adirondack Mountains (Crumley et al..
2016). Plant architecture, plant tissue N, and stable isotope N data suggested a growth
and carnivory optima for plants in bogs receiving 3.8-4.1 kg N/ha/yr. At higher
deposition rates of 4.4-4.9 kg N/ha/yr, there was no change in plant tissue N, but there
were negative effects upon growth (see Appendix 11.4.4) and a 40% decrease in
carnivory (Crumley et al.. 2016). The wetland obligate pitcher plants have previously
been shown to experience detrimental impacts of N deposition at physiological and
population levels lYGotelli and Ellison. 2002) in 2008 ISA], Bott et al. (2008) reciprocally
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transplanted Sarracenia purpurea ssp. purpurea [listed as exploitably vulnerable in New
York, threatened in New York, and endangered in Georgia and Illinois; (USDA. 2015b)!
between ombrotrophic Sapa bog and the rich Cedarburg fen, both in Wisconsin. Leaf %N
was positively and linearly correlated with surface water NOs concentration at the site
where the plants were transplanted (leaf %N = 0.9581 + 0.1667 x |iM NO3 ). confirming
Gotelli's designation of pitcher plants as sensitive indicators ofN deposition (Bott ct al..
2008).
A deposition-induced shift away from insectivory can lead to deficiency in other nutrients
for carnivorous plants. Across a natural deposition gradient in Sweden, the carnivorous
plant Drosera rotundifolia experienced significant declines of insectivory-derived N at
3.81 and 11.30 kg N/ha/yr (Millctt et al.. 2012). Across a broader deposition gradient of
European bogs (0.5-27.0 kg N/ha/yr), D. rotundifolia insectivory declined in a linear
relationship with increasing N deposition, and plants experienced increasing P limitation
as evidenced by increasing N:P tissue concentrations (Millett et al.. 2015). These studies
support previous research showing negative effects of N deposition upon D. rotundifolia
populations (see Appendix 11.7).
Research conducted in European fens and reviewed in the 2008 ISA has shown that N
addition favors the growth of grass and sedge species over peat-forming moss species
(U.S. EPA. 2008a). New research suggests that North American sedge and grass species
experience positive effects of N loading, and that their responses can alter N dynamics of
North American bogs and fens. In a bog in Gogebic County, MI, where sedge Carex
oligosperma [listed as Special Concern in Connecticut, threatened in Ohio and
Pennsylvania, and endangered in Illinois, Massachusetts, and North Carolina; (USDA.
2015b)] was dominant in the plant community, the obligate wetland species responded to
60 kg N/ha/yr addition by decreasing N use efficiency by 89% and N response efficiency
by 84% (Iversen et al.. 2010). The response of the sedge altered ecosystem responses in
the bog, increasing plant community N uptake by 125% over uptake in unamended plots
(Iversen et al.. 2010). These changes increased N stored in biomass pools 171% above
biomass N in control plots and increased the concentration of N in vascular plant tissues
by 19% (Iversen et al.. 2010). Similarly, the grass Calamagrostis canadensis [classified
as wetland facultative or obligate in different U.S. regions by (USDA. 2015b)l. which
was dominant in a closely adjacent rich fen, responded to N addition of 60 kg N/ha/yr by
increasing N uptake. Along with increased productivity (Appendix 11.4.5). this response
expanded the total pool of N stored in biomass by 300% (Iversen et al.. 2010). A similar
mesocosm study of bogs in western Europe across a deposition gradient of 2 to 47 kg
N/ha/yr found that at the highest deposition level, plant tissue N of grasses and sedges
was 120% higher than grasses and sedges at the lowest deposition level (Zaiac and
Blodau. 2016).
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In the same set of experiments in Gogebic County, MI, woody plants increased their
productivity in response to N loading, but their efficiency in using N declined, suggesting
a mechanism by which N loading lowers N retention of bogs and fens
(Appendix 11.3.1.5). The wetland obligate shrub Chamaedaphne calyculata (threatened
in Illinois and Maryland) responded to 60 kg N/ha/yr addition by increasing plant N
uptake by 75%, which increased productivity (Appendix 11.4) and increased the pool of
N stored in aboveground biomass by 100%. However, nitrogen addition decreased the
efficiency of photosynthesis and metabolism in C. calyculata, as N use efficiency was
81% lower and N response efficiency was 91% lower in fertilized plots (Iversen et al..
2010). In an intermediate fen, tree species Alnns incana ssp. rugosa, speckled alder
[hereafter referred to as A. rugosa, endangered in Illinois; (USDA. 2015b)l. responded to
an added 60 kg N/ha/yr with a 45% decline in N response efficiency [net primary
productivity relative to available soil N; (Iversen et al.. 2010)1. A European study of
mesocosms collected across a deposition gradient of bogs found that 12 kg N/ha/yr
decreased N retention of shrub species, and 47 kg N/ha/yr increased ecosystem total N
stored in shrub biomass (Zaiac and Blodau. 2016).
A long-term fertilization experiment at the Mer Bleue Bog in Ontario, Canada, has
documented the effect of more than a decade of elevated N inputs on the same species in
an ombrotrophic peat bog. Chamaedaphne calyculata responded to 32 g N/ha/yr by
increasing P resorption by 42%, and responded to 64 kg N/ha/yr by increasing P
resorption by 33% (Wang et al.. 2014b). N Plots that had received 64 kg N/ha/yr for
6 years decreased leaf calcium 22% (Wang et al.. 2014b). and leaf Ca was 34% lower in
C. calyculata from plots that received 9 years of 16 kg N/ha/yr than in leaves from
control plots (Bubier etal.. 2011). An observational study across a deposition gradient in
the Adirondack Mountains found effects on C. calyculata stoichiometry at much lower
deposition levels. There was a 27% increase in plant tissue N in C. calyculata when
deposition increased from a range of 3.4-3.6 kg N/ha/yr to 4.9-5.0 kg N/ha/yr (Crumley
etal.. 2016).
After 9 years of N addition at Mer Bleue, leaf C:N ratios were 15% lower in C.
calyculata leaves that received 16 kg N/ha/yr than in leaves from control plots, with
dependent increases in leaf alanine and y-aminobutyric acid (GABA; see Table 11-6 for
equations), amino acid pools that indicate N saturation in the plant (Bubier et al.. 2011).
Leaf aluminum concentrations were 44% lower in these leaves than in leaves from
control plots (Bubier etal.. 2011). In plots that received 64 kg N/ha/yr, there was no
change in leaf C:N, but there were increases in proteins indicative of increased plant N
uptake. C. calyculata leaf total chlorophyll was 84% higher than in control leaves, and
there were 115% increases in leaf alanine and 94% increases in leaf GABA, indicating N
saturation. Leaf manganese was 45% lower in these leaves than in control plot C.
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calyculata, indicating that shrub productivity may be limited by micronutrients in a
N-fertilized bog (Bubier et al.. 2011).
At Mer Bleue, the wetland obligate shrub Rhododendron groenlandicum (synonym:
Ledum groenlandicum, rare in Pennsylvania, threatened in Connecticut, and endangered
in Ohio) responded to 4 years of 64 kg N/ha/yr with a 16% increase in leaf N and a
dependent increase in leaf glutamic acid (Bubicr etal.. 2011). Leaf P declined by 54% of
control leaf P after 4 years of 64 kg N/ha/yr (Bubicr etal.. 2011). After 7 years of 64 kg
N/ha/yr, R. groenlandicum responded with a 23% decrease in leaf magnesium (Mg) and a
186% decrease in Mg resorption from senescing leaves (Wang et al.. 2014b). indicating
increasing limitation by magnesium, which could negatively affect photosynthesis and
enzyme activity within the plant. After 7 years of 32 and 64 kg N/ha/yr addition, R.
groenlandicum increased P and K resorption from senescing leaves; P resorption
increased 12-12.2 times the rate of resorption in leaves from unamended plots, and
increased K resorption by 92-123% (Wang etal.. 2014b).
In the same experiment, after 7 years, N addition shifted the seasonality of gross
photosynthesis, with addition of 64 kg N/ha/yr reducing C uptake 29-45% of control
rates in early summer (May-July), and increasing C uptake 25% above rates in control
plots in September (Larmola et al.. 2013). After 9 years of 16 kg N/ha/yr addition, R.
groenlandicum exhibited altered leaf physiology, doubling maximum carboxylation
capacity (Vcmax) on a per-leaf-mass basis (Bubier et al.. 2011). There was no significant
effect of 4 years of 64 kg N/ha/yr upon VCmax measured during the same growing season
(2008). The study authors posited that plants responded to this level of N availability with
a stress response, allocating N to storage in glutamic acid (nmol glutamic acid/g
leaf = 37.914 + 252.2 x leaf %N) instead of using it in photosynthetic enzymes (Bubier et
al.. 2011).
At Mer Bleue, the wetland facultative shrub Vaccinium myrtilloides (designated sensitive
in Washington, special concern in Connecticut, threatened in Ohio and Iowa, and
endangered in Indiana) responded to nitrogen addition with declines in other leaf
nutrients. Leaf calcium declined 23% in response to 9 years of 16 kg N/ha/yr and
declined 56% in response to 4 years of 64 kg N/ha/yr (Bubier et al.. 2011). indicating that
increasing N will increasingly limit leaf Ca and by extension decrease signaling and
structural integrity of plant tissues. Leaf P decreased by 55% after 9 years of 16 kg
N/ha/yr and by 68% after 4 years of 64 kg N/ha/yr. Leaf manganese (Mn) decreased 57%
under 64 kg N/ha/yr (Bubier et al.. 2011). which could affect photosynthesis or N
assimilation because Mn is a component of enzymes involved in these processes. V.
myrtilloides leaf potassium also declined: 47% in response to 9 years of 16 kg N/ha/yr
and 34% in response to 4 years of 64 kg N/ha/yr (Bubier et al.. 2011). Leaf potassium
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regulates stomatal movement, activates enzymes, and balances electrical charges
associated with ATP production, so K deficiency can decrease photosynthesis and slow
growth. Leaf aluminum decreased 45% of control leaf aluminum concentrations in plots
receiving 64 kg N/ha/yr for 4 years (Bubier et al.. 2011).
N addition can alter tissue nutrients in nonvascular plants as well. In the Mer Bleue bog,
N addition of 32 kg N/ha/yr decreased moss [Ca] by 34%; N addition of 64 kg N/ha/yr
increased moss [P] by 39%, and decreased moss [Ca] by 42% (Wang et al.. 2016a).
These results are consistent with N addition effects of Ca deficiencies in vascular plants
(see above) and ecosystem P limitation (see Appendix 11.3.2.1.3) within this bog
experiment. Surveys of two European bogs located in different mountain ranges found
that N deposition altered the stoichiometry of Sphagnum moss species. Sphagnum tissue
K concentrations were lower in samples collected from the high N deposition (20-25 kg
N/ha/yr) Jizera Mountains, 64% lower in S. rubrum and 62% lower in S. magellanicum,
than in samples from the lower deposition (12.5 kg N/ha/yr) Jeseniky mountain range. S.
magellanicum [Mg] was also 45% lower at the high N deposition site than at the lower N
deposition site (Jirousek et al.. 2015). Fritz etal. (2014) performed a peat mesocosm
study to test how a history of atmospheric N deposition alters the N uptake rates of
Sphagnum magellanicum, a widely distributed peat-forming species. The researchers
collected blocks of living S. magellanicum from a pristine bog in Argentina, which
received an annual load of 1-2 kg N/ha/yr, and also from a polluted bog in the
Netherlands, which received an annual load of 20-30 kg N/ha/yr. Under controlled
conditions, nitrogen solutions of 1, 10, or 100 |imol N/L were added to the block. Rates
of N uptake were significantly different among bog samples (Argentine or Dutch) for the
10 or 100 |imol solutions, with the pristine (Argentine) Sphagnum absorbing nitrogen at
1.4-2.6 times the rate of the uptake rate of the Dutch bog, which receives a higher annual
N load (Fritz et al.. 2014). A mesocosm study of four European moss species found that
applying rainwater (elevated N and P) rather than groundwater increases moss tissue [N]
and [P], while applying additional N in controlled treatments decreases moss tissue [P]
(Andersen et al.. 2016). A different European mesocosm study found that Sphagnum spp.
plant tissue N increased across a N deposition gradient of 2 to 47 kg N/ha/yr, such that
C:N ratio declined as a linear function of the natural log of increasing N deposition
(Zaiac and Blodau. 2016).
Recent research has addressed differential responses of bogs and fens to reduced or
oxidized forms of N addition. In an Irish fen dominated by Sphagnum contortum and
brown moss Scorpidium revolvens, 50 kg N/ha/yr was applied as NOs, in one treatment
and NH4+ in another treatment (Paulissen et al.. 2016). Reduced N forms elicited stronger
responses in tissue chemistry for each moss species. In S. contortum, oxidized N
increased tissue N 29%, tissue P 28%, and concentrations of alanine (a free amino acid
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hypothesized to be a mechanism of N storage) 31%; an equivalent addition of reduced N
decreased tissue N 13%, tissue K 49%, and dramatically increased (by half to 66 times
the control) the concentrations of five free amino acids storing excess N in plant tissue. In
the more sensitive species S. revolvens, reduced N increased tissue N 76% but decreased
tissue K 53%, and increased concentrations of arginine (a free amino acid) 7.3 times
concentrations in control plots; these same metrics were unaffected by the equivalent
oxidized N treatment (Paulisscn et al.. 2016). The stronger effects of reduced N on plant
chemistry in S. contortum than S. revolvens may represent species-level differences in N
tolerance, as S.contortum biomass did not change in response to N addition (see
Appendix 11.4.4).
The Whim Bog experiment in Scotland added N as either ammonium or nitrate to
simulate wet deposition, which with ambient N deposition of 8 kg N/ha/yr resulted in
treatment N loads of 16, 32, and 64 kg N/ha/yr. As in other wetlands, the bog had
divergent responses to oxidized and reduced N addition. Sphagnum capillifolium plant
tissue N and N:P increased in response to increasing oxidized N in a saturating function,
but responded to increasing reduced N in a linear function (Chiwa et al.. 2016). Reduced
N addition also resulted in increasing cation concentrations in pore water, suggesting
cation leaching by the moss species as a result of N uptake. These results indicate a
stronger effect of reduced N upon Sphagnum growth or P limitation, and these
physiological changes also affected N cycling and water quality within the bog (see
Appendix 11.3.1.5).
11.5.6. Summary Table
Table 11-6 Nitrogen loading effects upon plant stoichiometry and physiology.
Type of Additions or Load Biological and
Ecosystem (kg N/ha/yr)	Chemical Effects Study Site Study Species	Reference
Coastal salt
marsh
Addition: 100, 200,
400, 800, 1,600,
3,200 kg N/ha/yr as
urea
N dep = 3-5 kg
N/ha/yr as reported
in Tonnesen et al.
(2007)
Salicornia leaf %N
increases in a
saturating response to
added N (Nadd, as g
N/m2/yr), leaf
N = -0.73
Nadd + 2.99.
x e
Morro Bay
National
Estuary,
Carpinteria
Salt Marsh
Reserve,
Tijuana River
Reserve
Estuary, CA
Salicornia
depressa
(Salicornia
virginica) stands
Vivanco et al.
(2015)
June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Coastal salt
marsh
3,000 kg N/ha/yr as
NH3NO3
N dep = not
reported
S dep = not
reported
Plant [N] increased by
224 and 33%, and N
pools in AG vegetation
increased by 60 and
84% in successive
summers.
Elkhorn
Slough,
Monterey Bay,
CA
Sarcocornia
pacifica
dominant,
Distichlis spicata,
Frankenia salina,
and Jaumea
carnosa
Nelson and
Zavaleta (2012)
Estuarine salt
marsh
1,337 kg N/ha/yr as In S. pacifica, tissue N China Camp Sarcocornia
urea
N dep = not
reported
S dep = not
reported
increased 28-46%. D.
spicata tissue N
increased 19%. J.
carnosa tissue N
increased by 54%.
State Park, CA
pacifica
dominant,
Distichlis spicata,
and Jaumea
Ryan and Bover
(2012)
carnosa
Mangrove/
salt marsh
ecotone
1,400 kg N/ha/yr
N dep = not
reported
S dep = not
reported
Mangrove leaf
production increased
by 42%. Leaf tissue
%N increased 62%.
Root %N increased
9% and root C:N
decreased 25%.
Merritt Island
National
Wildlife
Refuge, FL
Avicennia
germinans
Simpson et al.
(2013)
Mangroves
No areal
reported;
N/yr per
N dep =
reported
S dep =
reported
rate
; 11,200 kg
tree
not
not
Fertilization increases
growth rate (shoot
elongation) by 290% in
the fringe zone and by
1,340% in the scrub
zone. In the fringe
zone, fertilization
decreases resorption
efficiency of N from
dying leaves by 7%,
resulting in senesced
leaves containing 39%
more N than control
senesced leaves.
Indian River
Lagoon, FL
Rhizophora
mangle
Feller etal. (2009)
Freshwater
marsh
Fall 2009: 32 (low),
65 (medium), 108
(high) kg N/ha/yr as
(NH^SCM; summer
2010: 59 (low), 119
(medium), or 198
(high) kg N/ha/yr as
(NH4)2S04
N dep = not
reported
S dep = not
reported
N addition increased
S. maritimus %N of
aboveground tissue
31-114% in 2009 and
33-67% in 2010. N
addition (medium,
high) increased S.
maritimus %N of
belowground tissue
26-126% in 2009 and
37-47% in 2010.
Mesocosms at
UC Riverside
research
station, CA
Schoenoplectus
maritimus, Culex
tarsalis, and
Anopheles
hermsi
Duquma and
Walton (2014)
June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Freshwater
tidal marsh
(growth
chamber)
250 g N/ha/yr as
NH4CI solution
N dep = not
reported
S dep = not
reported
Nitrogen increases
construction cost of
leaves by 5% in native
Phragmites. Nitrogen
under elevated [CO2]
increases construction
costs by 6% in
invasive Phragmites.
Mesocosms at
Smithsonian
Environmental
Research
Center, MD
Phragmites
australis (native F
haplotype and
invasive M
haplotype)
Caplan et al. (2014)
Freshwater
tidal marsh
500 kg N/ha/yr as
NH4CI or urea
N addition decreased Altamaha Zizaniopsis
leaf [N] by 99% and River, GA miliacea
leaf [P] by 25%.
Frost et al. (2009)
Freshwater
marsh
500 kg N/ha/yr as
NH4CI or urea
N dep = not
reported
S dep = not
reported
In Z. miliacea, leaf %N
increased 25-43% so
that leaf C:N
decreased 18-28%
and leaf N:P increased
25-61%. Total N and
total P in aboveground
pools increased 236
and 169%,
respectively.
Altamaha
Estuary, GA
Zizaniopsis
miliacea,
Pontederia
cordata, and
Sagittaria
lancifolia
Ket et al. (2011)
Freshwater
tidal marsh
50, 200, or 1,200 kg
N/ha/yr as
Nutralene
methylene urea
N dep = not
reported
S dep = not
reported
In S. lancifolia, N
addition (Nadd as kg
N/ha/yr) increased
plant tissue [N],
[N] = 0.00226 x Nadd +
23.271; and N:P ratio,
N:P = 0.00404 x
Nadd + 16.853.
In S. lancifolia,
resorption efficiencies
(RE) declined with
increasing N addition,
for nitrogen:
NRE = -0.01048 xNadd
+ 31.796, for
phosphorus:
PRE = -0.00915 xNadd
+ 62.778.
Tchefuncte
River,
Madisonville,
LA
Oligohaline plant
community
dominated by
Sagittaria
lancifolia,
Eleocharis fallax,
and Polygonum
punctatum
Graham and
Mendelssohn
(2010)
Freshwater
Addition: 670 kg
Foliar N increased in
Nanticoke Plant community Baldwin (2013)
tidal marsh
N/ha/yr
A. calamus (by 23%)
River, MD and

S dep = not
and Typha spp. (by
DE

reported
47%).


N dep = not



reported


June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Riparian
floodplain
successional
forest
100 kg N/ha/yr
N dep = not
reported
S dep = not
reported
Specific leaf mass
decreased by 12 and
11% in early and late
successional forest.
Leaf N increased 10%
in late successional
forest. Leaf P
resorption decreased
21%.
Bonanza
Forest LTER,
AK
Alnus incana ssp.
tenuifolia and
associated
Frankia strains
Ruess et al. (2013)
Ombrotrophic
bog
Ambient deposition
= 3.4-5.0 kg
N/ha/yr across the
region, divided for
this study into the
following ranges:
3.4-3.6, 3.8-4.1,
4.4-4.9, 4.9-5.0 kg
N/ha/yr
Pitcher plant foliar N
content increased 78%
between low N
deposition (3.4-3.6 kg
N) and medium N
deposition (3.8-4.1 kg
N). There was no
difference in foliar N
between medium and
higher deposition
rates.
Stable isotope (815N)
and mass balance
suggest that 68% of
pitcher plant N was
derived from carnivory
under low N
(3.4-3.6 kg N), 92%
under medium N
(3.8-4.1 kg N), and
55% under higher N
(4.4-4.9 kg N).
Increase from medium
to higher N deposition
decreased pitcher
plant carnivory by
40%.
Chamaedaphne
calyculata foliar N
increased 27% from
the lowest (3.4-3.6 kg
N) to the highest
deposition (4.9-5.0 kg
N) sites.
Adirondack
Mountains, NY
(11 sites)
Sarracenia
purpurea
Crumley et al.
(2016)
Bog
N deposition
gradient: 1.94,
and 11.30 kg
N/ha/yr
N from insectivory
3.81, decreases with
deposition: 55% from
insects under 1.94 kg
N/ha/yr, 20-30%
under 3.81 and
11.30 kg N/ha/yr.
Three bogs in
Sweden
Drosera
rotundifolia
Millett et al. (2012)
June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of Additions or Load Biological and
Ecosystem (kg N/ha/yr)	Chemical Effects Study Site Study Species	Reference
Ombrotrophic N deposition	N derived from	13 bogs across Drosera	Millett et al. (2015)
bog	gradient:	insectivory decreases Europe	rotundifolia
0.5-27.0 kg N/ha/yr with increasing N
deposition (Ndep as g
N/m2/yr) in a linear
relationship:
815N = -2.090 x Ndep
- 0.199, R2 = 0.43.
Plant tissue N:P
increases with
increasing N
deposition in a
logarithmic
relationship: tissue
N:P = 1.29 x In(Ndep)
+ 10.00, R2 = 0.62.
June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Ombrotrophic
peat bog
16 (low), 32
(medium), or 64
(high) kg N/ha/yr as
N (NH4NO3) or NPK
(NH4NO3 and
KH2PO4)
N dep = 8 kg
N/ha/yr
S dep = not
reported
In R. groenlandicum,
leaf%N increased by
25% (low N) and 16%
(high N), and leaf C:N
ratios decreased by
19% (low N) and 15%
(high N). Leaf P
declined by 54% (high
N). In low N addition,
R. groenlandicum
Vcmax increased by
76% on a per-area
basis or doubled on a
per-mass basis. In C.
calyculata, leafC:N
ratios decreased by
15% (low N) and 33%
(high N), and total
chlorophyll increased
by 84% (high N), just
as concentrations of
leaf alanine increased
115% (high N) and
GABA increased 94%
(high N). Leaf Ca
declined 34% (low N)
and 17% (high N), leaf
Mn declined 45% (high
N), and leaf Al
declined 44% (low N).
In V. myrtilloides, leaf
Ca declined 23% (low
N) and 56% (high N),
leaf P declined by 55%
(low N) and 68% (high
N), leaf Mn declined
57%, leaf K declined
47% (low N) and 34%
(high N), and leaf Al
declined 45% (high N).
Amino acid leaf
concentrations (y, as
nmol/g) increased with
leaf N (x, as %N): in R.
groenlandicum,
glutamic acid
increased
y = 252.2x+ 37.914; in
C. calyculata, alanine
increased
y = 221.07x- 52.931,
and GABA increased
y = 276.82x-98.682.
Mer Bleue
Bog, Ontario,
Canada
Three ericaceous
shrubs:
Vaccinium
myrtilloides,
Rhododendron
groenlandicum
(formerly Ledum
groenlandicum),
and
Chamaedaphne
calyculata
Bubieretal. (2011)
June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Ombrotrophic
peat bog
16 (low), 32
(medium), or 64
(high) kg N/ha/yr as
N (NH4NO3) or NPK
(NH4NO3 and
KH2PO4)
Under high N, gross
PSN declined 29-45%
in May-July, but
increased 25% in
September.
Mer Bleue
Bog, Ontario,
Canada
Shrub species
(Vaccinium
myrtilloides,
Rhododendron
groenlandicum,
Chamaedaphne
calyculata), and
mosses
(Sphagnum
magellanicum,
Sphagnum
capillifolium,
Polytrichum
strictum)
Larmola et al.
(2013)
Ombrotrophic
peat bog
16 (low), 32
(medium), or 64
(high) kg N/ha/yr as
N (NH4NO3) or NPK
(NH4NO3 and
KH2PO4)
N dep = 5-6 kg
N/ha/yr
S dep = not
reported
High N decreased leaf
Ca (mg/cm2) by 22%
in C. calyculata and
decreased leaf Mg
(mg/cm2) by 23% in R.
groenlandicum. In C.
calyculata, medium N
increased P resorption
by 42% as high N
increased it by 33%. In
R. groenlandicum, N
addition (medium N,
high N) increased P
resorption by
12-12.2x, and
increased K resorption
by 92-123%. High N
decreased Mg
resorption 186% in R.
groenlandicum.
Mer Bleue
Bog, Ontario,
Canada
Chamaedaphne
calyculata and
Rhododendron
groenlandicum
(formerly Ledum
groenlandicum)
Wang et al. (2014b)
Rich fen and
No addition
Leaf N (y, as %N) was
Cedarburg fen
Sarracenia
Bott et al. (2008)
ombrotrophic
N dep = not
reported
positively correlated
and Sapa bog,
purpurea ssp.

bog
with surface water
Wl
purpurea


NO3" concentration




S dep = not
(x):
y= 0.1667X+ 0.9581




reported



June 2018
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of Additions or Load
Ecosystem (kg N/ha/yr)
Biological and
Chemical Effects Study Site
Study Species	Reference
Plant community Iversen et al. (2010)
consisting of
Sphagnum spp.,
ericaceous
shrubs, dominant
vascular plants
Carex
oligosperma and
Chamaedaphne
calyculata
Ombrotrophic
bog
60 kg N/ha/yr as
urea
S dep = not
reported
N dep = not
reported
In Chamaedaphne Gogebic
calyculata, fertilization County, Ml
increases plant N
uptake by 75%, N
stored in biomass by
100%, productivity by
87%.
Fertilization decreases
C. calyculata N use
efficiency by 81% and
N response efficiency
by 91%.
In Carex oligosperma,
fertilization decreases
N use efficiency by
89% and N response
efficiency by 84%.
The entire plant
community responded
to N addition with
increases in N uptake
(125%), N stored in
biomass (171%),
productivity (82%), and
[N] in new tissue
(19%).
Fen
60 g N/ha/yr as
urea
N dep = not
reported
S dep = not
reported
In A. rugosa,	Gogebic
fertilization decreases County, Ml
N response efficiency
by 45%.
Dominant	Iversen et al. (2010)
vascular plants
Calamagrostis
canadensis and
Alnus rugosa
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Ombrotrophic
peatland
Ambient
deposition = 2 kg
N/ha/yr at DS;
12 kg N/ha/yr at
WM; 47 kg N/ha/yr
at FS
Mesocosms were
established with
NH4NO3 solutions
that mimicked
relative N
deposition levels
15N tracer was
added in
48 applications over
6 mo at a rate of
23 kg N/ha/yr
At 12 kg N/ha/yr, 15N
retention efficiency in
shrubs decreased
65% compared to DS.
At 47 kg N/ha/yr, plant
tissue N of graminoids
increased 120%
compared to DS. Total
N stored in shrubs
increased 360%
compared to DS.
Across five sites and
all mesocosms,
Sphagnum C:N was a
function of N
deposition (Ndep, as g
N/m2/yr):
C:N = -22.9 * ln(NdeP)
+ 68.6.
Data from LV and CF
mesocosms are not
considered because
mesocosm N addition
levels (CF: 300%
increase over DS N
solution, LV: 700%
increase over DS) did
not reflect ambient N
deposition differences
(CF: 750% increase
over DS deposition,
LV: 300% increase
over DS).
Across five sites and
all mesocosms,
Sphagnum C:N was a
function of N
deposition:
C:N = -22.9 * ln(NdeP)
+ 68.6.
Mesocosms
Sphagnum
constructed
capillifolium, S.
using peat bog
fallax, S.
cores from
magellanicum, S.
sites in
papillosum, S.
northern and
pulchrum, S.
western
rubellum,
1 Europe—
Andromeda
Degero
polifolia, Calluna
Stormyr,
vulgaris, Erica
Sweden (DS);
tetralix, Rubus
Fenn's,
chamaemorus,
Whixall, and
Vaccinium
Bettisfield
oxycoccos,
Mosses NNR,
Eriophorum
U.K. (WM);
vaginatum,
and
Eriophorum
Frolichshaier
angustifolum
Sattelmoor,

Germany (FS)

(Peat cores

also collected

at Little

Vildmose,

Denmark [LV]

and Cors

Fochno,

Wales, U.K.

[CF])

Zaiac and Blodau
(2016)
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Ombrotrophic
peatland
Addition = 16 kg
N/ha/yr (5N, or 5x
background
deposition), 32 kg
N/ha/yr (1 ON), or
64 kg N/ha/yr
(20N), all as
NH4NO3
Ambient (wet)
deposition = 8 kg
N/ha/yr and up to
0.26 kg P/ha/yr
N addition of 32 kg
N/ha/yr decreased
moss [Ca] by 34%. N
addition of 64 kg
N/ha/yr increased
moss [P] by 39%, and
decreased moss [Ca]
by 42%.
Mer Bleue bog,
Ottawa,
Canada
Evergreen
shrubs:
Chameadaphne
calyculata and
Rhododendron
groenlandicum
dominant; sparse
Kalmia
angustifolia
Deciduous
shrubs: Vaccinium
myrtilloides
Mosses:
Sphagnum
capillifolium, S.
magellanicum,
and Polystrichum
strictum
Wang et al. (2016a)
Ombrotrophic
bog
Wet + dry
deposition,
estimated by
Jirousek et al.
(2011)
Jireza: 20-25 kg
N/ha/yr (high N)
Jeseniky: 12.5 kg
N/ha/yr (low N)
S. rubellum [K] was
64% lower at high N
site. S. magellanicum
[K] was 62% lower and
[Mg] was 45% lower at
high N site.
Two sites:
Sphagnum fallax,
Jizerka in
Sphagnum
Jireza
magellanicum,
Mountains
and Sphagnum
(warm
rubellum/russowii
suboceanic

climate) and

Vozka in

Jeseniky

Mountains

(2°C colder)

Laboratory
Sphagnum
incubation of
magellanicum
moss from

pristine site in

Argentina and

N polluted site

in Netherlands

Jirousek et al.
(2015)
Bog
Laboratory
incubation. Moss
from Argentina
(54.75°S, 68.33°W)
and Netherlands
(52.82°N, 2.42°E)
Pristine site
(Argentina): 1-2 kg
N/ha/yr, N polluted
site (Netherlands):
20-30 kg N/ha/yr
N uptake of 10 and
100 pmol N/L solutions
was 40-160% more
rapid by Sphagnum
from the pristine
(Argentinian) bog than
the Dutch bog.
Fritz et al. (2014)
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Rich fen	Addition: two
aqueous
treatments:
groundwater
(0.18 mg N/L,
0.02 mg P/L, 90 mg
Ca2+/L,
pH = 8.0-8.6) and
rainwater
(0.58-0.98 mg N/L,
0.03 mg P/L, pH
6.5-7.0)
Three N addition
levels dissolved in
each aqueous
treatment: no
additional N (low),
1 mg N/L added
(medium), 3 mg N/L
added (high)
Ambient
deposition = not
specified, but
rainwater N is
220-440% higher
than groundwater N
Groundwater vs.
rainwater: tissue [N]
and tissue [P] were
higher in rainwater
than in groundwater
treatment.
N addition treatment:
tissue [P] was lower in
medium and high N
addition mesocosms.
Mesocosms
constructed
from sand,
peat, and
chalk, with all
four species
added to each
mesocosm,
fully factorial
design: two
aqueous
treatments * 3
N addition
levels x 3 P
addition levels
Two bryophyte
species common
in rich fens in
Denmark:
Calliergonella
cuspidata and
Bryum
pseudotriquetrum
Two bryophyte
species that have
been declining in
abundance over
the past
100 years:
Hamatocaulis
verrucosus and
Paludella
squarrosa
Andersen et al.
(2016)
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of
Ecosystem
Additions or Load
(kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Calcareous,
rich fen
Addition = 50 kg
N/m2/yr as NO3" or
50 kg N/m2/yr as
NH4+
Ambient
deposition = 7-10
kg N/m2/yr, based
Aherne and Farrell
(2002)
In Scorpidium, NHx
increased tissue N
76% and decreased
tissue K 53%. NHx
increased Scorpidium
free arginine
concentration 7.3x
control.
In Scorpidium, NOx
increased
phosphomonoesterase
(PMEase) 60%,
indicating increased P
limitation.
In Sphagnum, NHx
increased PMEase
70%, indicating
increased P limitation.
NHx decreased
Sphagnum tissue N
13% and tissue K
49%. NHx increased
the concentrations of
free amino acids in
Sphagnum: alanine by
55%, arginine 66x
control concentrations,
asparagine 39x
control, glutamine 65*
control, and serine
3.8x control.
In Sphagnum, NOx
increased tissue N
29% and tissue P
28%. NOx increased
Sphagnum free
alanine concentration
31%.
Scragh Bog,
Central Ireland
Scorpidium
revolvens and
Sphagnum
con tort um
Paulissen et al.
(2016)
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Table 11-6 (Continued): Nitrogen loading effects upon plant stoichiometry and
physiology.
Type of Additions or Load
Ecosystem (kg N/ha/yr)
Ombrotrophic Addition = 8, 24,
peatland	and 56 kg N/ha/yr
as either NH4+ or
NO3" since 2002
Ambient deposition
total N = 8 kg
N/ha/yr, 3 kg
N/ha/yr as wet NOx,
3 kg N/ha/yr as wet
NHx, and 2 kg
N/ha/yr as dry NHx
Biological and
Chemical Effects Study Site
Tissue N in Sphagnum Whim bog,
capitulum, and tissue Edinburgh,
N and N:P in	Scotland
Sphagnum stems,
increased in a
saturating
(i.e., logarithmic)
relationship with
increasing NOx
deposition.
Tissue N and N:P in
Sphagnum capitulum,
and tissue N in
Sphagnum stems,
increased in a linear
relationship with
increasing
NHxdeposition.
Study Species	Reference
Sphagnum moss Chiwa et al. (2016)
(Sphagnum
capillifolium, a
hummock-
forming species)
Heathland
community:
Calluna vulgaris,
Eriophorum
vaginatum,
Hypnum
jutlandicum,
Pleurozium
schreberi, and
Cladonia
portentosa
AG = aboveground; Al = aluminum; ANPP = aboveground net primary productivity; C = carbon; Ca = calcium; cm = centimeter;
C02 = carbon dioxide; dep = deposition; GABA = gamma-Aminobutyric acid; ha = hectare; K = potassium; kg = kilogram;
KH2PO4 = potassium phosphate; L = liter; LTER = Long Term Ecological Research; |jmol = micromole; mg = milligram;
Mn = manganese; N = nitrogen: NH4CI = ammonium chloride; NH4NO3 = ammonium nitrate; (NH4)2S04 = ammonium sulfate;
N03" = nitrate; NPK= nitrogen, phosphorus, potassium; NRE = nitrogen resorption efficiency; P = phosphorus; PRE = phosphorus
resorption efficiency; S = sulfur; yr = year.
11.6. Plant Architecture
Plant architecture (plant height, branching) in wetlands is an important endpoint because
the architecture of the dominant plants determines the availability of nesting habitats or
refugia from flooding and predators. Plant architecture is defined as the
three-dimensional organization of the plant body. For the parts of the plant that are
aboveground, this includes the branching pattern, as well as the size, shape, and position
of leaves and flower organs (Table 11-7). The 2008 ISA did not have sufficient data on N
addition effects on wetland plant architecture to consider plant architecture as a specific
endpoint.
11.6.1. Salt Marsh
There are several new studies on salt marsh plant architecture. Darby and Turner (2008)
reported that for Spartina alterniflora, stem height increased linearly with N load starting
at 230 kg N/ha/yr, although regression equations were not included. Stem density was
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significantly higher than in unamended plots only for the highest N load, 3,720 kg
N/ha/yr, which increased stem density by 43%. There are two other addition studies with
N addition rates above 500 kg N/ha/yr that evaluated effects in U.S. salt marshes (Ryan
and Bover. 2012; Davevetal.. 2011).
The Plum Island tidal eutrophication experiment added nitrate and phosphate to salt
marsh tidal creeks to achieve nitrate concentrations 15 times the concentrations of
unamended creeks. In the last 2 years of this experiment, there was N but no P
enrichment, and 620 kg N/ha/yr increased Spartina alterniflora plant height 4% in low
marsh (Johnson et al.. 2016a). In the high marsh, 140 kg N/ha/yr decreased Distichlis
spicata height 4% (Johnson et al.. 2016a). Across the course of the N and P addition,
alterations in plant carbon allocation decreased vascular plant root: shoot biomass ratio by
31% (Dccgan et al.. 2012). With the decrease in belowground support and the decreases
in shoot lignin (which provides structural support in plant shoots, see Stoichiometry
section), a 5% increase in shoot height made plants top-heavy. As a result, lodging (plant
stems collapsing to marsh substrate where they are inundated by tides), which was not
observed in control marshes, affected 41% of the marsh area in enriched marsh (Deegan
et al.. 2012).
11.6.2. Mangrove
In Florida mangroves, Whigham et al. (2009) added 100 kg N/ha/yr to plots of dwarf
Avicennia germinans, black mangrove, which increased the number of new branches
produced 150% above control plots.
11.6.3. Freshwater Tidal Marsh
There are several new studies on freshwater tidal wetlands. In freshwater tidal marshes of
Altamaha River Estuary, GA, N addition (500 kg N/ha/yr) increased the number of leaves
per m2 52% and increasing plant height by 25-40% over the course of the 5-year
experiment (Ket et al.. 2011). Duguma and Walton (2014) constructed freshwater
wetland mesocosms at UC Riverside, planting Bolboschoenus maritimus (formerly
Schoenoplectus maritimus) and alkali bulrush, and monitoring plant responses as well as
the use of mesocosms by mosquito species (Appendix 11.8.2). N addition of 32 kg
N/ha/yr increased plant height 23% with no effect upon plant height when 65 or 108 kg
N/ha/yr were added in 2009. However, in 2010 when the experiment was initiated earlier
in the year (5 August start date in 2009, 22 March start date in 2010), all N addition
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levels (59, 119, 198 kg N/ha/yr) increased plant height 28-44% above control plants
(Duguma and Walton. 2014).
11.6.4. Riparian Wetland
In developing riparian forests at Bonanza Forest LTER, AK, N addition of
100 kg N/ha/yr altered leaf structure of the tree species Alnus incana ssp. tenuifolia
(Ruess et al.. 2013). In the mostly recently colonized sand bars, N addition decreased
specific leaf mass (SLM) of A. incana ssp. tenuifolia 12%, and in midsuccessional
riparian forest SLM decreased 11% with N addition.
11.6.5. Summary Table
Table 11-7 Nitrogen loading effects upon architecture.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Coastal salt
marsh
4,200 kg N/ha/yr
as NH4NO3
S dep = not
reported
N dep = not
reported
N addition increased
rhizome diameter by
1% in shallow and 5%
in deep (10-20 cm)
sediments.
Goat Island,
SC
Spartina
alterniflora
Davev et al. (2011)
Coastal salt
marsh
230, 465, 930,
1,860, or 3,720 kg
N/ha/yr as
(NH4)2S04
S dep = not
reported
N dep = not
reported
Stem density and stem
height increased
linearly with N load
(regressions not given).
Stem densities
increased 43% in
response to 3,720 kg N.
LUMCON,
Cocodrie, LA
Spartina
alterniflora
Darby and Turner
(2008)
Estuarine
marsh
N/ha/yr as urea.
S dep = not
reported
N dep = not
reported
In S. pacifica, height
China Camp Sarcocornia
increased 207%,
State Park, CA pacifica
number of branches
dominant (C3
increased 135-283%.
succulent
J. carnosa height
shrub),
increased 504%.
Distichlis

spicata (C4

grass), and

Jaumea

carnosa (C3

semisucculent

forb)
Ryan and Bover
(2012)
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Table 11-7 (Continued): Nitrogen loading effects upon architecture.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species
Reference
Salt marsh
Nitrate and
phosphate
dissolved in tidal
inflows to raise
aqueous NO3"
concentrations to
70-100 |jM and
raise PO43" to
5-7 |jM.
Background load
in tides: 5 |jM
NO3-, 1 |JM PO43-
N dep = not
reported
S dep = not
reported
Root:shoot ratio
decreases 31%.
Shoot height increases
5%.
Lodging affects 41% of
area in enriched
marshes, is not
observed in control
marsh.
Plum Island
Estuary, MA
Primary tidal
creeks with
Spartina
alterniflora in
low marsh
along creeks,
and Spartina
patens in high
marsh
platforms
Deeaan et al.
(2012)
Salt marsh
Addition of nitrate
dissolved in
incoming tide
2011-2012,
addition of N at
70-100 |jM
NaNC>3" in tide for
added load of
620-1,200 kg
N/ha/yr in low
marsh and
70-140 kg N/ha/yr
in high marsh
Ambient
deposition = not
specified
Low marsh in enriched
creeks received an
additional N load of
620 kg N/ha/yr in 2012.
That year, N addition
increased shoot height
4% above S.alterniflora
in low marsh at
reference creeks.
High marsh in enriched
creeks received an
additional N load of
140 kg N/ha/yr in 2011.
In D. spicata, N addition
decreased shoot height
4%.
Plum Island
Sound
Estuary, MA
Spartina
alterniflora,
Distichlis
spicata, and
Spartina patens
(high marsh);
Spartina
alterniflora (low
marsh)
Johnson et al.
(2016a)
Mangrove
100 kg N/ha/yr
N addition increased
Indian River
Avicennia
Whiqham et al.


the number of new
Lagoon, FL
germinans and
(2009)


branches 150% in
(impoundment
associated



Avicennia.
SLC-24)
sediments

Freshwater
marsh
Fall 2009: 32
(low), 65
(medium), 108
(high) kg N/ha/yr
as (NH4)2S04;
Summer 2010: 59
(low), 119
(medium), or 198
(high) kg N/ha/yr
as (NH4)2S04
N dep = not
reported
S dep = not
reported
N addition increased S.
maritimus height 23% in
2009 (low N), and
28-44% in 2010.
Mesocosms at
UC Riverside
research
station, CA
Schoenoplectus
maritimus,
Culex tarsalis,
and Anopheles
hermsi
Duauma and
Walton (2014)
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Table 11-7 (Continued): Nitrogen loading effects upon architecture.
Type of Additions or	Biological and
Ecosystem Load (kg N/ha/yr) Chemical Effects Study Site Study Species Reference
Tidal
500 kg N/ha/yr as
In Z. miliacea, leaf
Altamaha
Zizaniopsis
Ket et al. (2011)
freshwater
NH4CI or urea
number increased 52%
Estuary, GA
miliacea,

marsh

and plant height
increased 25-40%.

Pontederia
cordata, and
Sagittaria
lancifolia

Riparian
100 kg N/ha/yr
Specific leaf mass
Bonanza
Alnus incana
Ruess et al. (2013)
floodplain
N dep = not
reported
S dep = not
reported
decreased by 12 and
Forest LTER,
ssp. tenuifolia

successional
11% in early and late
AK
and associated

forest
successional forest.

Frankia strains

Ombrotrophic
16 (low), 32
In medium and high
Mer Bleue
Bog plant
Juutinen et al.
peat bog
(medium), or 64
NPK treatments, shrub
Bog, Ontario,
community:
(2010)

(high) kg N/ha/yr
canopies grew 72 and
Canada
dwarf shrub


as N (NH4NO3) or
82% taller than

species and


NPK (NH4NO3 and
controls.

mosses


KH2PO4)
There were no

Sphagnum


S dep = not
significant effects of

magellanicum,


reported
N dep = 8 kg
medium or high N on
shrub canopies.

Sphagnum
capillifolium,
and


N/ha/yr


Polytrichum
strictum

AG = aboveground; ha = hectare; kg = kilogram; KH2P04 = potassium phosphate; LTER = Long Term Ecological Research;
LUMCON = Louisiana Universities Marine Consortium; N = nitrogen; NH4CI = ammonium chloride; NH4N03 = ammonium nitrate;
(NH4)2S04 = ammonium sulfate; P = phosphorus; S = sulfur; yr = year.
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11.7. Demography
Plant demography is the change in plant population size and structure through time.
Changes to plant demography will determine the long-term stability of wetlands.
Information on the effect of N loading on wetland plant demography, including
establishment, survival, and reproduction is sparse. Evidence from salt marsh ecosystems
shows mixed responses to N loading: two pre-2008 studies found negative effects of N
addition upon reproduction in invasive Spartina foliosa and Spartina hybrids (Tyler et al..
2007). and a positive effect upon reproduction of native Salicornia bigelovii (Bover and
Zedler. 1999). Work reviewed in the 2008 ISA on Sarraceniapurpurea, or northern
pitcher plant, found that increases in N deposition increased population extinction risk
(Gotclli and Ellison. 2006. 2002). which serves as the scientific basis for the wetland
critical load values (Appendix 11.9). Additionally, an older study of Drosera rotundifolia
in a Swedish bog found that N addition levels of 20 and 40 kg N/ha/yr significantly
decreased survivorship in comparison to control plots or N addition of 5 or 10 kg N/ha/yr
(Rcdbo-Torstcnsson. 1994). The sections below describe two new studies that find
negative impacts of high N upon the demography of wetland plants.
11.7.1.	Mangrove
Lovelock et al. (2009) considered a global data set of mangrove responses to N
fertilization experiments conducted for 3-12 years. N addition (900-2,700 kg N/tree/yr)
decreased survival probability by 10%, with trees growing on the land side in hypersaline
conditions particularly vulnerable when rainfall was low.
11.7.2.	Riparian Wetland
The Great Salt Lake is an inland salt lake ecosystem that provides an important food
source for migrating bird species as well as an economically important fishing industry
by providing habitat for Artemia franciscana, brine shrimp. Carling et al. (2013) studied
impounded and floodplain wetlands along the rivers that feed into the GSL, and found
that increasing nitrate and nitrite concentrations in water flowing through the wetlands
decreased reproduction of the submerged aquatic plants Ruppia cirrhosa and Stuckenia
spp. (equations in Table 11-8).
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11.7.3. Summary Table
Table 11-8 Nitrogen loading effects upon demography.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and Chemical
Effects
Study
Site
Study
Species
Reference
Mangrove
No areal rate
given; 900 kg/yrto
2,700 kg N/yr per
tree as urea
Mangrove survival probability
decreased 10% with N addition.
12 global Mangrove
sites, species
including
Indian
River
Lagoon,
FL
Lovelock et al.
(2009)
Riparian
wetlands
and
floodplain
wetlands
Not reported
Reproduction (Repro, g/m2
drupelets) declines with
increasing surface water nitrate
(NO3 in mg/L):
Repro = -4.15 * ln(NC>3) + 4.03,
R2 = 0.51; as well as with
increasing surface water NO2"
(NO2" in mg/L):
Repro = -13.43 x ln(N02")
-25.27, R2 = 0.54).
Great
Salt
Lake, UT
Submerged
aquatic
vegetation:
Ruppia
cirrhosa and
Stuckenia
spp.
Carlinq et al. (2013)
g = gram; ha = hectare; kg = kilogram; L = liter; m = meter; mg = milligram; N = nitrogen; N02 = nitrite; N03 = nitrate; yr = year.
11.8. Biodiversity/Community
High biodiversity is a characteristic of wetlands. Fluctuating water levels in open systems
such as riparian strips and tidal marshes create zones with distinct abiotic characteristics
and multiple niches for microbes, plants, and animals with different biotic and abiotic
requirements. More closed systems, such as bogs and fens, are nutrient-poor,
high-organic-acid ecosystems that contain rare species including carnivorous plants,
ericaceous plants, and bryophytes characterized by low growth rates. Most wetlands
contain species with a range of N tolerance and N-acquisition strategies, and one result of
eutrophication is to shift the composition and relative abundance of species so that
nitrogen-tolerant species become more common as species that are adapted to low
nitrogen availability become less common or disappear from the system. Thus, plant
cover (percentage of the total wetland area occupied by a particular species) is an
important metric of biodiversity. Similarly, communities of aquatic producers contain
both nitrogen-tolerant and low-nitrogen-adapted species, and changes in phytoplankton
species composition represents a change in biodiversity important for dependent food
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webs (see Appendix 9.3.2. Appendix 10.3.2. and Appendix 10.3.3). Changes in the
abundance of herbivorous, omnivorous, or predatory invertebrates are indicative of
alterations to trophic interactions and effects, which could cascade up the food web
towards economically or culturally important fish, bird, and mammal species.
The 2008 ISA found that the evidence was sufficient to infer a causal relationship
between N deposition and the alteration of species richness, species composition, and
biodiversity in wetland ecosystems. The 2008 ISA noted there are 4,200 native plant
species in U.S. wetlands, 121 of which are federally endangered. Wetlands provide
habitat to a disproportionally high number of rare plants given their relative area; for
example, fens occupy 0.01% of the land area of northeast Iowa, yet contain 17% of the
endangered, threatened, or listed species of concern in the area. Wetland species have
evolved under N limited conditions, including endangered species in the Isoetes (three
endangered species) and Sphagnum (15 endangered species) genera, as well as the
endangered insectivorous plants Sarracenia oreophila and Drosera rotundifolia.
The 2008 ISA provided evidence from American bogs and fens that showed that
Sarracenia purpurea (purple pitcher plant) responded to N deposition with changes in
morphology and projected population extinction under N deposition levels of 4.5-6.8 kg
N/ha/yr. Coastal wetlands responded to N enrichment with increased primary production,
shifting microbial and plant communities, and altering pore water chemistry, although
many of the studies in coastal wetlands used N enrichment levels more similar to that of
wastewater than of atmospheric deposition.
11.8.1.
11.8.1.1.
Plants
Salt Marsh
Nitrogen was an important factor driving plant diversity in a study on the Patuxent River,
where marsh soil NOs -N and pore water salinity best explained the variation in plant
species richness out of all factors considered, with NO3 -N accounting for 26% of the
variation in the richness of all species. Presence-absence data for all species were used to
determine the geographic range for each species, and the 50% of species per river with
the smallest range was designated as a "restricted-range" subset. On the Patuxent River,
the best model of variation in species richness of restricted range species contained only
soil NO3 -N as an explanatory variable, with a positive relationship between NO3 -N and
species richness [see Table 11-9; (Sharpe and Baldwin. 2013)1.
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Several new field addition studies further confirm N addition alters plant community
composition in salt marshes. In Kirkpatrick Marsh, MD, a limited plant community and
resampling of fixed plots allowed researchers to measure the biomass of plots as a proxy
for plant cover. C4 plants Spartina patens and Distichlis spicata plot cover increased over
the course of the 4-year fertilization experiment so that C4 biomass was 129% higher in
the 4th year than in the 1st year of receiving 250 kg N/ha/yr, while C4 biomass in the
control plots remained unchanged over 4 years. The C3 plant species Schoenoplectus
americanus declined in biomass in the control plots by 8-22% of that produced in the
1st year, but decreased to a greater extent in N addition plots, declining after 4 years by
52% compared to the 1st year plot biomass (Langlcv and Megonigal. 2010). In the
long-term nutrient addition experiment at Great Sippewissett Marsh, MA, 30 years of
addition of sewage sludge altered community composition. Nitrogen addition at 170, 520,
and 1,560 kg N/ha/yr altered the relative composition of Spartina alterniflora and
Distichlis spicata, with percentage cover of S. alterniflora decreasing 3% with every
additional 100 kg N/ha/yr, and percentage cover of D. spicata increasing 3% with every
additional 100 kg N/ha/yr of total N load (Fox et al.. 2012).
In Elkhorn Slough, CA, marsh community was assessed at six sites. At one site, where
impoundment restricted tidal exchange, N addition of 150 kg N/ha/yr altered community
composition, decreasing cover of native marsh plants by 52% and increasing cover of
non-native upland plants by 80% (Goldman Martone and Wasson. 2008). There were two
studies with N addition above 500 kg N/ha/yr (Baldwin. 2013; Ryan and Bover. 2012).
but this N addition level was not useful in inferring effects of N deposition upon
ecosystems.
11.8.1.2. Freshwater Tidal Marsh
In a freshwater tidal marsh on the Tchefuncte River, LA, N addition shifted the relative
dominance of the perennial wetland-obligate monocots dominant at the site (Graham and
Mendelssohn. 2010). Increasing N loads increased the relative dominance of Persicaria
punctata (formerly Polygonum punctatum) with a 0.13% increase (percentage of total
biomass consisting of P. punctata) for every 10 kg N/ha/yr added. Increasing N loads
decreased the relative dominance of Eleocharis fallax, with a 0.08% decrease (percentage
of total biomass consisting of E. fallax) for every 10 kg N/ha/yr added (Graham and
Mendelssohn. 2010) N loading did not affect the relative dominance of Sagittaria
lancifolia, but did alter its storage ofN in plant tissues (Appendix 11.5).
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11.8.1.3. Freshwater Wetland
There are two new studies of N loading in freshwater systems; they do not quantify N
deposition or N loading in terms of an areal rate, but rather find correlations between
surface water or soil N and changes in plant community diversity across a gradient of N
loads. Nitrogen loading is often correlated with invasions of non-native plants and
associated reductions in native species abundance and richness (Green and Galatow itsch.
2002). An observational study of 48 wetlands along the coasts of the Great Lakes found 7
floristically distinct plant communities: 1 Sphagmim-dominated wetland, 3 marsh types
dominated by native plants, and 3 marsh types dominated by invasive species. A
classification and regression tree that successfully classified 79% of the sites to the
correct plant community used surface water total N, conductivity, and pH as explanatory
variables: one of the invasive plant communities (dominated by the invasive Typha spp.)
was associated with high total N (Johnston and Brown. 2013).
In a similar observational study of 28 restored wetlands in Illinois, soil nitrogen
significantly correlated with plant invasion and a decline in native plant species richness
(Matthews et al.. 2009). There were linear, positive correlations of soil nitrate with
relative percentage cover of non-native plants, and with relative percentage cover of
invasive Phalaris arundinacea (see Table 11-9 for full equations). For each 0.1 ppm
increase in soil nitrate, non-native plant cover increased 4%, and P. arundinacea cover
increased 5%. The surveyed wetlands were diverse communities: there were 483 vascular
plant species across all 20 wetlands, and each wetland contained 40-100 native species.
For every 0.1 ppm increase in soil ammonium across the wetlands, seven native species
disappeared; and for every 0.1 ppm increase in soil nitrate, three native species
disappeared (Matthews et al.. 2009).
In addition to these studies, the National Wetlands Condition Assessment (NWCA),
conducted by U.S. EPA in 2011, collected soil and water chemistry data as well as plant
biodiversity data. In California wetlands sampled for this project, higher surface water
nitrate + nitrite concentrations correlated with lower wetland condition scores as
measured by the California Rapid Assessment Method (CRAM) (U.S. EPA. 2016i).
11.8.1.4. Intermittent Wetland
Vernal pools are intermittent wetlands, which typically flood in the late winter and
spring, and serve as important habitat for amphibians and endemic invertebrates.
California vernal pools host a number of endangered and threatened animal species
during their wet phase, as well as a high diversity of plant species during their dry phase.
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There are no studies published that directly test the effect of N deposition or N addition to
vernal pools, but there is a mesocosm study, which may allow some inference ofN
effects on vernal pools. This study added combined N and P solutions to mesocosms and
found increased algal biomass, which decreased vascular plant cover and richness in
vernal pool mesocosms (Kncitcl and Lessin. 2010). N and P additions were not quantified
in areal rates.
11.8.1.5. Bog and Fen
In Mer Bleue Bog, Ontario, N addition decreased the dominance of Sphagnum mosses in
the northern ombrotrophic peat bog. After 4 years of N fertilization at the rate of 64 kg
N/ha/yr, cover of Sphagnum species (of Sphagnum magellanicum and Sphagnum
capillifolium) in fertilized plots was 43% of Sphagnum cover in control plots (Juutinenet
al.. 2010). At Year 12, cover of Sphagnum species in 16 kg N/ha/yr fertilized plots was
64% of cover in control plots (Larmola et al.. 2013). Plots measured in the same growing
season (2011) that had received 7 years of elevated N at higher N addition rates also
experienced declines in Sphagnum cover: 26% decline in cover at 32 kg N/ha/yr and 54%
decline in cover at 64 kg N/ha/yr (Larmola et al.. 2013). The following year, moss
species (Sphagnum spp. and Polystrichum strictum) cover was 63% lower in 64 kg
N/ha/yr plots than in control plots, while the percentage of cover of the vascular plant
Chamaedaphne calyculata was 202% higher in 32 kg N/ha/yr plots than in control plots
(Wang et al.. 2016a). These declines in peat-forming species corresponded to changes in
ecosystem C fluxes (Appendix 11.3.2).
N deposition or addition effects upon plant community composition have been studied in
European systems. In an Irish fen, which was also the site of aN addition experiment (see
Appendix 11.4.4 and Appendix 11.5.5). researchers observed plant community
composition shifts in control plots over 4 years (Paulissen et al.. 2016). Under ambient
deposition of 7-10 kg N/ha/yr, cover of the N-sensitive brown moss Scorpidium
revolvens decreased 25%, while cover of peat moss Sphagnum contortum increased 20%
and cover of vascular plants increased qualitatively (Paulissen et al.. 2016). N addition
increases vacular plant cover in multiple systems. In an Italian fen, N addition altered
plant cover of 4 of the 17 species documented in the marsh (Gerdol and Brancaleoni.
2015). Vascular species cover did not shift in response to 10 kg N/ha/yr; but in response
to 30 kg N/ha/yr addition, cover of shrub Calluna vulgaris increased 96% and cover of
grass Molinia caerulea increased 5.6 times that of control plots. Like the vascular plants,
moss species Sphagnum fuscum (native in the U.S., listed as endangered by North
Carolina) did not respond to 10 kg N addition, but its cover decreased 45% in response to
30 kg N addition. The moss species Polytrichum strictum (native in the U.S., listed as
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1	endangered in Kentucky) increased in response to both 10 or 30 kg N addition, with
2	cover 190-240% higher after 8 years of N addition. Researchers continued to monitor the
3	plots for 3 years after ceasing N addition treatments to assess wetland recovery: vascular
4	plants and P. strictum remained dominant, while S. fuscum cover continued to decline,
5	and a different moss species, Sphagnum magellanicum, increased its coverage in plots
6	recovering from elevated N addition (Gcrdol and Brancaleoni. 2015).
11.8.1.6. Summary Table
7
Table 11-9 Nitrogen loading effects upon plant biodiversity and communities.
Type of
Additions or
Biological and



Ecosystem
Load (kg N/ha/yr)
Chemical Effects
Study Site
Study Species
Reference
Coastal high
Addition: 150 kg
At a site where tidal
Six sites at
Marsh
Goldman
marsh
N/ha/yr as urea
inflow was restricted,
Elkhorn
community
Martone and

S dep = not
reported
N dep = not
reported
fertilization increased
Slough,
dominated by
Wasson

cover of non-native
Watsonville,
Sarcocornia
(2008)

upland plants by 80%
CA
pacifica, also


and decreased cover of

contains Jaumea


native marsh plants by
52%.

carnosa,
Frankenia salina,
Spergularia
salina, Distichlis
spicata, and
Atriplex
californicaAriang
uiaris

Estuarine salt
Addition: 1,337 kg
S. pacifica cover
China Camp
Sarcocornia
Rvan and
marsh
N/ha/yr as urea
increased at 6.2x the
State Park,
pacifica
Bover (2012)

S dep = not
rate of increase in
CA
dominant (C3


reported
N dep = not
reported
control. J. carnosa cover

succulent shrub),


declined 5.4x as fast as
in control.

Distichlis spicata
(C4 grass), and
Jaumea carnosa
(C3
semisucculent
forb)

Estuarine salt
Addition: 250 kg
S. americanus
Kirkpatrick
Schoenoplectus
Lanalev and
marsh
N/ha/yr
aboveground biomass
Marsh, MD
americanus (C3),
Meqoniqal

S dep = not
reported
N dep = not
reported
decreased by 19 and
(measured in
Spartina patens
(2012. 2010)

45% in the 3rd and 4th
3rd and 4th
(C4), and


yr, respectively.
yr,
Distichlis spicata



2008-2009)
(04)

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Table 11-9 (Continued): Nitrogen loading effects upon plant biodiversity and
communities.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
Coastal salt
marsh
Addition: 170 kg
N/ha/yr, 520 kg
N/ha/yr, 1,560 kg
N/ha/yr, all added
as sewage sludge
(10% N, 6% P2O5,
4% K2O)
S dep = not
reported
N dep = not
reported
Background total
N
load = 2 kg/ha/yr
(as reported in
Bowen and Valiela
(2001).
S. alterniflora cover (y, as
% cover) declines in plots
with increasing N load
(Nadd, as kg N/ha/yr),
y = -0.0292 x Nadd +
64.6.
D. spicata cover (z,as %
cover) increases in plots
with increasing N load,
z= 0.0327 x Nadd + 3.04.
Great
Sippewissett
Salt Marsh,
MA
Salt marsh
community,
dominated by
Spartina
alterniflora,
Spartina patens,
Distichlis spicata,
and Iva
frutescens.
Fox et al.
(2012)
Estuarine tidal
marsh
Addition: none
S dep = not
reported
N dep = not
reported
On the Patuxent River,
pore water NCV-N and
salinity best explained
plant species richness
(model R2 = 0.67), with
NC>3"-N accounting for
26% of variation in the
full data set and for 5% of
variation in widely
distributed species. In
species with more
restricted geographic
ranges, N03"-N alone
best predicted plant
species richness
(positive relationship,
model R2 = 0.71).
Nanticoke
River and
Patuxent
River,
Chesapeake
Bay, MD and
DE
Plant community
Sharpe and
Baldwin
(2013)
Estuarine tidal Addition: 670 kg
marsh	N/ha/yr
S dep = not
reported
N dep = not
reported
Typha spp. cover
increased by 120%.
Nanticoke
River, MD
and DE
Plant community Baldwin
(2013)
Freshwater
estuarine
marsh
50, 200, or
1,200 kg N/ha/yr
as Nutralene
methylene urea
S dep = not
reported
N dep = not
reported
N addition (Nadd as kg
N/ha/yr) decreased the
dominance of E. fallax
(Efa, as % biomass)
Elfa = -0.00812 x
Nadd + 15.410.
Addition increased the
dominance of P.
punctatum (Ppu, as % of
biomass)
Ppu = 0.03144 x Nadd + 9
.353
Tchefuncte
River,
Madisonville,
LA
Oligohaline plant
community
dominated by
Sagittaria
lancifolia,
Eleocharis fallax,
and Polygonum
punctatum
Graham and
Mendelssoh
n (2010)
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Table 11-9 (Continued): Nitrogen loading effects upon plant biodiversity and
communities.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
Restored
freshwater
wetlands
Ambient
deposition; rates
not reported
Soil available N (ppm
ammonium, abbreviated
as Nnh4, or ppm nitrate,
as Nno3) correlated with
decreasing native
species richness
(number of species, y):
y = -73.8A/a/h4 +122,
r2 = 0.41
y = -30.9Nno3 + 97.4,
r2 = 0.18
Soil available N (as ppm
nitrate, or Nno3),
correlated with increasing
cover by all nonnative
plants (relative % cover
all non-natives, z):
z = AMNno3 -3.5,
r2 = 0.31
Soil available N (as ppm
nitrate, or Nno3),
correlated with increasing
cover by invasive
Phalaris arundinacea
(relative % cover,
abbreviated PHAR):
PHAR= 52.7/Va/03-29.6,
r2 = 0.58
28 wetlands
restored or
created in
1992-2002,
Illinois
All wetlands
surveyed
summer
2006
Vascular plant
community:
483 species
across all
wetlands,
including
109 non-native
species
Matthews et
al. (2009)
Ombrotrophic
peat bog
16 (low), 32
(medium), or 64
(high) kg N/ha/yr
as N (NH4NO3) or
NPK (NH4NO3 and
KH2PO4)
S dep = not
reported
N dep = 8 kg
N/ha/yr
In high N treatment,
Sphagnum cover
declined by 57%.
In all NPK treatments,
Sphagnum cover
declined over 8 yrto
1-25% of control, as
Polytrichum cover
increased by 5-1 Ox
control and then
declined.
Mer Bleue
Bog,
Ontario,
Canada
Bog plant
community:
dwarf shrub
species and
mosses:
Sphagnum
magellanicum,
Sphagnum
capillifolium, and
Polytrichum
strictum
Juutinen et
al. (2010)
Ombrotrophic
16 (low), 32
After 12 yr of low N,
Mer Bleue
Shrub species
Larmola et
peat bog
(medium), or 64
Sphagnum cover
Bog,
(Vaccinium
al. (2013)

(high) kg N/ha/yr
decreased 36%. After
Ontario,
myrtilloides,


as N (NH4NO3) or
5 yr of medium and high
Canada
Ledum


NPK (NH4NO3 and
N, Sphagnum cover

groenlandicum,


KH2PO4)
decreased 26 and 54%,
respectively.

Chamaedaphne
calyculata) and
mosses
(Sphagnum
magellanicum,
Sphagnum
capillifolium,
Polytrichum
strictum)

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Table 11-9 (Continued): Nitrogen loading effects upon plant biodiversity and
communities.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
Ombrotrophic
peatland
Addition = 16 kg
N/ha/yr (5N, or 5x
background
deposition), 32 kg
N/ha/yr (1 ON), or
64 kg N/ha/yr
(20N), all as
NH4NO3
Ambient (wet)
deposition = 8 kg
N/ha/yr and up to
0.26 kg P/ha/yr
N addition increased
cover of Chamaedaphne
calyculata by
202%(1 ON), and
decreased cover of moss
species by 63% (20N).
Mer Bleue
bog, Ottawa,
Canada
Evergreen
shrubs:
Chameadaphne
calyculata and
Rhododendron
groenlandicum
dominant; sparse
Kalmia
angustifolia
Deciduous
shrubs: Vacciniu
m myrtilloides
Mosses:
Sphagnum
capillifolium, S.
magellanicum,
and Polystrichum
strictum
Wang et al.
(2016a)
Calcareous,
rich fen
Addition = 50 kg
N/m2/yr as NO3" or
50 kg N/m2/yr as
NH4+
Ambient
deposition =
7-10 kg N/m2/yr,
based on Aherne
and Farrell (2002)
Sphagnum cover
increased, Scorpidium
cover decreased, and
vascular plant cover
increased in all plots
(including control plots)
over the course of the
study (2000-2004).
Scragh Bog,
Central
Ireland
Scorpidium
revolvens and
Sphagnum
contort um
Paulissen et
al. (2016)
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Table 11-9 (Continued): Nitrogen loading effects upon plant biodiversity and
communities.
Type of
Ecosystem
Additions or
Load (kg N/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
Fen
(transitional
mire)
Treatment,
2002-2009,
solution of
NH4NO3 added at
10 kg N/ha/yr
(Low N) or 30 kg
N/ha/yr (High N)
Deposition = not
reported
C. vulgaris cover does
not change in response
to low N addition, but 8 yr
of high N increased cover
(# intercepts) 96%, and
cover did not change 3 yr
after cessation of N
treatment (recovery
period).
S. fuscum cover does not
change in response to
low N, but 8 yr of high N
decreased cover (%)
45%, and cover
continued to decline
during recovery, when
cover was 60% less than
in control plots.
P. strictum cover
responded to 8 yr of low
and high N with
190-240% cover
increase, and although
this response declined in
recovery period, when
cover was 120-140%
higher than in control
plots.
M. caerulea cover did not
change in response to
low N, but increased to
5.6x control cover with
high N, and was 6.7x
control cover during
recovery period.
S. magellanicum cover
did not respond to low or
high N during 8 yr
experiment, but in
recovery, cover in plots
that had previously
received high N
increased to 4.3x control
plot cover.
Torbiera di
Passo San
Pellegrino,
Italy
Hummocks and
lawns:
13 species,
Calluna vulgaris,
Carex nigra,
Carex pauciflora,
Carex rostrata,
Eriophorum
vaginatum,
Molinia caerulea,
Polytrichum
strictum,
Sphagnum
angustifolium,
Sphagnum
fuscum,
Sphagnum
magellanicum,
Sphagnum
russowii,
Vaccinium
uliginosum,
Vaccinium vitis-
idaea
Lawns had an
additional
four species:
Nardus stricta,
Potentilla erecta,
Trichophorum
cespitosum,
Vaccinium
myrtillus
Gerdol and
Brancaleoni
(2015)
11.8.2. Consumers
Increased reactive N can alter dynamics among consumers in wetlands. Studies show that
increasing N load increases parasitism, decreases overall health condition, and increases
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the abundance of mosquitos (vectors for zoonotic diseases). There is also a new
meta-analysis evaluating consumer behavior in coastal wetlands. He and Silliman (2015)
conducted a meta-analysis of herbivory in globally distributed coastal wetlands (marshes
and mangroves) and found that N and N + P additions generally increased herbivore
abundance and herbivory across salt marsh studies, but not across mangrove studies.
In a survey of the native mud snail Ilyanassa obsoleta collected from 15 New England
salt marshes, increased N correlated with increasing parasitism of I. obsoleta. Sediment N
correlated positively with parasitic trematode prevalence (If = 0.41) and with infection
by the parasitic flatworm Stephanostomum spp. (If = 0.42).
Wigand et al. (2010) developed a metric for marsh condition using the Nitrogen Loading
Model (Wigand et al.. 2003) and data collected on abundance and richness of plants,
invertebrates, and soil conditions of 10 coastal marshes in Narragansett Bay, RI. Marsh
condition declined as annual N load to the marsh increased.
Ialeggio and Nvman (2014) collected plants dominant in freshwater (Panicum
hemitomon), freshwater-brackish (Sagittaria lancifolia), and brackish (Spartina patens)
marshes, fertilized the plants with 619 kg N/ha/yr in mesocosms, and then offered the
plants along with unfertilized controls in feeding trials to the introduced aquatic rodent
Myocastor coypus (nutria). Fertilization increased plant tissue N by 140%, from 0.9% N
to 2.2% N by mass, and nutria preferentially fed on fertilized plants. N loading increased
plant mass loss due to nutria herbivory by 750% (Ialeggio and Nvman. 2014).
N addition increases the abundance of mosquito species Culex tarsalis and Anopheles
hermsi, which act as vectors for zoonotic diseases. Duguma and Walton (2014)
constructed freshwater wetland mesocosms at UC Riverside, planting Schoenoplectus
maritimus in all mesocosms and monitoring plant growth as well as colonization of
mesocosms by mosquito species. N loads were 0, 32, 65, and 108 kg N/ha in 2009, and 0,
59, 119, and 198 kg N/ha in 2010. Sampling retrieved elevated numbers of Culex tarsalis
(western encephalitis mosquito) larvae and pupae in elevated nitrogen treatments in both
2009 and 2010, and of Anopheles hermsi in 2010. In 2009, immature mosquito counts
were 56-100% higher in fertilized treatments than in control treatments, but there were
no differences among different N loads in mosquito counts. In 2010, total mosquito
counts were similar in unenriched mesocosms and in mesocosms receiving 59 kg
N/ha/yr, but increased by 64% with 119 kg N/ha/yr, and by 129% with 198 kg N/ha/yr
(Duguma and Walton. 2014).
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11.9.
Critical Loads
In the 2008 ISA, no critical loads studies had been published. Since then, critical loads
have been published for wetlands in the U.S. (Greaver et al.. 2011).
11.9.1. Freshwater Wetland
There are two published critical loads for North American freshwater wetlands. Greaver
et al. (2011) determined that the critical load for altered peat accumulation and NPP is
between 2.7 and 13 kg N/ha/yr, based on observations from Aldous (2002a). Moore et al.
(2004). Rochefort et al. (1990). and Vitt et al. (2003). The upper end of this critical load
range is based on measurements of wet deposition only [10 to 13 kg N/ha/yr; (Aldous.
2002a. b) | and, therefore, does not reflect total N loading from all sources. There is
evidence showing that N deposition alters both the morphology and population dynamics
of the purple pitcher plant. The empirical evidence suggests a critical load to protect the
population of purple pitchers of 10-14 kg N/ha/yr (Gotelli and Ellison. 2006). while
matrix modeling to forecast long-term population sustainability based on observations of
population demographics suggests a lower value of 6.8 kg N/ha/yr (Gotelli and Ellison.
2002).
11.9.2. Coastal Wetlands
Coastal wetlands have open hydrologic and nutrient cycles that receive N loads from
sources other than atmospheric deposition, so a critical load for atmospheric N deposition
has been difficult to establish (Greaver et al.. 2011). Typically, the amount of N added in
experimental treatments in these systems simulates total N input and, therefore, far
exceeds the atmospheric deposition received by U.S. coastal wetlands. Only two studies
have considered N loads below 100 kg N/ha/yr. Based on the results ofWigand etal.
(2003). a critical load to protect the community structure of salt marshes is likely to be 63
to 400 kg N/ha/yr. Caffrev et al. (2007) provided additional evidence that 80 N/ha/yr
alters microbial activity and biogeochemistry. Greaver et al. (2011) also established a
critical load for eelgrass growing in estuarine and marine waters associated with coastal
wetlands (see Appendix 10.6). The critical load for coastal wetlands is based on total N
loading to salt marshes. Total loading includes N deposition directly to the marsh surface,
N deposited to the watershed and transported via surface or groundwater, and runoff from
agriculture, urban areas, or other sources. Additional experimental evidence on
ecosystem response to N loads more similar to N loading from atmospheric deposition is
needed to improve the critical load calculation for coastal wetlands in the U.S.
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11.9.3.
Comparison to Critical Loads from Europe
An assessment of N critical loads in Europe was published in 2003 and designated critical
loads for multiple types of wetlands, including raised and blanket bogs, poor fens, rich
fens, mountain rich fens, and intertidal wetlands (Bobbink et al.. 2003). There are
numerous publications on N effects to wetlands in Europe compared to the U.S. In
general, documented responses include effects on growth and species composition,
competition among species, peat and peat water chemistry, decomposition, and nutrient
cycling. A brief summary of the European critical loads for wetlands is presented here.
Bobbink et al. (2003) assigned a critical load of 5 to 10 kg N/ha/yr for bog ecosystems,
based on plant community and species responses to N deposition, and indicated that
precipitation and P limitation should be used to assign critical loads to individual sites.
The observed changes in the plant communities of ombrotrophic bogs included the
replacement of Sphagnum-form i ng species with nitrophilous moss species [20 to
40 kg N/ha/yr in Dutch bogs; (Grcvcn. 1992)1 and the absence of characteristic
Sphagnum species in British bogs [30 kg N/ha/yr; (Lee and Studholme. 1992)1. A recent
study along a deposition gradient in Alberta, Canada, compared Sphagnum bog and fen
responses in that region to previously published European bog and fen responses to N and
S deposition. Similar N deposition gradients elicited significant Sphagnum and water
chemistry responses in European peatlands but no responses in Albertan peatlands
(Wieder et al.. 2016). This suggests that critical loads for European Sphagnum species,
which have experienced a longer history of higher N deposition than North American
wetlands, may not be applicable in North American bogs and fens. Bobbink et al. (2003)
also suggested a critical load for the carnivorous bog plant roundleaf sundew
[10 kg N/ha/yr in Swedish bogs; (Bobbink et al.. 2003; Redbo-Torstensson. 1994)1 that
may be relevant to American bogs where this species also grows. The European critical
load for this carnivorous plant is similar to the range of critical loads suggested for
freshwater wetlands in the U.S.
There are European critical loads for fens, but they are based on moss decline, and the
amount of N causing ecological effects may not be directly relevant to U.S. ecosystems
for the reasons outlined by Wieder etal. (2016). European poor fens have a critical load
of 10 to 20 kg N/ha/yr based on increased sedge and vascular plants and negative effects
on mosses. The critical load for rich fens in Europe is 15 to 35 kg N/ha/yr based on
increased tall graminoids and decreased diversity. The critical load for montane rich fens
in Europe was 15 to 25 kg N/ha/yr based on increased vascular plants and decreased
bryophytes. These community changes are similar to the increasing vascular plant and
declining moss cover observed in American N addition experiments (see
Appendix 11.10). although there has been no critical load established for American bogs
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and fens based on these community composition changes. In Europe, changes in the
vegetation composition and structure likely affect fauna species assemblages, such as
ground-breeding birds, spiders, and beetles, living in the originally open bog vegetation.
Increased nutrient availability results in an increase of the nutrient content of plant
material (Limpens et al.. 2003a; Tomassen et al.. 2003) and algal growth (Limpcns et al..
2003b; Gulati and Demott. 1997). which affects herbivorous, detritivorous, and
carnivorous invertebrates (van Duinen et al.. 2004).
The European critical load for salt marshes, based on expert judgment, is 30 to
40 kg N/ha/yr (Bobbink et al.. 2003). but studies of European salt marshes are limited.
There are no European wetland systems composed of native plants directly equivalent to
North American low salt marshes; in European systems, these zones consist of
unvegetated mud flats, or of monocultures of invasive plants. High levels of N input (65
to 70 kg N/ha/yr) significantly increased biomass production in the Netherlands (van
Wijnen and Bakker. 1999); however, no changes in species composition and in diversity
have been observed at deposition of 15 to 25 kg N/ha/yr at sites in the Netherlands and
Germany (Bobbink et al.. 2003). The critical load for N deposition in North American
coastal wetlands may be close to these values when considered as part of the total N load
that these systems receive.
11.10. Summary
There are chemical and biological effects of N deposition in wetlands. Wetland
vegetative communities are adapted to high levels of natural organic acidity, so it is
unlikely that S or N deposition would cause any acidification-related effects at levels of
acidic deposition commonly found in the U.S. IYU.S. EPA. 2008a). in Annex B],
Wetlands can be sensitive to eutrophication caused by N deposition. These effects were
documented in the 2008 ISA. In this appendix, new information published since 2008 is
integrated with the information published in the 2008 ISA. Synthesis is made across
wetland types for causality statements. Information is also synthesized for freshwater and
coastal wetlands because saltwater and estuarine wetlands are typically adapted to much
higher N loading.
11.10.1. Causality across Wetland Types
New studies published between 2008-2015 support and extend the findings of the 2008
ISA. As in the 2008 ISA, the body of evidence is sufficient to infer a causal
relationship between N deposition and the alteration of biogeochemical cycling in
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wetlands. N deposition changes N and C cycling in wetlands. There is new evidence of
N deposition changes to plant physiology, which was not included in the 2008 ISA as an
endpoint due to lack of evidence. This evidence expands our understanding of the causal
relationship between N deposition and species diversity. The body of evidence is
sufficient to infer a causal relationship between N deposition and the alteration of
growth and productivity, species physiology, species richness, community
composition, and biodiversity in wetlands.
N deposition contributes to total N loading in wetlands, with the relative contribution of
deposition to total loading variable among wetland types. Chemical indicators of N
deposition in wetlands include NOs and NH4+ leaching, DON leaching, N
mineralization, denitrification rates, and N2O emissions. A wetland can serve as a source,
sink, or transformer of atmospherically deposited N (Dcvito et al.. 1989). Source/sink N
dynamics in wetlands vary with season, hydrological conditions, vegetation type, climate,
and surface geology (Mitchell et al.. 1996; Arhcimcr and Wittgrcn. 1994; Koerselman et
al.. 1993; Devito et al.. 1989).
Several new studies provide evidence of N deposition alterations to biogeochemical
cycling of N within a wetland ecosystem (Appendix 11.3). A synthesis study of multiple
wetland types across the globe shows wetland N removal is proportional to N load (log[N
removal] = 0.943 x log[N load] - 0.033 N) and removal efficiency is 26% higher in
nontidal than tidal wetlands (Jordan et al.. 2011). Meta-analysis of 19 N addition
observations (N addition 15.4 to 300 kg N/ha/yr) found that N enrichment increased
wetland N2O emissions by 207% (Liu and Greaver. 2009). The emission factor for the
amount of N added to a wetland that is converted to N2O ranges between 0.007-0.07 (Liu
and Greaver. 2009).There are also new studies that evaluate the effects of N loading/N
addition on other endpoints related to N cycling in peat bog, riparian, mangrove, and salt
marsh wetlands. The endpoints evaluated include ecosystem N retention, wetland export
of N to surface waters, N fixation, N mineralization, denitrification, emission ofN20; and
bacterial abundance, activity, and composition in wetland soils. The results of North
American studies are summarized in Figure 11-2. Across studies, N loading decreases the
ability of wetlands to retain and store nitrogen, which may diminish the wetland
ecosystem service of improving water quality.
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¦f N03- and NH4+ in peat A
M
O
CQ
sl/ N retention efficiency B
0
c
.ro
fU
Ł sl/ BNF rates, sl/N-fixing symbionts c
100

o
bO
™ 'T* Denitrification rate D
100
"5 n)/ N mineralization E
fU
\1/ N retention tidal N export) F
-t—'
1/1
fU
° \1/ N retention (^porewater NH4+, ^N20 emissions) G
P. australis expansion and^porewater NH4+ h
100
180
250
250
0 100 200 300 400 500
N addition (kg N/ha/yr)
BNF =biological nitrogen fixation; ha = hectare; kg = kilogram; N = nitrogen.
aPinsonneault et al. (2016).
"Xing et al. (2011).
cRuess et al. (2013).
dWhiqham et al. (2009).
eVivanco et al. (2015).
fBrin et al. (2010).
gPastore et al. (2016).
"Mozdzer et al. (2016).
Figure 11-2 Summary of the levels of nitrogen addition where a change to
nitrogen cycling is observed.
1	There is new information on how N deposition alters biogeochemical cycling of C in
2	wetlands. In the 2008 ISA, evidence from Canadian and European peatlands showed that
3	N deposition had negative effects on Sphagnum bulk density and mixed effects on
4	Sphagnum productivity depending on the history of deposition. In Canadian
5	ombrotrophic peatlands experiencing deposition of 2.7-8.1 kg N/ha/yr, peat
6	accumulation increased with N deposition, but accumulation rates had slowed by 2004,
7	indicating a degree of N saturation (Moore et al.. 2004). N addition alters belowground
8	and aboveground pools of carbon and also increases wetland methane emissions, as
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summarized by meta-analysis in the 2008 ISA and published in Liu and Greaver (2009).
In a study of 90 wetlands around the Gulf Coast from Florida to Texas, Nestlerode et al.
(2014) found that higher N in the soil correlated with lower bulk density (ln[soil
%N] = -1.9233 x [g/cc] + 0.4165) which indicates that the marsh is less resilient to
physical stresses from tidal or storm flooding, and the marsh platform is more likely to
shear off or wash away. New evidence shows that N loading increases methane
production from wetland soils. In N addition experiments replicated in three California
marshes (Vivanco et al.. 2015; Irvine et al.. 2012). field-measured methane flux from
soils increased linearly with increasing N load (0, 100, 200, 400, 800, 1,600,
3,200 kg N/ha/yr), such that CH4 flux from soils increased by 1.23 |ig CHVmVday for
each 10 kg N/ha/yr added, (Vivanco et al.. 2015; Irvine et al.. 2012). Methane flux
increased from low or nearly negative to net positive values above -100 kg N/ha/yr,
indicating a shift from net methanotrophy to methanogenesis in the microbial community
in response to N loading consistent with other studies (Liu and Greaver. 2009).
Additional literature evaluates the effects of N deposition, N loading, or experimental N
addition on C cycling in salt marshes, mangroves, freshwater tidal marshes, riparian or
intermittent marshes, bogs, and fens. Significant effects of N loading upon
biogeochemical cycling of C in North American wetlands (in which the N addition was
500 kg N/ha/yr or lower) are summarized in Figure 11-3. Ecological endpoints evaluated
to assess N loading effects on C cycling include (1) measures of C pools—plant
aboveground biomass and productivity, plant belowground biomass of roots and
rhizomes, soil organic matter, and total soil C or peat C; (2) measures of C fluxes—
microbial mineralization, decomposition rates, CO2 emissions, and CH4 emissions; and
(3) measures of physical marsh stability as a proxy for long-term C storage—peat
chemical composition, temperature, bulk density, physical resistance, and soil water
content. In general across wetlands, nitrogen loading increases aboveground C pools,
decreases or does not change belowground C pools, and does not change or increases C
fluxes. This shift from belowground to aboveground C storage may diminish the wetland
ecosystem services of long-term carbon storage and flood protection, as well as reduce
the stability and persistence of wetlands on the landscape. N loading effects upon C
cycling are uneven across species, which may affect wetland biodiversity and the wetland
ecosystem service of provisioning.
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s
Ł*5
*
AG: 4-3-21% growth 5. purpurea (STE)
BG: 4" 17% biomass in top 10 cm peat
BG: P and K limitation of microbial activity
BG: altered peat temperature, •!¦¦ peat [C02]
AG: ^ 30% biomass of moss community
AG: 1s 105% biamass C. catyculata (STE)
AG: -T1 600% biomass C. oligosperma (STE)
AG: i" 2.6 times productivity plant community
BG: 4- 46% net ecosystem C exchange
BG: 4-51% established root biomass
ANPP responsiveto N
EG: 4^ 71% rhizome biomass
BG: 4^ 33% macro-organic matter
AG: ^ 1.4-3.8 times biomass Z. miliacea (STE)
A


B
0

C
0

C
0

c
0

a
0

0
0

E
0
E

60
I

60
F

[m
G


H


i
I.J


500
500 I
500
ip
5
Ł
Ł
AG: ^ biomass A. germinans
AG: ^ biomass of 5. depressa
M
AG: ^"3-14% shoot mass in high marsh
AG: 1s 36-53% biomass of 5. parifka*
BG: 4- 42-34% fine root production
O.P
i
BG: T* roots of invasive P. australis
p
AG: ^ 25-7596 biomass colonizing plants
AG: 'V129% biomass of C4 plants S. patens and D. spicataC
AG: 4^55% bioma5sofC3 5. americanusC
AG: ^ 47-79% plant community biomass*
AS
[250J
f==°1
S
I 250 I
N addition (kg N/ha/yr)
AG = aboveground; ANPP = aboveground net primary productivity; BG = belowground; C02 = carbon dioxide; FW = freshwater;
ha = hectare; kg = kilogram; N = nitrogen; NPP = net primary productivity; STE = state-listed threatened or endangered species;
yr = year.
ACrumlev et al. (2016); BWendel et al. (2011); cPinsonneault et al. (2016); DXing et al. (2011); Elversen et al. (2010); FLarmola et al.
(2013); GGraham and Mendelssohn (2016); "Graham and Mendelssohn (2010); 'Ket et al. (2011); JFrost et al. (2009); KWhiqham et
al. (2009); LVivanco et al. (2015); "(Johnson et al.. 2016a); "Goldman Martone and Wasson (2008); "Lanalev and Meaoniaal (2010);
pLanalev and Meaoniaal (2012); "Mozdzer et al. (2016) RLanalev et al. (2013); sLanalev et al. (2009).
Grey boxes with black borders represent ambient deposition studies, and white boxes with blue borders represent field nitrogen
addition experiments.
Figure 11-3 Summary of new literature of nitrogen load effects on
belowground and aboveground carbon cycling.
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N addition effects on plant stoichiometry and plant architecture were not addressed in the
2008 ISA. Plant stoichiometry theory considers the balance of multiple chemical
elements in living tissues, and plant architecture is defined as the three-dimensional
organization of the plant body. The stoichiometry of plant tissue is often connected to the
tissue's physiological function. Most new studies on physiology and stoichiometry have
been conducted in bogs and fens, where N addition typically caused an increase in plant
N content, a decrease in N use efficiency and resorption, and an increase in plant
production, particularly of vascular plants. At high N loads or cumulative exposure to
years of lower N loads plants may experience leafN saturation and micronutrient
limitation (e.g., P, K, and Ca, indicated by altered leaf tissue concentrations or altered
reabsorption efficiencies; see Figure 11-4). which in turn cause leaf damage (Bubicr et
al.. 2011; Xing etal.. 2011) or decreasing plant abundance of sensitive species (see
biodiversity section). N loading can disrupt nutrient acquisition of carnivorous plants via
their adaptation of capturing and digesting insects; for example, along a narrow N
deposition gradient (3.4-5.0 kg N/ha/yr) across bogs in the Adirondack Mountains,
purple pitcher plant (S. purpurea) experienced negative effects upon growth and
insectivory at deposition of 4.4-4.9 kg N/ha/yr (Crumley et al.. 2016). A new European
study also found negative effects of N deposition upon insectivory of the wetland plant
Drosera rotundifolia at 3.81-11.30 kg N/ha/yr (Millctt et al.. 2012). Historically low N
loads to bogs and fens have made these wetlands and their endemic plant species
particularly sensitive to N deposition. New studies of physiology and plant architecture in
North American riparian wetlands, freshwater tidal marshes, mangroves, and salt marshes
showed that N addition in these systems increased plant tissue N with a cascading effect
that increased plant primary production and changed plant architecture. N loading
increased plant height in salt and freshwater marshes, which Deegan et al. (2012) showed
made salt marshes top-heavy and resulted in a loss of marsh platform stability, with and
shredding and loss of marsh surface. In general effects of N loading upon physiology and
architecture vary by species across North American wetlands. Even species tolerant of N
loading may experience negative effects if N loading causes physiological limitation by
other nutrients, including Ca, P, and K. Plant physiology and architecture altered by N
loading may affect wetland biodiversity and resiliency to other disturbances.
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4/40% insectivory in S. purpurea (STE) |	|
^ 27% leaf N in shrub C. calyculata (STE) | |
•4/ 15% leaf C:N, 4/ 34% leaf Ca, ^ amino acids in shrub C. calyculata (STE) [»D
4' 23% leaf Ca, 4' 55% P, 4^ 47% K in shrub V. myrtilloides (STE) 1161
^ 100% Vcmax in shrub R. groenlandicum (STE) QU
4^29-62% N uptake in moss S. magellanicum 1201
4/ 34% tissue Ca in mosses 1321
1* 42% P resorption in shrub C. calyculata (STE) \El
"Is 12 times P resorption, ^ 92-123% K resorption in R. groenlandicum (STE)
4/ 81% NUE, 4^ 91% NRE in shrub C. calyculata (STE)	[io]
4^ 89% NUE, 4/ 84% NRE in sedge C. oligosperma (STE)	| 601
¦T N uptake by grass C. canadensis	1601
4/45% NRE in tree A. rugosa	| 601
4^ 42%tissue Ca, "T 39%tissue P in mosses	1641
^ 16% leaf N, 4^ 54% leaf P, /t" glutamic acid, in shrub R. groenlandicum (STE)	1641
4/ 23% leaf Mg, 4- 186% Mg resorption in R. groenlandicum (STE)	1641
^ 84% leaf chlorophyll, /\* amino acids in shrub C. calyculata (STE)	1641
¦T P resorption 33%, 4/ 22% leaf Ca, 4/ 45% Mn in shrub C. calyculata (STE)	1641
Shift in bog seasonal gross PSN: 4^ summer PSN, ^ fall PSN	1641
10% leaf N , 4/ 21% P resorption in tree A. tenuifolia
AG plant tissue [N], in sedge B. maritimus (STE) 1321
planttissue N, plant N:P, 4^ NRE, 4^PRE in forb S. lancifolia	1501
"T 26% BG plant tissue [N] in sedge B. maritimus (STE)	1651
5% leaf construction cost in native haplotype of grass P. australis
25-43% leaf N, 4^ 18-28% leaf C:N, 1s 25-61% leaf N:P in tidal FW marsh
^ leaf N in succulent forb S. depressa	| ioo|
0	100
kg N/ha/yr
AG = aboveground; BG = belowground; C = carbon; Ca = calcium; FW = freshwater; K = potassium; Mg = magnesium;
Mn = manganese; N = nitrogen; NRE = nitrogen resorption efficiency; NUE = nutrient use efficiency; P = phosphorus; PP = primary
productivity; PRE = phosphorus resorption efficiency; PSN = photosynthesis; STE = state-listed threatened or endangered species;
Vcmax = maximum velocity of carboxylation.
Grey boxes with black borders represent ambient deposition studies, and white boxes with blue borders represent field nitrogen
addition experiments.
Figure 11-4 Summary of the level of nitrogen load that caused a change in the
response variables of plant stoichiometry and physiology in
wetlands.
1
2	The 2008 ISA found that the evidence was sufficient to infer a causal relationship
3	between N deposition and the alteration of species richness, species composition, and
4	biodiversity in wetland ecosystems. Wetlands provide habitat to a disproportionally high
5	number of rare plants given their relative area; for example, fens occupy 0.01% of the
6	land area of northeast Iowa yet contain 17% of the endangered, threatened, or listed
7	species of concern in the area. The 2008 ISA noted that there were 4,200 native plant
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species in U.S. wetlands, 121 of which are federally endangered. Wetland species have
evolved under N limited conditions, including endangered species in the Isoetes
(3 endangered species) and Sphagnum (15 endangered species) genera, as well as the
endangered insectivorous plants Sarracenia oreophila and Drosera rotundifolia. Plant
demography was not specifically addressed in the 2008 ISA, although demography is an
important aspect in maintaining species richness and biodiversity. Plant demography is
the change in plant population size and structure through time. In bogs and fens, N
addition experiments suggest positive population growth rates for Sarracenia purpurea at
0 or 1.4 kg N/ha/yr, but population losses at 14 kg N/ha/yr (Gotclli and Ellison. 2006).
Demographic modelling of S. purpurea found that populations remained stable for
100 years at deposition rates of 4.5-6.8 kg N/ha/yr, but extinction risk for populations
increased above 6.8 kg N/ha/yr (Gotclli and Ellison. 2002). Field N addition studies
suggest that N addition to bogs and fens will also affect community composition
(Figure 11-5). In Mer Bleue Bog, Ontario, N addition decreased the dominance of
Sphagnum mosses in the northern ombrotrophic peat bog, while increasing dominance of
woody shrub species (Larmola et al.. 2013; Juutinen et al.. 2010). This result is consistent
with the earlier European literature summarized in the 2008 ISA, which showed that N
deposition decreased moss dominance and increased the dominance of vascular plants in
European bogs and fens.
Across freshwater wetlands, N load is correlated with an increase in the abundance of
invasive plant species and a decrease in the number of native plant species (see
Figure 11-1 and Appendix 11.8.1). as well as an increase in larvae of mosquito species
that are vectors for zoonotic diseases (Appendix 11.8.2). In a freshwater tidal marsh on
the Tchefuncte River, LA, N addition shifted the relative dominance of the perennial
wetland-obligate monocots dominant at the site (Graham and Mendelssohn. 2010). In salt
marshes, N addition caused negative reproductive effects in invasive Spartina foliosa and
Spartina hybrids (Tyler et al.. 2007). and a positive effect on reproduction in Salicornia
bigelovii (Bover and Zedler. 1999). N loading in salt marshes is correlated with species
richness. A study on the Patuxent River found marsh soil NOs -N and pore water salinity
best explained variation in plant species richness out of all factors considered (R2 = 0.67),
with NO3 -N accounting for 26% of the variation in total species richness., In the same
system, the best model for rare species richness contained only soil NO3 -N as an
explanatory variable, with a positive relationship between NO3 -N and species richness
[i?2 = 0.71 (Sharpe and Baldwin. 2013)1. In salt marshes in Maryland and California, N
addition was shown to alter the relative abundance of plant species (Langlev and
Megonigal. 2010; Goldman Martone and Wasson. 2008) (Figure 11-5). N deposition
effects can cascade beyond plants up trophic levels to consumers. Meta-analysis showed
that N loading increased invertebrate abundance and herbivory, and a field experiment
found increased herbivory by the invasive mammal nutria in freshwater tidal and salt
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1	marshes (Appendix 11.8.2). In New England, N loading decreased salt marsh condition,
2	evaluated by a multidimensional metric that included abundance and richness of plants
3	and invertebrates as well as soil conditions (Appendix 11.8.2). In general, N loading
4	across North American wetlands decreases richness or abundance of sensitive species
5	while promoting invasive species, with negative effects on biodiversity.
CuO
o
CO
tc
o
u
4/ 36% cover of mosses S. magelj^inicum and...
1* 202% cover of shrub C. calyculata (8 yr Ex) B
4/ 26% cover of mosses S. mageHanicum and...
4/ 63% cover of mosses (8 yr Ex) B
4/ 54% cover of mosses S. mage^lanicum and...
4/ 57% cover of mosses S. magellarjicum and S....
4/ relative dominance of sedge E. fallax D
'T relative dominance of forb P. punctata D
/t" 80% cover of non-native upland plant species E
4/ 52% cover of native marsh plant species E
'T 129% cover of C4 grasses S. patens and D....
4/ 52% cover of C3 sedge S. americanus F
100
250
250
200 300
kg N/ha/yr
400
500
Ex = experimental exposure length; FW = freshwater; h = hectare; kg = kilogram; N = nitrogen; yr = year.
ALarmola et al. (2013).
BWang et al. (2016a).
cJuutinen et al. (2010).
"Graham and Mendelssohn (2010).
EGoldman Martone and Wasson (2008).
FLanalev and Meaoniaal (2010).
Numbers indicate the lowest addition level in which change is observed.
Figure 11-5 Summary of nitrogen addition studies on wetland biodiversity.
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11.10.2. Coastal versus Freshwater Wetlands
Coastal and freshwater wetlands tend to have different sensitivity to N
deposition/loading. The effect of N deposition on wetland ecosystems depends on the
fraction of rainfall in the total water budget, and the sensitivity to N deposition has been
suggested as: bogs (70-100% rainfall) > fens (55-83% rainfall) > coastal wetlands
[10-20% rainfall (Morris. 1991)1.
Greaver et al. (2011) suggested a critical load to protect biodiversity and biogeochemistry
of coastal wetlands based on salt marsh community composition, microbial activity, and
biogeochemistry (63-400 kg N/ha/yr as total N load). A comparison of those critical
loads with data on N addition levels (15-500 kg N/ha/yr) and associated effects,
published since the last ISA, is given in Figure 11-6.
In freshwater systems, Greaver et al. (2011) determined a critical load for wetland C
cycling as quantified by altered peat accumulation and NPP, between 2.7 and
13 kg N/ha/yr. Greaver et al. (2011) also set a critical load to protect biodiversity based
on morphology and population dynamics of the purple pitcher plant (Sarracenia
purpurea) between 6.8-14 kg N/ha/yr. A more recent study across an N deposition
gradient suggests that purple pitcher plant populations experience negative effects of N
deposition at 4.4 kg N/ha/yr (Crumley et al.. 2016). A comparison of freshwater wetland
CLs to data from N addition levels (16-500 kg N/ha/yr) and associated effects, published
since the CLs for wetlands were set, is given in Figure 11-7. There is information on the
relationship between N addition and numerous endpoints. At the lowest experimental
addition level (16 kg N/ha/yr), there are observations of altered C and N cycling and
altered biodiversity. The endpoints affected include decreases in moss cover, increased
peat biomass, decreased N retention efficiency, and altered/damaged leaf stoichiometry in
vascular plants.
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aj



o
Denitrification rate
\wo]

c
ra
2
AG: ^ biomass A. germinans
1100 I


4- N mineralization
[ 100 I


't methane emissions
1100 I


AG: biomass of succulent forb S. depressa
1ml


Is leaf N in S. depressa
1100 I


1s shoot mass in high marsh
140


AG:biomass of S. pacifica
150


'T" cover of non-native upland plant species
150

-C
4- cover of native marsh plant species
150

ra
5
4- N retention ("T1 tidal IM export)
M

a:
ra
4- N retention (^porewater NH4+, *tN20 emissions)
250

l_)
BG: 4, native fine root production
jlso'


BG: T" roots of invasive P. australis
no


AG: 1s plant community biomass
250


AG: T1 biomass of C4 plants S. patens and D. spicata
I2501


AG: 4- biomass of C3 S. americanus
(13


colonization by invasive P australis
HED


"t1 cover of C4 grasses S. patens and D. spicata
250


4/ cover of C3 sedge S. americanus
0§]
a: 3
-a 2


t- Ł
ID "C
U
3
c
Community structure, microbial activity, biogeochemistry
63-400
0 100 200 300 400 500
kg N/ha/yr
AG = aboveground; BG = belowground; CH4 = methane; GTR = general technical report; N = nitrogen.
Figure 11-6 Summary of field nitrogen addition studies for coastal wetlands (blue
borders) versus critical load (black border).
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BO .52
O J3 ¦
CO g
a 2'
4^growth of S. purpurea
4*40% insectivory in S. purpurea (STE)
1s 27% leaf N in shrub C. calyculata (STE)
4, N retention efficiency of bog
peat DIN
BG: biomass in top 10 cm peat
BG: recalcitrantC in peat
BG: T microbial metabolism of C
BG: P and K limitation of microbial activity
BG: altered peat temperature, 4- peat [C02]
AG: 1* biomass of moss community
4-leaf C:NP 4-leaf Ca, "t1 amino acids in shrub C. calyculata...
4, leaf Ca, 4/leaf P, 4'leaf K in shrub V. myrtilloides (State T& E)
1"" Vcmax in shrub R, groenlandicum (STE)
4-cover of mosses S. magellanicum and capillifolium(12 yr Ex)
4' N uptake in moss S. magellanicum
4>tissue Ca in mosses
1" P resorption in shrub C. calyculata (STE)
resorption, 'T* K resorption in R. groenlandicum (STE)
4' cover of mosses S. magellanicum and capillifolium (7 yr Ex)
4* NUE, 4- NRE in shrub C. calyculata (STE)
4' NUE, 4- NRE in sedge C. oligosperma (STE)
1" N uptake by grass C. canadensis
4- NRE in tree A, rugosa
AG: 't biomass of grass C. calyculata (STE)
AG: T biomass of sedge C. oligosperma (STE)
AG: "T productivity plant community
4/tissue Ca, 'f tissue P in mosses
^ leaf N, 4-P, 'T glutamic acid, in R. groenlandicum (STE)
4' leaf Mg, 4. Mg resorption in R. groenlandicum (STE)
T* leaf chlorophyll, amino acids in shrub C. calyculata (STE)
1" P resorption, 4- leaf Ca, 4' Mn in shrub C. calyculata (STE)
BG: 4» net ecosystem C exchange
Shift in bog seasonal gross PSN: 4' summer PSN, 1" fall PSN
4' cover of mosses S. magellanicum and capillifolium (4 yr Ex)
4' BNF rates, 4-N-fixingsymbionts
leaf N, 4- P resorption in tree A. tenuifolia
4/ SLM in tree A. tenuifolia
I" AG plant tissue [N], in sedge B. maritimus (STE)
"T larval and pupal mosquito abundance
T plant tissue N, T N:P, 4' NRE, 4-PRE in forb S. lancifolia
4- relative dominance of sedge E. fallax
'f relative dominance of forb P. punctata
BG plant tissue [N] in sedge B. maritimus (STE)
"T total mosquito abundance
ANPP responsive to N
^ leaf construction cost of grass P. australis
BG: 4- rhizome biomass
BG: 4' macro-organic matter
AG: -f biomass Z. miliacea(STE)
T1 leaf N, sb leaf C:N, 1" leaf N:P in tidal FW marsh
peat accumulation and NPP
Pitcher plant community change
16
16
16
16
16
16
16
16
16
16
16
16
|20|
I<200I
1250]
kg ^/fia/yr
AG = abovegrourid; ANPP = aboveground net primary productivity; BG = beiowgrourid; BNF = biological nitrogen fixation;
C = carbon; Ca = calcium; C02 = carbon dioxide; Ex = experimental exposure length; FW = freshwater; GTR = general technical
report; K = potassium; Mg = magnesium; Mn = manganese; N = nitrogen; NPP = net primary productivity; NRE = nitrogen resorption
efficiency; NUE = nutrient use efficiency; P = phosphorus; PRE = phosphorus resorption efficiency; PSN = photosynthesis;
SLM = specific leaf mass; STE = state-listed threatened or endangered species; Vcmax = maximum carboxylation velocity; yr = year.
Values indicate biotic or chemical changes observed in response to experimental nitrogen addition (boxes with blue borders) or an
ambient gradient of N deposition (boxes with black borders).
Figure 11-7 Summary of nitrogen load studies for freshwater wetlands as well
as current critical loads.
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APPENDIX 12. NONACID SULFUR ENRICHMENT
EFFECTS
Appendix 12 summarizes research on the nonacidifying impacts of S deposition on
ecosystems, synthesizing literature published since the 2008 ISA for Oxides of Nitrogen
and Sulfur-Ecological Criteria (hereafter referred to as the 2008 ISA) with earlier key
studies. The causal statements are presented in the Introduction (Appendix 12.1).
Appendix 12.2 discusses the effects of sulfur (S) deposition on S storage and cycling in
ecosystems, sulfide phytotoxicity in wetlands, internal eutrophication in freshwater
aquatic systems, methane emissions from wetlands and lakes, and the microbial
communities responsible for methanogenesis.
Increasing bioavailability of mercury (Hg) due to S addition is a major focus, and Hg
sources, pools, and fluxes in terrestrial and aquatic ecosystems are briefly described in
Appendix 12.3.1 and in the Supplemental Material. Appendix 12.3.2 includes a
discussion of the current understanding about the distribution of Hg methylation potential
across the prokaryotic domains and the key role of sulfur-reducing prokaryotes (SRPs).
Mercury methylation occurs in lakes in shallow sediments or below the oxycline; in peat
wetlands, at the interface with upland areas and generally near the water surface in peat;
in periphyton; in saturated decomposing litter or algal biomass; in saturated forest soils;
and in estuarine and marine sediments. In freshwater systems, methylation is strongly
seasonal, with methylmercury (MeHg) concentrations peaking in summer or fall.
Methylation rates and MeHg concentrations are determined by the biological niche and
environmental requirements of microbial methylators; the current state of our
understanding of those environmental requirements is briefly summarized in
Appendix 12.3.3.
In Appendix 12.3.4. multiple lines of evidence are presented establishing the relationship
between SOx deposition and increases in bioavailable Hg in the environment, including
experimental S addition experiments and case studies from well-studied wetlands where
S inputs are primarily agricultural. Appendix 12.3.5 considers observational evidence of
the correlations between SOx deposition and Hg burdens in fish. Appendix 12.3.5 also
presents the results of observational studies in prairie pothole wetlands, peat bogs,
freshwater marshes, streams, and rivers that show correlations between ambient sulfate
and MeHg concentrations in water and sediment samples. Appendix 12.4 summarizes
studies that find higher Hg burdens in mosquitos and fish from ecosystems with increased
experimental or anthropogenic S loading. Ecosystems especially sensitive to the effects
of S deposition are described in Appendix 12.5.
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Appendix 12.6 describes critical loads for Hg from European systems; at this time, there
are no established critical loads for nonacid S effects in North American ecosystems.
Appendix 12.7 is a summary including causal determinations based on the synthesis of
new information and previous evidence.
Introduction
In the 2008 ISA, the evidence was sufficient to infer a causal relationship between S
deposition and increased methylation of Hg in aquatic environments where the value of
other factors was within adequate range for methylation. Direct atmospheric deposition of
S to wetland and aquatic systems, as well as sulfate leaching from terrestrial systems with
current or historical S deposition, alters biology and chemistry of these systems.
Nonacidifying effects of S deposition in these nonterrestrial systems include toxicity to
wetland plants, increased phosphorus (P) availability leading to internal eutrophication,
altered methane emissions, and increased Hg methylation and Hg concentration in biota.
Consistent with the causal statement in the 2008 ISA, the body of evidence is sufficient
to infer a causal relationship between S deposition and the alteration of Hg
methylation in surface water, sediment, and soils in wetland and freshwater
ecosystems. The 2008 ISA described sulfide phytotoxicity in European systems, but
there is more recent research demonstrating sulfide phytotoxicity under current
conditions in North American wetlands. The body of evidence is sufficient to infer a
causal relationship between S deposition and changes in biota due to sulfide
phytotoxicity including alteration of growth and productivity, species physiology,
species richness, community composition, and biodiversity in wetland and
freshwater ecosystems.
S deposition contributes to Hg accumulation by stimulating the activity of
anoxic-sediment-dwelling sulfur-reducing bacteria (SRB), which convert inorganic Hg to
MeHg in the course of metabolism. The 2008 ISA described the activity of SRB, but
more recent research indicates that S reducing archaea are also active in wetland
sediments, which is why this document will use the broader term sulfur-reducing
prokaryotes (SRPs) to denote both bacteria and archaea involved in S reduction.
Laboratory and mesocosm-scale experiments reviewed in the 2008 ISA established that
only trace amounts of MeHg could be produced in the absence of sulfate. These results
were confirmed with larger scale observational studies. MeHg enters the food chain and
bioaccumulates in higher trophic levels, and research summarized in the 2008 ISA
suggested that numerous wild populations of fish, birds, and mammals experienced
MeHg exposures high enough to cause substantial reproductive, behavioral, or health
impairment.
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At the time of the 2008 ISA, dose-response relationships between S deposition and
Hg-methylation rates had not been established, in part because oxygen content,
temperature, pH, and labile carbon supply also control the rates of the activity of SRB in
aquatic environments. Watersheds with conditions known to be conducive to Hg
methylation were identified in the northeastern U.S. and southeastern Canada, although
significant biotic Hg accumulation had been observed in other regions that had not been
studied as extensively. The U.S. EPA set the fish tissue criterion in 2001 for MeHg at
300 ng/g fish tissue (reported as 0.3 mg/kg) for the protection of human health, which
resulted in 2,436 fish consumption advisories for Hg in 2004, 2,682 in 2005, and 3,080 in
2006. Forty-eight states, one territory, and two tribes had issued Hg fish-consumption
advisories at the time of the 2008 ISA.
Appendix 12 summarizes research on the nonacidifying impacts of S deposition on
ecosystems, synthesizing literature published since the 2008 ISA with earlier, key studies
(Figure 12-1).
CH4 = methane; Fe = iron; Hg = mercury; MeHg = methylmercury; P = phosphorus; S042 = sulfate; SOx = sulfur oxides.
Figure 12-1 Effects of sulfur oxide deposition on chemical (blue boxes),
biological (green boxes), and atmospheric (yellow boxes)
indicators of ecosystem change, as documented by the previous
Integrated Science Assessment and more recent research.
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12.2.
Ecosystem Effects of Altered Sulfur Cycling
This section provides background on the sulfur cycle in terrestrial and aquatic ecosystems
(Appendix 12.2.1). It describes effects of S enrichment (SOx deposition or experimental
S amendments) upon aquatic S cycling (Appendix 12.2.2). sulfide toxicity on wetland
plants (Appendix 12.2.3). S enrichment on phosphate cycling in aquatic ecosystems and
on uptake of toxic elements by aquatic plants (Appendix 12.2.4). and S enrichment in
altering microbial competition and methanogenesis (Appendix 12.2.5).
12.2.1. The Sulfur Cycle
S cycling in terrestrial ecosystems is relatively well characterized compared to S cycling
in aquatic systems. Sulfate is the dominant S source in supplying plant nutrient S
demand. Sulfate in the soil solution, whether its source is weathering of sulfur-containing
minerals, mineralization of internal ecosystem S, or atmospheric deposition of SOx, has
several possible routes through the S cycle (Figure 12-2).
Wet
Deposition
Dry
Deposition
Litter
Inputs
Uptake
Desorption
Source: Adapted from Mitchell et al. (2011).
Figure 12-2 Sulfur cycle in terrestrial, forested ecosystems.
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Sulfate can be immobilized by incorporation into living cells (microbes, fungi, plants),
which ultimately end up in the organic soil S pool. In most North American terrestrial
ecosystems, total S pools in soil greatly exceed S stored in living biota (Figure 12-2).
with soil S in organic or inorganic molecules. The largest pool of S in soil is organic,
typically comprising more than 95% of total soil S in temperate ecosystem soils (Wang et
al.. 2006; Solomon et al.. 2001). although in drained wetland soils of the Florida
Everglades Agricultural Area (EAA), the organic fraction was quantified as 87% of total
S (Ye et al.. 2010). Organic S pools in grassland soils are negatively correlated with mean
annual temperature in the North American Great Plains, indicating that climate is a strong
control on S cycling (Wang et al.. 2006). Organic soil S is found in ester-bound forms
(produced via microbial activity) or bound directly to carbon (C) in organic compounds
of varying sizes [whose ultimate sources were found to be plant roots or litter, (Edwards.
1998)1. Organic soil S cannot be taken up by plants directly, but organic S compounds
are mineralized back to sulfate by soil microbes under favorable moisture, temperature,
and oxic conditions (Wang et al.. 2006). Organic S pools can be indirectly enhanced by
SOx deposition via microbial and plant incorporation, as indicated by measures of soil
[S] and S isotopes. Sulfate in the soil solution can remain in the soil S pool in its
inorganic form as it adsorbs to hydrated soil particles. Soil adsorption of sulfate generally
results in a smaller S pool than soil organic matter, but this pool can be directly enhanced
by SOx deposition and associated acidification. Acidification may increase the storage
capacity of soils for sulfate. In the case of nonspecific adsorption, increasing acidity and
more H+ held in colloidal suspension create more sulfate binding capacity in soil, while
ecosystem recovery towards a more neutral pH may cause desorption of sulfate and
higher sulfate concentrations in the soil solution. In the case of specific adsorption,
sulfate binds with metal atoms in mineral particles; sulfate specific adsorption to soil also
increases with lower pH. However, the rate at which sulfate bound to soil particles is
released from adsorption in the soil solution as the ecosystem recovers (see Appendix 4)
towards a more neutral pH will heavily depend upon soil factors such as clay mineralogy,
soil age and weathering, calcium availability, content of amorphous iron and aluminum
oxides, and glaciation history. Inorganic S storage within terrestrial soils can be directly
affected by SOx deposition, as indicated by measures of sulfate concentrations in soil
fractions and soil solution.
Sulfate in the soil solution may be transported out of the terrestrial ecosystem if it leaches
to surface water and into downstream wetland and aquatic ecosystems. Sulfate
concentrations in surface water are often measured as an indicator of S fluxes from small
terrestrial watersheds. In watersheds that have received historically high S deposition,
dissolved organic sulfur (DOS) may also comprise a substantial portion of the S-leaching
from terrestrial to aquatic watersheds. In Archer Creek in the Adirondacks, NY, DOS
averaged 21% of total dissolved sulfur in the creek (Kang et al.. 2014). Freshwater
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aquatic S cycling is not as well described in the scientific literature as is terrestrial soil S
cycling. The preponderance of evidence is concerned with the locations of redox S
reactions within these aquatic systems. Sulfate leached from terrestrial ecosystems, as
well as sulfate from direct deposition to surface water, is transformed in aquatic systems
under particular conditions. High oxygen levels in the water column or inundated
sediments of a wetland, lake, or stream favor S in oxidized form as sulfate. Sulfate in
anoxic (very low O2) waters or sediments is transformed into more reduced forms such as
sulfide or elemental S, which are less soluble and may precipitate into sediments.
Sulfur-reducing prokaryotes (SRPs) couple their metabolism to the transformation of
sulfate into these reduced forms. SRPs are obligate anaerobes but require sulfate for
metabolism, and as a result, they are often present and active at oxic-anoxic boundaries in
aquatic environments. In shallow lakes, this boundary (and the majority of sulfate
reduction) occurs at the water-sediment interface, in the top 1-2 cm of sediment (Rudd et
al.. 1986; Kelly and Rudd. 1984). In deeper lakes that experience seasonal stratification,
the water column may account for up to 15% of sulfate reduction (Ingvorscn and Brock.
1982). and the anoxic boundary may occur in the flocculant layer (suspended sediment);
for example, the boundary occurs 10 cm above the top of the sediment in Little Rock
Lake, WI (Sherman et al.. 1994). In wetlands, particularly in peat bogs that depend
entirely upon precipitation as their water supply, the anoxic-oxic boundary often
coincides with the water table and can be much more dynamic. Finally, recent research
suggests that periphyton occupying high-oxygen waters can create anoxic
microenvironments that promote the growth and activity of SRPs (see Appendix 12.3.3).
The location and activity of SRPs are important environmentally because some SRPs
couple sulfate reduction with Hg methylation (see Appendix 12.3.2). The transformation
of S between reduced and oxidized forms in aquatic and wetland ecosystems occurs in
particular locations within these ecosystems but can have broader impacts on chemistry
and biota.
Pools of S in freshwater systems (lakes, rivers, streams, and wetlands) and residence
times of S in ecosystem pools are addressed in only a few ecological studies. S cycling in
lakes removes sulfate from the water column and stores reduced S in sediments. Sulfate
in the water column diffuses into sediments where it binds to particulate matter or is
incorporated by microbes. Water column sulfate is reduced (by chemical or biological
processes) and released into the sediments as sulfide, which can bind to iron and
precipitate, or as elemental S, which can precipitate directly. An early study by Rudd et
al. (1986) developed a budget for reduction-oxidation cycling of S based on Adirondack
lakes, and estimated that 47% of water column sulfate was reduced and stored in
sediments. A broader study that included lakes in the Adirondacks, Ontario, the
Experimental Lakes Area, Northern Wisconsin, and Southern Norway estimated that
39-80% of annual sulfur load was retained in sediments based on mass balance, or
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11-65% based on measured sulfate reduction rates, across all lakes (Kelly et al.. 1987).
On average across lakes, one-third to one-half of annual lake sulfate load was transferred
to sediments, and variation around this mean correlated with water residence times of
lakes. Longer hydraulic residence times correlated with more sulfate removal. The
release of reduced S into the water column is responsible for some of the other
nonacidifying effects of S deposition, including internal eutrophication and protective
effects against certain heavy metals (see Appendix 12.2.4). The rest of the S originally
taken up by microbes as sulfate is incorporated into organic compounds. The growth and
activity of SRPs are responsible for Hg methylation in many sediments (see
Appendix 12.3). About one-third of the reduced inorganic S and 57% of the organic S are
buried in lake sediments annually, or about 47% of the total sulfate input. The remaining
reduced sulfur, 53% of the original sulfate input, is oxidized by chemical reaction with
water molecules (Bates et al.. 2002) and returns to the S cycle (Rudd et al.. 1986). These
studies suggest that changes in S inputs to lakes—for example, reductions in SOx
deposition—could rapidly reduce sulfate concentrations in the water column, assuming
other environmental factors remain steady (see below for climatic influences on S
cycling).
Wetland S budgets resemble lake budgets in S reduction and burial of reduced S in
sediments, but also may include plant uptake and S export if hydrology allows. In a study
of depressional freshwater wetlands in two lakes in Germany, the sulfate reduction rate
was 2.5 to 7.9 times higher annually in the high sulfate lake (22.7 mg sulfate/L) than in
the low sulfate lake (9.0 mg/L) (Kleeberg et al.. 2016). Sulfur budgets have also been
determined for the peat soils and freshwater marshes of the Florida Everglades. The EAA
has large impacts on S cycling in freshwater marshes in the surrounding designated
Water Conservation Areas (WCA), which serve as a water quality buffer around
Everglades National Park, a Class I area. In the EAA, the addition of elemental S as an
agricultural soil amendment increased sulfate in the soil from 13 to 19% of total soil S,
although at the end of the growing season sulfate was 1% of total S due to uptake by
growing sugarcane (Ye et al.. 2010) and runoff into surrounding canals and wetlands
(Bates et al.. 2002). In an early study quantifying S budgets in the Everglades WCA,
(Bates et al.. 2002) determined that most of the sulfate in the system has its source in S
agricultural amendments applied in the EAA. In these wetland systems, the pool of S in
organic matter is high, and additional sulfate could be rapidly incorporated into organic
matter or exported in surface water, in addition to the reduction or burial, which occurs in
other aquatic systems.
Weather and climatic factors can disrupt S storage in more stable pools within aquatic
and wetland systems. In smaller stratified lakes, sulfate transfer from the epilimnion into
lower layers occurs slowly by diffusion between the stable water layers, while in larger
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lakes, the layers are more prone to perturbation by wind and allow larger transfers of
sulfate to zones where SRPs are active (Ingvorsen and Brock. 1982). Sulfur cycling, like
most microbially driven processes, is seasonal; sulfate fluxes to the sediment are 10 times
higher in the summer than in winter (Sherman et al.. 1994). and the growing season
coincides with peak sulfate reduction in many freshwater systems (see
Appendix 12.3.3.3V In the Experimental Lakes Area (ELA) in Ontario, 10-39% of
reduced S ended the summer as sulfide dissolved in the hypolimnion, and this S was
reoxidized to sulfate when circulation occurred in the fall (Kelly et al.. 1982). In winter,
O2 can penetrate deeper into sediments or peat, reoxidizing the sulfide (Sherman et al..
1994). This reoxidation can also occur in water bodies where water levels are controlled
by humans, for example in reservoirs (Ecklev et al.. 2015). Extreme events like droughts
can also result in reoxidation of reduced S (Wasik et al.. 2015) and can contribute to Hg
methylation hotspots upon rewetting and restoration of anaerobic aquatic conditions.
Deposition and Sulfur Stores
In terrestrial ecosystems, atmospheric deposition of S at levels below amounts that trigger
acidification may still induce biochemical effects. These effects were described in the
2008 ISA and include nutrient enrichment of plants, increased S storage in soils, and
elevated sulfate concentrations in water leaching from soils. New information regarding
the effects of S deposition on terrestrial biogeochemical cycling is sparse, but can be
found in Appendix 4. Terrestrial plant productivity is commonly limited by N, so SOx
deposition at levels low enough to provide S enrichment without associated acidification
effects is unlikely to have significant effects upon terrestrial plant productivity. Instead,
the nonacidifying effects of SOx deposition are strongest in freshwater systems such as
wetlands and lakes, where hydrology controls storage and release of S.
Changes in water availability and water levels in aquatic systems alter the S cycle and
may release additional S to the aquatic environment from storage in sediment. During a
drought at Little Rock Lake in Wisconsin, lake sulfate concentrations increased
0.18 mg/L/yr, from 1.5 mg/L in 1998 to 2.9 mg/L in 2006 (Watras and Morrison. 2008).
During this time period, lake water volume decreased only 30%, which would not
account for the 93% increase in water sulfate concentration, suggesting that drought
released additional S stored in lake sediment or biomass into the water column. Watras
and Morrison (2008) calculated that an additional load of 5 kg S/ha/yr from internal
ecosystem sources of S would account for the increase of S in Little Rock Lake. In the S
addition experiment at the S6 bog in Marcell Experimental Forest, MN (see
Appendix 12.3.4.1). rewetting of the bog after a drought released a pulse of sulfate into
surface water in both control and experimental S addition plots. In plots to which
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32 kg S/ha/yr had been added, concentrations of sulfate in peat pore water after the
drought were 438% higher than in control plots ("Wasik et al.. 2012). indicating that S
addition decreases the retention of S in drought-stressed wetlands (Table 12-1). This is
particularly important because intensification of drought frequency and severity is one of
the possible effects of climate change, and suggests that climate change and SOx
deposition may synergistically contribute to increased sulfate concentrations in surface
water and associated water quality impairment.
Table 12-1 New study on sulfur (S) deposition effects on sulfur (S) cycling.
Type of
Ecosystem
Additions or Load
(kg S/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
Bog	Ambient S deposition:
5.5 kg/ha/yr (mid 2000s) as
quantified at NADP site
MN16
S addition: 32 kg S/ha/yr
dissolved in pond water and
delivered by sprinklers
(mimics 4x the 1990s
deposition rate)
Following 2007
drought, pore water
SO42" concentrations
were 438% higher in
experimental
treatment.
S6 peatland,
bog section,
Marcell
Experimental
Forest, MN
Wasik et al.
Pore water,
peat, and Culex (2012)
spp.(mosquito)
larvae
ha = hectare; kg = kilogram; NADP = National Atmospheric Deposition Program; S = sulfur; S042 = sulfate; yr = year.
12.2.3. Sulfide Toxicity
Wetland ecosystems are characterized by occasional or total soil inundation and typically
include water and sediment zones with low oxygen levels. Under low-oxygen conditions,
sulfate can be used as an electron acceptor by a number of microorganisms and is quickly
converted to sulfide during decomposition of organic C. Sulfide inhibits plant root
nutrient uptake at high levels [>34.1 mg/L or >1 mM; (Koch etal.. 1990)1. and sulfide
toxicity due to increased sulfate loads was documented by the 2008 ISA and references
therein for wetland plants, including Carex spp., Juncus acutiflorus (not native to or
documented in North America), Galium palustre (endangered in Ohio, special concern in
Tennessee), other species within Gramineae (the grass family), as well as the aquatic
macrophytes Stratiotes aloides (listed as an Class C noxious weed in Alabama, Class 1
prohibited aquatic plant in Florida) and Elodea nuttallii (threatened in Kentucky, special
concern in Tennessee). As reviewed in the 2008 ISA, Smolders et al. (2003) set a
threshold value of <48 mg SO42 /L in surface water to protect the sensitive aquatic
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species Stratiotes aloides and Potamogeton acutifolius (not native to CONUS), as well as
to protect Potamogeton zosteriformis and Utricularia vulgaris, which are both native and
widely distributed in CONUS. This threshold would protect against sulfide phytotoxicity
and internal eutrophication and was based on correlations between surface water
concentrations and plant presence/absence in Dutch fens (Smolders et al.. 2003). A recent
survey of water quality and plant presence/absence data from four water management
districts in the Netherlands confirmed that the probability of encountering the S-sensitive
plants S. aloides and six S-sensitive Potamogeton (pondweed) species declined
significantly at sulfate concentrations greater than 50 mg/L (Vcrmaat et al.. 2016).
More recent research confirms sulfide toxicity in wetland habitats and suggests that
sulfide toxicity can determine plant community composition in freshwater wetlands. A
review of the sulfide toxicity literature by Lamers et al. (2013) found sulfide toxicity
occurred between 0.3-29.5 mg S27L (originally reported as 10-920 |imol L ') in
freshwater wetland emergent plants and aquatic submerged macrophytes native to North
America (see Table 12-2 for species-specific phytotoxicity levels). A recent study
sampled pore water chemistry and plant community composition in Junius Pond Fen,
Seneca County, NY, and found that sulfide concentrations (range: not detectable to
5.73 mg/L or 168 |iM H2S) had negative effects on the plant community. Sulfide
concentration was negatively correlated with total plant cover and had strong effects on
the cover of the dominant plants in the fen (Simkin et al.. 2013). In models, high sulfide
concentrations decreased cover of the moss species Campylium stellatum, decreased
cover of the monocot and USDA-classified (USDA. 2015b) wetland obligate Eleocharis
rostellata (endangered in Florida and Pennsylvania; threatened in Illinois, Minnesota, and
Wisconsin; sensitive or special concern in Rhode Island and Washington), and decreased
cover of obligate monocot Cladium mariscoides (endangered in Pennsylvania). Sulfide
concentrations also correlated negatively with dicot diversity (Simkin et al.. 2013). This
study shows that sulfide toxicity effects on sensitive plant species can cascade to affect
plant community composition in fens. There are also sensitive species in other freshwater
wetlands. A greenhouse experiment with Taxodium distichum (baldcypress) seedlings
collected from a North Carolina riparian field showed negative effects of sulfate addition
(sulfate concentrations: 48 mg/L in drought treatment, 129 mg/L in flooded treatment) on
seedling height, which the researchers attributed to sulfide toxicity (Powell et al.. 2016).
Baldcypress is a foundation species in southeastern freshwater swamps, which if located
in estuaries, may also experience elevated sulfate, sulfide, and salinity as a result of
saltwater intrusion. SOx deposition may exacerbate sulfide toxicity in baldcypress
swamps already stressed by sea level rise.
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Table 12-2 Quantitative effects of sulfide on wetland and aquatic plant species.
Species
Effects of Sulfide
Sulfide
(mg S27L)
Listing
Reference
Calla palustris
Decreased aboveground
productivity
4.8
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Caltha palustris
Decreased aboveground
productivity
5.5
NRCS wetland
obligate
Van der Welle et al.,
2007b in Lamers et
al. (2013)
Carex nigra
Decreased aboveground
productivity
0.3-0.6
NRCS wetland
facultative
Lamers et al., 1998 in
Lamers et al. (2013)
Ceratophyllum
demersum
Decreased aboveground
productivity
16
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Cladium jamaicense
Decreased leaf elongation
rate
Decreased net
photosynthetic rate
Aboveground die-off and root
(and rhizome) die-off
7.0
22.1
29.5
NRCS wetland
obligate
Li et al, 2009 in
Lamers et al. (2013)
Elodea nuttallii
Decreased aboveground
productivity
3.2
NRCS wetland
obligate
Van der Welle et al.,
2007a in Lamers et
al. (2013)
Eiodea nuttallii
Decreased aboveground
productivity
4.8-16
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Equisetum fluviatile
Decreased aboveground
productivity
1.6-16
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Juncus
alpinoarticulatus
seedlings
Decreased aboveground
productivity
1.0-1.6
NRCS wetland
obligate
Grootjans et al., 1997
in Lamers et al.
(2013)
Juncus effusus
Decreased aboveground
productivity
16
NRCS wetland
obligate or facultative
Geurts et al., 2009 in
Lamers et al. (2013)
Menyanthes trifoliata
Decreased aboveground
productivity
(unfertilized/fertilized)
4.8/>4.8
NRCS wetland
obligate
Geurts et al., 2009 in
(Lamers et al., 2013)
Menyanthes trifoliata
Decreased aboveground
productivity
>7.5
NRCS wetland
obligate
Armstrong and
Boatman, 1967 in
Lamers et al. (2013)
Nitella flexilis
Decreased aboveground
productivity
1.6

Van der Welle et al.,
2006 in (Lamers et
al., 2013)
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Table 12-2 (Continued): Quantitative effects of sulfide on wetland and aquatic
plant species.
Species
Effects of Sulfide
Sulfide
(mg S2"/L)
Listing
Reference
Panicum hemitomon
Decreased aboveground
productivity, root (and
rhizome) die-off
20.2
NRCS wetland
obligate
Koch and
Mendelssohn, 1989
in Lamers et al.
(2013)
Panicum hemitomon
Decreased aboveground
productivity, decreased root
ADH activity, decreased
nutrient uptake
32.1
NRCS wetland
obligate
Koch et al., 1990 in
Lamers et al. (2013)
Potamogeton
compressus
Decreased aboveground
productivity
4.8-16
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Sphagnum
cuspidatum
Aboveground die-off
1.9

Lamers et al., 1999 in
Lamers et al. (2013)
Thelypteris palustris
Decreased aboveground
productivity
4.8
NRCS wetland
obligate
Geurts et al., 2009 in
Lamers et al. (2013)
Typha domingensis
Decreased leaf elongation
rate, decreased net
photosynthetic rate,
aboveground die-off, root
(and rhizome) die-off
29.5
NRCS wetland
obligate
Li et al., 2009 in
Lamers et al. (2013)
Source: all columns except "Listing" are from Lamers et al. (2013).
Sulfide toxicity has been proposed as an important causal factor in the expansion of
Typha domingensis (southern cattail) in the Florida Everglades. Southern cattail is
displacing the historically dominant sawgrass, Cladium jamaicense, which is the
keystone species in the Everglades sawgrass prairies and which has minor agricultural
importance as a food source for small mammals and water birds (reported in. Everitt et
al.. 1999). In a hydroponic greenhouse study of both species, cattails had a higher
tolerance for sulfide in all plant metrics measured. Sulfide concentrations in growth
media were 0, 7.5, 15.7, 23.5, 31.4 mg/L sulfide (reported as 0, 0.22, 0.46, 0.69, or
0.92 mM). At 7.5 mg/L sulfide, sawgrass leaf elongation rates decreased 49%; at
15.7 mg/L sulfide, net photosynthetic rate decreased 34%; and at 23.5 mg/L, sawgrass
biomass in different compartments decreased 29-41% (Li et al.. 2009). Cattails
experienced sulfide toxicity effects on leaf elongation rates at 23.5 mg/L sulfide (38%
decrease) and on photosynthesis at 31.4 mg/L sulfide (22% decrease), with no effects of
sulfide upon biomass at any concentration (Li et al.. 2009). Together, these results
suggest that sulfide toxicity is shifting the competitive balance between these species
away from the dominant species in the Everglades. However, studies that assessed cattail
and sawgrass growth in mesocosms placed across the naturally occurring sulfate-sulfide
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gradient in the Everglades (DeBusk et al.. 2015) or in mesocosms with added sulfate (Li
et al.. 2009) have not found effects of sulfide toxicity upon sawgrass. Under controlled
conditions, sulfide toxicity favors cattails over keystone Everglades species sawgrass, but
evidence from field studies in the Everglades ecosystem suggests that geological or
ecological factors not reproduced in the greenhouse may mitigate sulfide toxicity.
Recent work by the Minnesota Pollution Control Agency (MNPCA) proposes to
incorporate geological and biological factors into water quality regulation in order to
prevent sulfide toxicity to an economically important species. Wild rice (Zizania
palustris) has been protected by the state of Minnesota as ecologically important food for
waterfowl and an economically important foraged crop (MPCA. 2015a). and the Ojibwe,
Menominee, and Dakota peoples value the plant as a cultural and economic resource
(Pastor etal.. 2017). The state first set a standard to protect wild rice in 1973, when the
water quality standard was set at 10 mg sulfate/L. In 2011, the state began collecting field
samples, conducting experiments in hydroponics and mesocosms, and modeling
environmental and experimental data in order to revise the standard. The previous
standard of 10 mg sulfate/L would only protect 58% of current wild rice sites from
sulfide toxicity (MPCA. 2015a). In March 2015, the Minnesota Pollution Control Agency
(MPCA) released its recommendations (MPCA. 2015a). Based on the new analyses, the
MPCA set a sulfide concentration threshold of 0.165 mg sulfide/L in sediment pore water
to protect wild rice, based on 2011-2013 monitoring of 112 Minnesota water bodies in
which this level of sulfide corresponded to a 10% decrease in the probability of the
species being present (MPCA. 2015b). Sulfide is the end product of microbial sulfate
reduction, which is heavily dependent upon other environmental factors, particularly
DOC, and sulfide phytotoxicity depends upon its residence time as a free ion in water,
which depends upon iron concentrations (see Figure 12-3). As a result, the MPCA did not
set a single value of sulfate as a standard to protect wild rice.
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High iron
in sedime
High organic
carbon levels in
sediment
*	bacteria
| iron	iroiT)( iron^)
Iron in sediment binds to sulfide
and neutralizes it, making it
nontoxic to wild rioe.
Oganic carbon in sediment is
food forthe bacteria, causing
more sulfide to be produced.
Source: from MPCA (2015b).
Figure 12-3 Schematic from Minnesota Pollution Control Agency that
illustrates the mitigating effect of iron on the toxicity of sulfide
and the stimulatory effect that organic carbon has on sulfide
production.
Instead, MPCA established a formula that would allow protective sulfate thresholds to be
calculated for all wild rice-producing water bodies based on sediment organic carbon and
sediment iron concentrations in each water body [Source: (MPCA. 2015b)l:
Sulfate = 0.0000136 x Organic Carbon—1.410 x Ironl.956
Equation 12-1
The range in sulfate values that protect wild rice from sulfide toxicity in Minnesota water
bodies is from 0.4 to >200 mg SO-r /L. This proposed standard will require large-scale
sediment sampling to determine sediment pore water DOC and Fe concentrations in
Minnesota water bodies, and is still under review. These regulatory efforts determine
protective surface water sulfate concentrations on the basis of a sulfide phytotoxicity
threshold protective of wild rice. In hydroponic experiments with Z. palustris, sulfate and
sulfide exposures did not affect seed germination, but sulfide concentrations of 0.32 mg/L
or higher reduced growth rates 88% compared with controls (Pastor et al.. 2017). The
same research group also conducted growth trials in outdoor tank mesocosms in which
surface water sulfate concentrations, and thus sediment sulfide concentrations, could be
manipulated. In the mesocosms, 50 mg S0.f7L did not affect seedlings, but elevated
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sulfate and sulfide at 100 mg SO42 /L or higher decreased seed mass, seed viability,
seedling emergence rates, and seedling survival rates (Pastor et al.. 2017).
Sulfide phytotoxicity as a result of SOx deposition alters freshwater wetlands and lakes,
but is unlikely to affect coastal wetlands. Seawater contains high concentrations of sulfate
[1 ppt seawater contains 750 |iM or 72 mg/L sulfate according to (Hackney and Avery.
2015)1. and as a recent study in the Cape Fear Estuary, North Carolina, demonstrates,
seawater intrusion and sulfate-reducing conditions structure the distribution of tidal
marsh and tidal swamp zones in the estuary. Specifically, tidal marsh plants are salt and
sulfide tolerant, and occur in the estuary in areas where sulfate reduction dominates
microbial mineralization (Hackney and Avery. 2015). Tidal swamp plants (such as
baldcypress, see above) have been shown to be sensitive to and negatively affected by
saltwater intrusion over the decade-long course of the study. This study suggests a
biologically plausible link between sulfide increases caused by SOx deposition and tidal
marsh intrusion upon tidal swamp zones. However, the authors did not consider sulfate as
originating from any source other than seawater, so the effect of salinity could not be
separated from the effect of sulfate reduction and resultant sulfide phytotoxicity. There is
no evidence of SOx deposition contributing to sulfide phytotoxicity in wetland and
aquatic ecosystems with significant seawater contributions.
Sulfide toxicity has not been documented for freshwater animal species, since the
reducing conditions that produce sulfide also cause hypoxia. However, animal species do
vary in their tolerance of sulfate concentrations in water. Carlisle et al. (2007) and
Meador and Carlisle (2007) used data from the U.S. Geological Survey's 1993-2003
National Water-Quality Assessment (NAWQA) to generate indicator values for widely
distributed North American macroinvertebrates and fish. Each indicator value is the
average value for a water quality parameter at which a taxon is detected in NAWQA
samples. A follow-up analysis of macroinvertebrates and stream quality in the eastern
and midwestern U.S. designated streams with more than 20% loss of expected taxa for
the region as degraded habitat (Carlisle and Meador. 2007). This dataset has identified
animals sensitive and tolerant to sulfate and other water quality parameters, and allows
the evaluation of sulfate impacts on animal species at concentrations lower than those
that cause acidification.
Internal Eutrophication
Internal eutrophication occurs in saturated soils in freshwater wetlands and lakes when
sulfate addition results in the release of bioavailable phosphate from soil forms that are
not bioavailable. Additional sulfate in surface waters can increase P mobility and cause
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internal eutrophication of the recipient ecosystem; farmers growing sugarcane in the
oxidized peatland soils of the Florida EAA use S as a soil amendment for this reason (Ye
et al.. 2010).
Wetland and aquatic systems often retain phosphate as FePC>4. When sulfate is reduced,
the resultant sulfide binds with iron to create FeS, which precipitates out of solution,
releasing P into surface waters (Figure 12-4). This internal eutrophication can alter plant
community composition (U.S. EPA. 2008a). or can speed the accretion of organic matter
and infilling of the wetland, reducing wetland habitat (Kleeberg et al.. 2016). In
mesocosms resembling shallow lakes planted with wild rice, experimentally elevated
surface water sulfate concentrations raised phosphate concentrations in the water column
(Pastor etal.. 2017). Water and sediment mesocosms of samples collected from Lake
Moshui in China showed that sodium sulfate addition to raise sulfate concentrations to
500 mg/L also raised peak [P] by 180% in the water column, and by 210% in the
sediment pore water, above unamended control mesocosms, and the peak [P] with sulfate
addition occurred a week later than in control mesocosms (Yu et al.. 2015). Limitation of
microbial sulfate reduction achieved through experimental NOs enrichment of the
hypolimnion in Onondaga Lake, NY, decreased soluble reactive phosphorus in the
hypolimnion by 95% (Matthews et al.. 2013). Sulfate addition can alter P dynamics and
cause eutrophication in freshwater systems.
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Fe(ll) = iron (II); Fe(lll) = iron (III); FeSx = iron-sulfide complexes; P043 = phosphate.
Open boxes indicate pore water components, and shaded boxes indicate solid-phase chemistry. Reduction reactions are indicated
by solid arrows, oxidation reactions are indicated by dotted lines, and phase changes (precipitation or dissolution) are indicated by
dashed lines.
Source: from Simkin et al. (2013).
Figure 12-4 Mechanisms of linked sulfur, iron, and phosphorus cycling in
wetland waters and soils.
Surface water sulfate may decrease heavy metal bioaccumulation in plants, but the
implications of this process for impacts of SOx deposition may be limited. Sulfate
addition can decrease the accumulation of toxic elements in aquatic plants. In soil, S
addition decreased dissolved arsenic (As) concentrations in soil solution, but increased
As(III):As(V), although the researchers suggested that the formation of As-Fe-sulfide
minerals prevented plant As uptake (Jia et al.. 2015). In microbially spiked soil
mesocosms, the addition of SRB had a stronger stimulatory effect on As(III) As(V) than
did S addition. In rice mesocosms, S addition decreased As concentrations in rice roots
and in the iron plaques on roots (Jia et al.. 2015). In laboratory mesocosms, accumulation
of selenium (Se) in aquatic plants was a function of both sulfate and Se concentrations in
water, but increasing sulfate concentrations decreased Se concentrations in both Lemnct
minor (aquatic macrophyte) and Pseiidokrichneriella subcapitata [unicellular green alga
(Lo et al.. 2015)1. Sulfate in surface waters may slow bioaccumulation of heavy metals in
biota by forming relatively insoluble S-metal complexes and preventing metal uptake by
aquatic plants. However, these findings may have limited applicability to SOx deposition.
If there is a common source of both SOx and heavy metal deposition to an aquatic
environment, it is unlikely that deposition will correlate with decreasing metal
bioavailability. Also, mercury is a focus of research into heavy metal environmental
transformations and bioavailability, and SOx deposition is linked instead to increased
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bioaccumulation of Hg (see Appendix 12.3 to Appendix 12.6). There is no research on
how SOx deposition affects Se or As accumulation in plants growing under field
conditions.
Sulfur Effects on Methane Emissions
The 2008 ISA documented the suppression of methane emissions from aquatic and
wetland ecosystems in response to increases in sulfate concentrations. When sulfate is
added to freshwater systems, SRPs outcompete methanogens, and methane emissions are
suppressed. The activity and environmental preferences of SRPs are key to understanding
the effects of SOx deposition upon methane and MeHg production. Under anaerobic
conditions, SRPs and methanogens compete for organic C as a metabolic substrate, and
the reactions of sulfate reduction coupled with organic C oxidation are
thermodynamically favored over the reactions of methanogenesis (Paulo et al.. 2015). In
many anaerobic environments, sulfate is in short supply, and methanogens are more
active than SRPs in mineralizing organic C. Sulfate addition may swing the competitive
balance back towards SRPs, decreasing methane production in saturated soils and surface
waters. However, studies summarized in other parts of this appendix show that SRPs and
methanogens can form syntrophies (see Appendix 12.3.2) or coexist and rapidly shift
dominance within a single wetland (see Appendix 12.3.2 and Appendix 12.3.3.1).
The relative primacy of methanogens versus SRB in oxidizing carbon is one of the
factors which distinguishes freshwater from coastal wetlands and water bodies. Older
literature suggests that when SOx deposition is not a factor, methanogens dominate
anaerobic decomposition in freshwater systems (including wetlands), with
methanogenesis responsible for 72-82% of organic C mineralization at the ELA, Ontario
(Kelly and Rudd. 1984). and methanogenesis responsible for 4 times the electron flow
and organic C mineralization of sulfate reduction in Lake Mendota, WI (Ingvorsen and
Brock. 1982). In estuarine and marine ecosystems, on the other hand, SRPs are the
dominant anaerobic decomposers. In the saltwater marshes at Kirkpatrick Marsh, MD,
methanogenesis is <10% of C mineralization, and sulfate reduction is responsible for
45-85% of total C mineralization in the summer, and 5-25% of C mineralization in the
fall (Mitchell and Gilmour. 2008). Older research suggests that in pristine ecosystems,
methanogens dominate anaerobic C mineralization in freshwater systems.
New research on microbial communities in saturated sediments contributes to the
biological plausibility of the reduction in methanogenesis caused by SOx deposition
established by the 2008 ISA . Adding sulfate to freshwater systems can depress
methanogenesis because anaerobic C mineralization by SRPs is energetically favored
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when sulfate and acetate are available at the sediment surface (Urban etal.. 1994). Newer
work also supports this finding (see Table 12-3). Twitchell Island is a restored wetland in
the San Joaquin delta, CA, where researchers sampled sediments and water along a
gradient of decreasing riverine S load (14-8 mg sulfate/L in February) from the wetland
inlet to a transitional site, and to an interior marsh site. In summer, sulfate was 60% lower
and methane emissions were 110% higher at the interior site than at the inlet, and 16S
rRNA sequencing confirmed that relative SRP abundance was negatively correlated with
methanogen abundance (He et al.. 2015). In the Everglades WCA, FL, similar gradients
of S loading occur from agricultural runoff in the freshwater marsh. Sampling at sites F4
(7.1 mg/L or 74 (j,M sulfate in pore water), U3 (3.7 mg/L or 39 (j,M sulfate), and W3
(<0.4 mg/L or <4 (j,M sulfate) found effects of sulfate upon SRB (sulfur-reducing archaea
were not sampled) and methanogens (Bae et al.. 2015). SRB abundance increased with
increasing sulfate pore water concentrations across the three sites, and was 6.9 times
higher at U3 and 16.8 times higher at F4 than at W3. Sulfate also affected the competitive
interaction between SRB and methanogens, which had equal abundances at the most
pristine W3 site. At the sites with higher S loads, SRB were 60-80% more abundant than
methanogens. The relative abundance of methanogens was determined by copy number
of the mcrA gene in environmental samples, which correlated positively with methane
production rates and mRNA (Bae et al.. 2015). so sulfate loading alterations to the
microbial community in the Everglades WCA may also be lowering methane emissions.
Sampling from natural microbial communities confirms that sulfate favors SRB over
methanogens.
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Table 12-3 New studies on nonacidifying sulfur effects on methane emissions.
Type of
Ecosystem
Additions or
Load (kg
S/ha/yr)
Biological and Chemical
Effects
Study Site Study Species
Reference
Restored River water is
peatland the source of
island in the S load,
river	more interior
sites have lower
S load, as
reflected by
February
sediment
surface
measurements:
Inlet [SO42"]
14 mg/L;
Transitional
[S0421
11.5 mg/L;
Interior [SO42"]
8 mg/L
SRP abundance decreases Twitchell
with increasing
methanogen abundance
(r= - 0.67).
Denitrifier abundance
decreases with increasing
methanogen abundance
(r= -0.64).
In August, sulfate
decreases 60% from inlet
to interior, and methane
emissions increases 110%.
SRP relative abundance
increases with sulfate
concentration (r= 0.96).
Island, San
Joaquin
delta, CA
Island covered He et al. (2015)
with Typhus spp.
and
Schoenoplectus
acutus, sampled
at inlet,
transitional, and
interior marsh
sites.
Microbial
community
quantified by 16S
rRNA sequencing
of peat sample.
Freshwater
marsh
W3 pore water
[SO42"] <4 |jM
U3 pore water
[SO42"] is 39 |jM
F4 pore water
[SO42"] is 74 |jM
Deposition = not
measured
SRB abundance is 6.9x
higher in U3 than W3 and
16.8x higher in F4.
In W3, abundance of SRB
and methanogens are not
significantly different. In
U3, SRB abundance is
80% higher than
methanogen abundance. In
F4, SRB abundance is
60% higher than
methanogens.
mcrA copy number
correlates positively with
mRNA and methane
production rates.
Acetotrophic methanogens
are dominant at W3, while
at high S U3 and F4 sites,
hydrogenotrophic
methanogens are
dominant.
F4, U3, and
W3 sites in
Water
Conservation
Area,
Everglades,
FL
Methanogens as
quantified by mcrA
copies.
Sulfur-reducing
bacteria as
quantified by dsrB
copies.
Bae et al. (2015)
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Table 12-3 (Continued): New studies on nonacidifying sulfur effects on methane
emissions.
Type of
Ecosystem
Additions or
Load (kg
S/ha/yr)
Biological and Chemical
Effects
Study Site Study Species
Reference
Coastal salt
marsh
Total load = not
measured
Deposition = not
measured
No significant correlation
between potential methane
emissions and pore water
S042", NO3-, and Fe3+.
Methanogen abundance is
84% lower in S. alterniflora
than in P. australis marsh.
SRP abundance is 73%
lower in S. alterniflora than
in P. australis marsh.
Methanogen abundance
(M, as 1,000 gene copies/g
soil) increases with pore
water NO3" concentrations
(Nbad, as |jM NO3"):
M = 0.3468 + 0.8149 x
Nload
Shanyutan,
Min River
Estuary,
China
Zones of Chinese
native Phragmites
australis, Chinese
invasive
(American native)
Spartina
alterniflora, and
Cyperus
malaccensis
(subtropical
species not
present in U.S.)
Methanogenic
archaea quantified
using 16S rRNA
SRPs are
quantified using
dsrA gene
Tona etal. (2015)
16S rRNA = 16S ribosomal ribonucleic acid; Fe3+ = iron; g = gram; ha = hectare; kg = kilogram; L = liter; |jM = micromolar;
mg = milligram; N03" = nitrate; S = sulfur; S042" = sulfate; SRPs = sulfur-reducing prokaryotes; yr = year.
A laboratory study with controlled redox conditions and microbial communities suggests
that even when SRB and methanogens are not in direct competition for anaerobic C,
methanogen activity will increase when sulfate concentrations decrease. In an incubation
of water and sediment collected from oligotrophic Lake Stechlin, Germany, researchers
were able to create fluctuating redox conditions while continuously sampling the water
column. When the system was reduced, expression of the dsrB gene (present in SRB)
increased, and sulfide concentrations increased, indicating that SRB abundance and
activity increased in response to anoxic conditions in the water column (Frindte et al..
2015). Methane also increased in the water column, indicating no competition between
SRB and methanogens in this system. Over time, dsrB diversity decreased, and
expression ofpmoA (present in methanogens) increased (Frindte et al.. 2015). indicating
that although there was no competition between SRB and methanogens, methanogens
were able to persist after SRB had exhausted the available sulfate and diminished in
abundance. Active methanogenesis occurs when SRB activity is constrained by low
sulfate availability. Reductions in SOx deposition could increase methanogen activity
typical of wetland and aquatic sediments.
As with sulfide phytotoxicity, impacts of SOx deposition upon methanogenesis may be
present in freshwater but not marine ecosystems, as the high endogenous concentration of
sulfate in seawater is likely to overwhelm the impact of anthropogenic S. In marine
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sediments, sulfate reduction accounts for 50-100% of anaerobic carbon mineralization
(Urban ct al.. 1994); methanogenesis often persists at depths in organic sediments below
the depth to which sulfate can diffuse. In estuarine wetlands in the Cape Fear River
system in North Carolina, Hackney and Avery (2015) stated that in vertical soil profiles,
a concentration of 28 mg/L (300 (j,M) sulfate was the threshold where S reduction
switched to methanogenesis as the dominant form of C mineralization. This study also
determined that in estuaries, the transition from methanogenic to S-reducing conditions
occurs when more than 25% of tidal flooding carries more than 1 ppt of seawater
(Hackney and Avery. 2015). In a study of the Min River Estuary in China, there was no
relationship between methane emissions and pore water sulfate, but dominant plant
species did affect microbial communities, with lower relative abundance of both
methanogens (84% decrease) and SRPs (73% decrease) in Spartina alterniflora
sediments than in Phragmites australis-associated sediments (Tong et al.. 2015). Both of
these species are dominant plants in North American eastern coastal marshes, and this
result suggests plant community may be more important than sulfate deposition in
determining methane emissions in coastal marshes. SOx deposition is unlikely to affect
methane emissions from coastal marshes and water bodies. Sulfate addition may also
result in lower methane emissions by stimulating microbial methane oxidation. In marine
sediments, this occurs in syntrophic associations (associations among heterotrophs
dependent on the metabolic byproducts of other microbes) between anaerobic
methanotrophic archaea and SRB, which pair anaerobic oxidation of methane with sulfate
reduction, or in a clade (monophyletic group) of anaerobic methanotrophs capable of
both sulfate reduction and methanotrophy (Jove. 2012). However, these processes have
been observed only in marine deep-sea sediments, and it is not yet clear how widely
distributed or important they are in other systems. In peat bog mesocosms, by contrast, S
addition of 96 mg/L decreased microbial biomass 23-57% and decreased methane
oxidation to almost zero (Lozanovska et al.. 2016).
Interactions between Sulfur Deposition and Mercury
The 2008 ISA found the evidence was sufficient to infer a causal relationship between S
deposition and increased methylation of Hg, in aquatic environments where the value of
other factors was within adequate range for methylation. Recent research, as well as key
early papers on the microbial methylation of mercury, is summarized in this section. New
evidence is consistent with the 2008 ISA in identifying sulfur-reducing bacteria as a
biotic link between increased SOx deposition and increased MeHg concentrations in the
environment and in biota. New developments since 2008 include identification of both
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the microbial genes linked to mercury methylation and the methylation capability in
certain archaeal strains (Appendix 12.3.1 and Appendix 12.3.2V
Ecosystem areas with high MeHg fractions (as a fraction of total Hg) are considered to
indicate active microbial methylation, as well as places where primary producers and
animals are at an elevated risk of MeHg accumulation. An active area of research is in
identifying areas or zones of ecosystems where %MeHg is elevated . Background
information on deposition and biogeochemical cycling of Hg, with particular emphasis on
identifying hotspots of mercury methylation, is presented in Supplemental Material.
Mercury methylation is a microbial process carried out by a subset of prokaryotes that are
active under reducing environmental conditions (anoxic water or sediments) while
requiring oxidized sulfur and organic C as substrates (see Figure 12-5). As a result, Hg
methylation rates are heterogeneous across time and the landscape. Transformation by
bacteria and archaea of inorganic Hg to organic MeHg occurs in wetland soils, in
Sphagnum mats of wetlands, and at the oxic-anoxic boundary in lakes, estuaries, and the
Arctic water column. Hg methylation by SRB embedded in periphyton is a new topic that
was not covered in the 2008 ISA (Appendix 12.3.2V Evidence of MeHg production in
periphyton has important implications for aquatic MeHg concentrations, because
periphyton may extend throughout water columns and are more connected to aquatic and
terrestrial food chains than are Hg methylation hotspots in sediment or peat. Regardless
of their location within the ecosystem, microbial Hg methylators have specific
environmental requirements which affect their rates of Hg methylation
(Appendix 12.3.3). There is new evidence from sulfur addition studies (Appendix 12.3.4)
and from studies of ambient conditions in North American ecosystems (Appendix 12.3.5)
to show that S deposition and ambient sulfate concentrations increase mercury
methylation under certain conditions. New evidence is consistent and coherent with the
conclusions of the 2008 ISA and the body of evidence is sufficient to infer a causal
relationship between S deposition and the alteration of Hg methylation in surface
water, sediment, and soils in wetland and freshwater ecosystems.
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Sulfate-reducing prokaryote
Sulfate | Ł
Dissolved
Organic
Carbon
i
increases
availability
~r
Inorganic
Mercury
dissimilatory
sulfur reduction
I
metabolism
of carbon
methylation
by hgcAB
proteins
Sulfide
Increased
SRP Activity
or Abundance
Methyl
Mercury
Figure 12-5.
Sulfate, dissolved organic carbon, and inorganic mercury are all
important determinants of the rate and amount of methyl mercury
produced by sulfate-reducing prokaryotes.
12.3.1. Mercury Cycle and the Importance of Methylation
Mercury is present in ecosystems as elemental Hg", as Hg2+ complexed to metals or
organic compounds, or as MeHg. Since the Industrial Revolution, Hg emissions have
increased Hg concentrations in water and soils above preindustrial levels. Although in
recent decades U.S. emissions have declined [79% decrease in Hg emission between
1990 and 2011 as reported by the NEI (U.S. EPA. 2016a)l. Hg concentrations remain
elevated in air, water, soil, and biota above inferred preindustrial values. Environmental
mercury sources and transformations are illustrated in Figure 12-6; important mercury
sources to ecosystems that this appendix will consider include geological formations, wet
and dry deposition of Hg, and legacy Hg stored in soils from historical deposition or
historical (i.e., no longer active) industrial use. Hg released from active industrial
sources, active mines, or historical mine sites to soil or water are not considered.
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Source: from Biaham et al. (2017).
Figure 12-6 Cycling of mercury (Hg) and methylmercury (MeHg) in
ecosystems.
Hg deposition is generally more bioavailable and readily methylated than is Hg already
stored in sediments or organic matter within the ecosystem, and sulfate stimulation of
methylation favors Hg newly added to the system over Hg already in the ecosystem
(Bigham et al.. 2017).
hi terrestrial soils, Hg forms complexes with organic matter (Poulin et al.. 2016). New
research suggests that disturbances that accelerate carbon fluxes in terrestrial systems
also accelerate Hg release from soils and sometimes stimulate methylation. In a study of
Picea cibies (Norway spruce) stands in Sweden, clear-cutting raised the soil water table
and created conditions favorable for Hg methylation. Clear-cutting did not alter Hg
concentrations in the organic soil (O) layer, but it did significantly increase MeHg
concentrations and %MeHg in the O layer 5-7 times over control conditions (Kronberg
et al.. 2016). hi a similar study of clear-cutting plantations planted on drained peat soils in
Finland, harvesting increased DOC and Hg in drainage waters (Ukonmaanaho et al..
2016). Peat SRB abundance correlated with MeHg concentrations and varied by geology,
with higher MeHg concentrations in watersheds with underlying schist formations (high
Hg content). In Oak Ridge, TN, riparian forest soils along East Fork Poplar Creek are
contaminated by historical industrial mercury releases and contain 2-3 times the Hg in
uncontaminated soils. Soil microcosms showed that flooding of these soils moved
mercury bound to soil organic matter into the soil solution, as aqueous Hg" or as HgS
precipitates, the latter produced by microbial sulfate reduction within the soils (Poulin et
al.. 2016). Both Hg" and nanoparticulate HgS are bioavailable to microbial methylators,
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and flooding resulted in a pulse of both aromatic DOC and MeHg in the riparian soil pore
waters (Poulin et al.. 2016).
Recent studies using stable isotopes of mercury reviewed by Paraniape et al. (2017)
reported a range of 0.1-15% MeHg in North American and Swedish wetlands, and
0.4-4.6% MeHg in lake sediments. There is new evidence that MeHg concentration in
Adirondack surface waters is positively correlated with Hg° deposition rate (Gcrson and
Driscoll. 2016). Specifically, Hg in terrestrial leaf litter from Arbutus Lake watershed
was measured 6 times over 2004-2014, and declined 40% in concert with a 25% decline
in Hg° atmospheric concentrations. Over the same time period, MeHg concentrations in
Arbutus Lake declined (p < 0.03), although there was no quantification of the trend
reported. Gerson and Driscoll (2016) suggested that these results show that decreases in
Hg° concentrations will limit the supply of bioavailable Hg to mercury methylators and
decrease MeHg in aquatic ecosystems.
Wetlands tend to have higher MeHg fractions than other water bodies at similar locations
and order within watersheds. In the Adirondack Mountains, NY, sampling of soils and
substrates (upland, riparian, open water, peat bog) at different points in the Sunday Lake
watershed found that the highest MeHg fraction was in Sphagnum mats (Yu et al.. 2010).
In the Archer Creek watershed, also in the Adirondacks, MeHg concentrations in
wetland-draining streams were elevated compared to MeHg in upland streams
(Selvendiran et al.. 2008a). At a larger scale survey of 44 Adirondack lakes sampled
2003-2004, land cover did not correlate with MeHg in lake water or biota, possibly
because variation in wetland cover was low across watersheds (Yu et al.. 2011). In a
survey that included 21 freshwater, brackish, and marine wetlands and lakes in the
Mississippi River delta near Lake Pontchartrain, LA, MeHg levels in wetlands were
higher than in nearby rivers or lakes with similar salinity levels. In brackish and
freshwater wetlands, total MeHg concentrations (ng/L) and MeHg fraction (MeHg/total
Hg) were 2 or 3 times as much as in brackish or freshwater rivers (Hall et al.. 2008).
The relationship between Hg and MeHg concentrations may depend heavily on the type
and salinity of water body. A recent report on Hg concentrations in streams across the
U.S. found no relationship between Hg and MeHg across sites (Wentz et al.. 2014). This
is in contrast to an earlier review of Hg methylation and demethylation dynamics found
that across studies, there is a positive correlation between Hg and MeHg in estuaries
(R2 = 0.78), and weaker positive correlations in rivers (If = 0.68) and lakes (If = 0.64),
but no correlation in wetlands (Benoit et al.. 2003). In a recent study of eight highly
Hg-contaminated (up to 226,000 ng Hg/g sediment, reported as 226 |ig Hg/g sediment)
lakes and water bodies in Sweden, more than 55% of the variation in sediment MeHg was
explained by total sediment Hg in three estuary sites, but there was no correlation in the
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remaining freshwater sites (Drott et al.. 2008). In the freshwater sites, MeHg was instead
strongly correlated with potential methylation rates [r1 > 0.85 (Drott et al.. 2008)1.
Biology of Sulfate-Reducing Prokaryotes
The 2008 ISA attributed Hg methylation to sulfur-reducing bacteria, and there is new
evidence of the link between microbial sulfate reduction and Hg methylation. New (and
key older) studies include experimental inhibition of different microbial groups in order
to demonstrate microbial mechanisms, molecular sequencing techniques to identify
potential microbial Hg methylating strains, and sequencing of environmental or
experimental samples to determine microbial dynamics of Hg methylation. These studies
were conducted with samples from rivers, lakes, wetlands, and laboratory mesocosms
(Table 124). Current knowledge suggests mercury is methylated in the environment by
certain strains of sulfur-reducing and iron-reducing bacteria in Deltaproteobacteria
(Desulfovibrio, Desulfotomaculum, Desulfobulbus, and Geobacter genera) as well as
methanogens in Archaea, while the genetic potential for methylation has been identified
in the bacterial Firmicutes (Paranjapc et al.. 2017).
Increasing SOx in the microbial environment increases the metabolic activity and growth
of sulfate reducers, and one of the metabolic activities that can be stimulated is the
methylation of Hg. Peat samples from the Bog Lake fen in the Macell Experimental
Forest showed that experimental increases of wet S deposition to historical levels (32 kg
S/ha/yr) changed entire bacterial community composition. Experimental S addition also
significantly changed the composition of deltaproteobacterial communities, which
contain the SRB and IRB, and this community change correlated with bog %MeHg
(Strickman et al.. 2016). In recovery sections of the same wetland (S treatment of 32 kg
S/ha/yr had ceased 3 years before sampling), entire bacterial community and
deltaproteobacterial community compositions were not significantly different from
control wetland communities (Strickman et al.. 2016). suggesting a rapid response time
of microbial communities to changes in SOx deposition.
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Table 12-4 New studies on the biology of sulfate-reducing prokaryotes.
Additions or
Type of	Load Biological and Chemical
Ecosystem (kg S/ha/yr)	Effects	Study Site Study Species	Reference
Peatland
Treatment:
Bacterial communities in
Marcell
Pore water. Strickman et al.

32 kg
areas with increased
Experimental
bacterial (2016)

SO42 /ha/yr
sulfur deposition had
Forest, MN
communities

addition,
significantly different



control/ambient:
compositions compared



4.6 SO42"
to control and recovery



kg/ha/yr
areas. Bacterial species




diversity was significantly




higher in the control than




sulfate treated or recovery




groups.




Deltaprotobacteria were




significantly related to




%MeHg in bogs and close




to significant in lagg areas




(p = 0.057).


River
Deposition not
Sulfate reduction rates
Mesocosms
Avramescu et al.

reported
were positively correlated
of sediments
(2011)


with potential Hg
from St.



methylation rates
Lawrence



(r2 = 0.98).
River



Methane production in




sediment slurries




correlated positively with




demethylation rates




(r2 = 0.61).




Direct inhibition of




methanogens decreased




demethylation by 83%.


Floating
2.0 mM
Addition of M0O42"
Sunday Lake,
SDhaanum sdd. Yu et al. (2010)
bog
(190 mg/L)
decreased MeHg
Adirondack


sulfate added to
production by 44%.
Mountains,


slurried
Potential Hg mthylation
NY


samples
rates (%MeHg/day) were




2.1x higher with SO4




amendment.


Lakes
Deposition not
Addition of M0O42"
Lake
Samples of Gascon Diez et al.

reported
decreased methylation by
Geneva,
Darticles from (2016)


60-90% in settling
Switzerland
water column and


particles, and by 80% in

from lake bottom


sediments.

sediments
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Table 12-4 (Continued): New studies on the biology of sulfate-reducing
prokaryotes.
Type of
Ecosystem
Additions or
Load
(kg S/ha/yr)
Biological and Chemical
Effects
Study Site Study Species
Reference
Lakes
Deposition not
reported
Addition of M0O42"
decreased MeHg
production by 30% for E.
crassipes and by 60% for
P. glabra.
qPCR of bacterial genes
indicated that 3.34% of
the total bacterial
community in the
periphyton belonged to
the SRB families
desulfovibrionaceae and
desulfobacteraceae.
MeHg fraction was
positively correlated with
relative abundance of
desulfobacteraceae in P.
glabra periphyton.
Amazon
Oxbow
Lakes,
Bolivia
Eichhornia
crassipes and
polygonum
densiflorum
(currently named
persicaria glabra)
Acha etal. (2011'
Freshwater
marsh
W3 pore water
[S042"] <4 |jM
U3 pore water
[SO4"] is 39 |jM
F4 pore water
[SO42"] is 74 |jM
Deposition = not
measured
SRB abundance is 6.9x
higher in U3 than W3 and
16.8x higher in F4.
In W3, abundance of SRB
and methanogens are not
significantly different. In
U3, SRB abundance is
80% higher than
methanogen abundance.
In F4, SRB abundance is
60% higher than
methanogens.
mcrA copy number
correlates positively with
mRNA and methane
production rates.
Acetotrophic
methanogens are
dominant at W3, while at
high S U3 and F4 sites,
hydrogenotrophic
methanogens are
dominant.
F4, U3, and
W3 sites in
Water
Conservation
Area,
Everglades,
FL
Methanogens as
quantified by mcrA
copies
Sulfur-reducing
bacteria as
quantified by dsrB
copies
Bae etal. (2015)
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Table 12-4 (Continued): New studies on the biology of sulfate-reducing
prokaryotes.
Additions or
Type of	Load
Ecosystem (kg S/ha/yr)
Biological and Chemical
Effects
Study Site Study Species
Reference
Marine Laboratory
Methylation increases	Laboratory
with increasing bacterial	cultures
biomass (r = 0.83).	Sediments
During exponential growth	from Berre
phase, S addition	Lagoon,
increases net MeHg	France
production by 50%.
hgcAB expression levels
do not correlate with
methylation capacities.
Desulfovibrio	Goni-Urriza et al.
dechloracetivorans, (2015)
strain BerOCI from
sediments
cultures with
different C
sources
(pyruvate,
fu ma rate,
lactate, ethanol,
or malate),
control and
additional
sulfate (30 mM)
C = carbon; ha = hectare; kg = kilogram; MeHg = methylmercury; mM = millimolar; |jM = micromolar; S = sulfur; S042 = sulfate;
SRB = sulfur-reducing bacteria; yr = year.
Sulfate-reducing bacteria (SRB) and iron-reducing bacteria (IRB) are primarily
responsible for Hg methylation in natural ecosystems. On a broad scale, these bacteria
metabolize organic C by pairing its oxidation with the reduction of sulfate and/or Fe3+.
The sulfide produced by the reduction of SOx can act as a negative feedback on SRP
activity and Hg methylation. Methanogens, SRB, and IRB are facultative anaerobes,
(Paraniape et al.. 2017). and their presence and activity within wetland and freshwater
ecosystems is limited to zones with low oxygen availability. Experimental inoculations
suggest that microbial methylators are capable of forming syntrophies, in which microbes
link their metabolisms by the exchange of metabolic substrates and products. This
suggests that sulfate stimulation of methylation may involve more types of organisms
than just the sulfate-reducing bacteria and has led to the classification of the group of
interest as sulfate reducing-prokaryotes.
Specific inhibition of SRPs using molybdate (M0O42 ) is a common tool in studies that
determine the contribution of SRPs to Hg methylation. An older study of acidic Dickie
Lake in Ontario used different microbial inhibitors to determine whether SRPs or
methanogens within the sediments were responsible for Hg methylation. Molybdate
inhibited sulfate reduction and decreased %MeHg by 75% of production in controls,
while 2-bromo-ethanesulfonate (BESA), an inhibitor of methanogens, did not
significantly decrease MeHg production (Kerry et al.. 1991). Researchers inferred that
SRPs, not methanogens, were responsible for methylation in the lake, and sulfate
reduction rates increased 0.35 mg sulfate reduced/L/day with a 1 mg/L increase in sulfate
addition. However, additions of sulfate to concentrations of 5-25 mg sulfate/L in
sediment slurries did not affect %MeHg, which was 0.20-0.25% MeHg/g sediment
(Kerry et al.. 1991). A similar relationship between MeHg and sulfate reduction was
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demonstrated using molybdate inhibition in coastal Georgia marsh soils, where Mo
addition decreased sulfate reduction by 90% and decreased MeHg production by 85%
(King et al.. 1999). indicating that SRPs were involved in both processes.
An early laboratory study cultivated strains of a SRB, a methanogen, or an acetogen with
an Hg spike. The SRBs were the only microbes that methylated added Hg, and they did
so at a rate that resembled the rate in aquatic sediments, while molybdate inhibited MeHg
production and demethylation of added MeHg (Pak and Bartha. 1998). In a more recent
study, incubations of St. Lawrence River sediments with different inhibitors showed that
molybdate inhibited sulfate reduction completely, but did not affect methane production
(Avramcscu et al.. 2011). indicating that methanogens did not play an important role in
reducing S. Moreover, sulfate reduction rates were strongly positively correlated with
potential Hg methylation rates (r2 = 0.98). However, when methanogens were inhibited,
methylation of Hg increased 16% (Avramcscu et al.. 2011). indicating that methanogens
compete with SRPs and can depress Hg methylation. In Everglades, FL sediment
incubations, molybdate decreased MeHg production by 90% (Gilmour et al.. 1998). In
incubations of Sphagnum sampled from Sunday Lake, NY, molybdate decreased MeHg
production by 44% (Yu et al.. 2010). In core incubations from lakes in Wisconsin,
including both basins of Little Rock Lake, molybdate reduced MeHg production by 50%
(Gilmour et al.. 1998).
In wetlands within the Allequash Creek watershed, WI, sediment incubations with added
14C-labelled substrate and sodium molybdate suggested a diverse variable microbial
methylation community (Crcswell et al.. 2017). In some incubations, addition of
molybdate increased 14CC>2 produced from the substrate, suggesting that iron-reducing
methylators were dominant; in some incubations molybdate decreased 14CC>2, suggesting
that sulfate-reducing methylators were dominant. Incubations from the lower part of the
wetland produced 14CH4 as their dominant product, suggesting that methanogens
dominated the microbial methylation at this location (Crcswell et al.. 2017). Dominant
microbial methylators varied seasonally and spatially, including by sediment depth.
In Lake Geneva, Switzerland, mercury methylation rates were an order of magnitude
higher in settling particles than in sediments, and molybdate inhibition suggested that
sulfate-reducing methylators were responsible for 60-90% of methylation in settling
particles, and 80% of methylation in sediments (Gascon Diez et al.. 2016). Gascon Diez
et al. (2016) suggest that this study may be relevant to the Great Lakes, which like Lake
Geneva, are large, deep, and experience algal blooms due to nutrient inputs.
New research suggests that SRPs are important methylators of mercury in disturbed
terrestrial soils as well. In a Swedish study of the effects of forest harvesting on mercury
dynamics, clear-cutting induced changes in mercury methylators as inferred from soil
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incubations (Kronbcrg et al.. 2016). Additions of sulfate and iron enhanced potential
methylation rates in clear-cut but not reference forest soils, indicating that SRPs and
iron-reducing bacteria were both active methylators in disturbed soils with high water
tables (see Appendix 12.3.2V The role of SRPs was confirmed by inhibition of
methylation by molybdate in soils from clear-cut stands, and a role of methanogens in
methylating mercury was shown by BES inhibition of methylation (Kronbcrg et al..
2016).
The 2008 ISA did not address the role that periphyton play in hosting SRPs and boosting
Hg methylation rates within the oxic water column of aquatic and wetland ecosystems.
This section summarizes new (and key older) papers that demonstrate that SRB in
periphyton methylate Hg in South American lakes as well as North American lakes and
wetlands. Periphyton are ecologically important as they form the base of the aquatic food
chain for many invertebrates and vertebrates. Periphyton mats are biofilms of algae,
bacteria, fungi, microinvertebrates, organic detritus, and inorganic particles, embedded in
a polysaccharide matrix, typically attached to a substrate of sediments or submerged
macrophytes. The microbial communities embedded within the periphyton can be quite
complex and diverse, including autotrophs and heterotrophs, SRB and S oxidizing
bacteria, and methanogens; periphyton microbial composition and microbial activity
depend on site conditions (Correia et al.. 2012). The polysaccharide matrix slows
exchange with the water column while intense aerobic metabolism at the matrix surface
quickly consumes oxygen diffusing into the periphyton, creating anoxic, reducing
microenvironments in the interior of periphyton where anaerobes including SRPs thrive.
In the case of periphyton attached to macrophytes, plant exudates are an important source
of carbon for periphyton microbes, boosting activity above rates of activity of similar
microbes residing in sediments.
SRPs embedded in periphyton are more efficient in methylating Hg than are SRPs in
sediments. Much of the research demonstrating Hg methylation by periphyton has
occurred in South America, in tropical lakes where the genus Eichhornia (water
hyacinth) is native, and where ambient temperatures can support high levels of bacterially
mediated Hg methylation (Guimaraes et al.. 2006). Eichhornia crassipes is a wetland
plant present in 24 states and listed as a noxious or invasive weed by 7 states (USDA.
2015b). Closely related Eichhornia azurea is currently present only in Florida; it is listed
by the Animal and Plant Health Inspection Service of the U.S. Department of Agriculture
(USDA-APHIS) as a noxious weed (USDA. 2015b). In incubations of chopped
Eichhornia crassipes roots spiked with 203HgCl2, net MeHg production was high,
1.6-30.2% of total Hg (Guimaraes et al.. 2006). Other studies have shown a similarly
wide range of methylation by periphyton, with high rates exceeding sediment methylation
rates. Intact E. crassipes periphyton incubations had potential methylation of 12.1-25.2%
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of added Hg, and periphyton isolated from Polygonum densiflorum [hereafter called by
its current accepted scientific name Persicaria glabra; a wetland obligate plant also
native to the U.S., classified as endangered in Maryland and New Jersey, (USDA.
2015b)! had potential methylation rates of 0.2-36.1%; by comparison, sediment from the
same tropical lakes produced 6.4-12.5% MeHg in incubations (Correia et al.. 2012).
Mo inhibition of sulfate reduction also inhibits MeHg production in periphyton,
confirming that SRPs are responsible for a significant portion of Hg methylation within
these complex microbial communities. Recently, Achaetal. (2011) and Correia etal.
(2012) collected macrophyte samples from oxbow lakes in the Bolivian Amazon for
inhibition experiments. Incubations with an array of guild-specific inhibitors show the
relative contribution of different microbes to Hg methylation. One set of incubations of
periphyton associated with the roots of E. crassipes or P. glabra suggested that SRB were
the most important component of the periphyton in producing MeHg. Molybdate, an
inhibitor of sulfate reduction, decreased MeHg production by 30% fori?, crassipes and
by 60% for P. glabra (Acha etal.. 2011). indicating that in intact periphyton, SRB alone
are responsible for 30-60% of Hg methylation. Quantitative PCR of bacterial genes
indicated that 3.34% of the total bacterial community in the periphyton belonged to the
SRB families Desulfovibrionaceae and Desulfobacteraceae, and MeHg fraction was
positively correlated with relative abundance of Desulfobacteraceae in P. glabra
periphyton (Achaetal.. 2011). A second set of incubations of six tropical macrophyte
species-associated periphyton confirmed that inhibition of sulfate reduction and SRB
using molybdate decreased MeHg production, but showed that inhibiting both SRB and
methanogens resulted in a 90% reduction of MeHg production across sediments and
periphyton (Correia etal.. 2012). Algaecide or fungicide additions significantly inhibited
MeHg formation in some but not all incubations, indicating that the microbial
composition or activity varied across periphyton samples, even when they were collected
from the same site or from the same host plant species grown at different sites (Correia et
al.. 2012). Together, these studies show that SRB are important in producing MeHg
within periphyton, but that other microbes residing in the periphyton contribute through
poorly understood mechanisms to total MeHg production.
Mercury methylation by periphyton has also been documented within North American
ecosystems. In an older study in Dickie Lake in Ontario, measured Hg concentrations in
periphyton were 56.5 ng/g, 22 times higher than Hg concentrations in sediments (Kerry et
al.. 1991). A more recent study of periphyton growing in the epilimnion of boreal Lake
Croche in Quebec demonstrated that Hg methylation rates are faster in periphyton,
reaching steady-state in 12 hours, than the 4-8 days to steady state of MeHg production
in sediments (Desrosiers etal.. 2006). MeHg production by periphyton in the boreal lake
was dependent on SRB as well as on photosynthesizers, as demonstrated by inhibition
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experiments. Despite the lower temperatures of incubations to reflect the lower
temperatures of the boreal lake, potential MeHg production was 16.6-18.5% (Dcsrosicrs
ct al.. 2006). similar to MeHg production in tropical periphyton.
In the Florida Everglades, periphyton are the base of the aquatic food chain. In an older
study, incubation of periphyton collected from different sites along the eutrophic gradient
between the EAA to the north and Everglades National Park to the south found MeHg
production in periphyton from all sites, although rates varied widely (C'lcckncr ct al..
1999). MeHg potential production rates were highest in periphyton collected at the
northern, eutrophic site, and MeHg production correlated positively with both S reduction
and S oxidation by photosynthesizing purple S bacteria (Cleckner et al.. 1999).
Periphyton may be an important direct source of MeHg to the food chain in the Florida
Everglades.
The gene pair conferring the ability to methylate Hg has been identified only recently,
after the 2008 ISA. No ecological or evolutionary advantage of the ability to methylate
Hg has been established (Kcrin et al.. 2006; Benoit et al.. 2003). and it appears to be an
inadvertent transformation by a corrinoid-dependent protein produced by prokaryotes for
a different metabolic pathway (Gilmour etal.. 2013). Recent phylogenetics work
indicates that the gene pair hgcAB confers the ability to methylate Hg, and that the gene
pair and Hg methylation are present in some but not all species within the SRB,
iron-reducing bacteria, syntrophic sulfur reducers in the Syntrophobacterales,
methanogens in the Archaea, and bacteria in the Clostridia (Gilmour et al.. 2013). The
confirmed and predicted Hg methylators from this study occur across alkaline, neutral, or
acidic environments, and are mostly heterotrophs that use sulfate, iron, and CO2 as
terminal electron acceptors.
MeHg production rates vary among species, and even among strains within species (Shao
et al.. 2012). A recent study of hgcAB in the freshwater marshes of the Florida Everglades
showed that syntrophs dominated the microbial methylators (49-65% of sequences) and
also were the dominant sulfate reducers, as shown by dsrB mRNA quantification
(75-89%) of sediments sampled within the water conservation areas (Bae et al.. 2014). In
a recent study of methylation using a strain of Desulfovibrio dechloracetivorans,
expression of the hgcAB gene did not correlate with methylation capacities of pure
cultures, which supports the supposition that Hg methylation is not the main purpose of
the proteins encoded by hgcAB. However, during the exponential growth phase of the
cultures, addition of 2,900 mg/L (30 mM) sulfate increased net MeHg production by 50%
(Goni-Urriza et al.. 2015).
Demethylation is also an important microbial Hg transformation to consider in natural
ecosystems. The capacity to transform MeHg to inorganic Hg (elemental Hg or Hg2+) is
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widely distributed and occurs by different mechanisms in different microbial groups.
Microbial demethylation by aerobic and facultative Hg-resistant microbes possessing the
mer operon break MeHg into methane and Hg° via reductive degradation, whereas
methanogens, SRPs, and other prokaryotes break MeHg into carbon dioxide, methane,
and Hg2+ via oxidative demethylation. A recent study of St. Lawrence River, Ontario,
Canada sediments showed that methane production in sediment slurries correlated
positively with demethylation rates (r2 = 0.61), and direct inhibition of methanogens
decreased demethylation by 83% (Avramcscu ct al.. 2011). An earlier study tested a pure
culture of the archaeon Methanococcus maripaludis and found that it demethylated added
MeHg despite no observed methylation potential (Pak and Bartha. 1998).
Microbial demethylation also occurs in the same prokaryotes that methylate Hg; a
laboratory incubation of pure cultures demonstrated that Desulfovibrio desulfuricans (an
SRB that methylates Hg) demethylated added MeHg at rates similar to those of incubated
aquatic sediments (Pak and Bartha. 1998). In the same study of sediments of the St.
Lawrence River, Ontario, inhibition of both SRPs and methanogens decreased
demethylation of Hg by 96% (Avramcscu et al.. 2011). indicating that while
methanogens are responsible for the bulk of demethylation in these sediments, SRPs are
involved in demethylation.
New evidence of the genetic basis of Hg methylation and the activity of microbial
methylators in the environment and in the laboratory is consistent with the 2008 ISA's
link between sulfate reduction and Hg methylation. Mercury methylation by SRP is
determined by the same suite of environmental factors that enhance SRP activity or
abundance: sulfate, organic matter, and anoxic-oxic conditions (Figure 12-7). Mercury
methylation occurs in wetlands, lakes, reservoirs, streams, and rivers. Algal blooms,
periphyton, stratification of the water column, and fluctuating water levels all create
favorable conditions for sulfate stimulation of mercury methylation.
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log S04, uM
Source: from Glimour (2011).
Figure 12-7 The effect of different environmental factors on the relationship
between sulfate and mercury methylation. Methylmercury (MeHg)
accumulation is minimal at low and high sulfate concentrations,
with an optimum near 100 |jM sulfate (blue line). High dissolved
organic matter (DOM) will increase the magnitude of MeHg
production across the range of sulfate concentrations (red line).
Sulfide produced by methylators will inhibit further methylation if
it accumulates in the aqueous environment where methylation
occurs, shifting the MeHg optimum left (purple line). However if
ecosystem chemistry (iron [Fe], reoxidation of sulfide, organic
matter [OM]) allows for rapid sequestration of sulfide to large
particles and sediments, the relaxation of the negative feedback
of sulfide to sulfur-reducing prokaryotes (SRPs) will shift the
MeHg optimum right (green line).
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1	12.3.3. Environmental Drivers of Mercury Methylation Potential
2	The 2008 causality statement linked S to increased Hg methylation when other conditions
3	are favorable, with a particular emphasis on dissolved organic carbon (DOC) as a control
4	on Hg methylation (U.S. EPA. 2008a'). This section is an overview of factors that control
5	Hg methylation. In a review, Bigham et al. (2017) summarized drivers of mercury
6	methylation in terms of their effect on mercury availability to methylators and on
7	microbial activity associated with Hg methylation (Table 12-5).
Table 12-5 Environmental factors that affect mercury (Hg) methylation.
Geochemical or Bioavailability Microbial
Physical Parameter of Hg to	Methylator
Affects	Methylators	Activity	Notes
Sulfur	Increase or Increase
decrease
Bioavailability of inorganic Hg will increase or decrease
depending on Hg-S bonds and chemical species
formed.
Increased sulfate can stimulate activity of
sulfate-reducing prokaryotes, some which may be
capable of Hg methylation.
Organic carbon	Increase or Increase
decrease
Availability of inorganic Hg can decrease if Hg binds to
organic carbon molecules of a size or structure that
makes the molecules resistant to microbial degredation.
Availability of inorganic Hg can increase if Hg binds to
organic molecules that are mobile in water and
accessible to microbial degredation.
Increased organic carbon can stimulate microbial
activity by increasing electron donors to metabolism.
Iron
Decrease	Increase or Bioavailability of inorganic Hg can decrease as Fe
decrease	reduces and binds to Hg.
Fe can increase microbial methylation by binding to
sulfide and preventing the negative feedback of this
product of sulfate reduction upon its production, which
coincides with Hg methylation in some SRP. Also, Fe
can stimulate the action of iron-reducing microbial
methylators.
Fe can decrease microbial methlyation by shifting
microbial communities towards nonmethlyating
organisms.
PH
Increase or
decrease
Increase or
decrease
Bioavailability of inorganic Hg is reduced at high pH.
pH between 4.5 and 9 is optimal for microbial activity.
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Table 12-5 (Continued): Environmental factors that affect mercury (Hg)
methylation.
Geochemical or
Physical Parameter
Affects
Bioavailability
of Hg to
Methylators
Microbial
Methylator
Activity
Notes
Temperature
Increase or
decrease
Increase or
decrease
Bioavailability of inorganic Hg can increase if
weathering of Hg-containing rocks or sediments, or
decomposition of older organic matter, increases with
higher temperatures.
Bioavailability of Hg can decrease if Hg volatilization to
the atmosphere increases.
Microbial methylators have optimal temperatures for
growth.
Oxygen availability
Increase
Decrease
Bioavailability of Hg in reduced chemical species is low.
Microbial methylators are anaerobes, methylating Hg
under reducing conditions.
Drying and wetting
cycles (drought,
seasonal variation in
water levels, etc.)
Increase
Increase
Fluctuating water levels can increase the release of Hg
from particles in sediment.
Fluctuating water levels can increase the reoxidation
and release of S from sediments as sulfate.
Adapted from Biaham et al. (2017)
Seasonal patterns of Hg methylation and transport of Hg in many systems can affect
MeHg accumulation. Also, a number of chemical constituents can stimulate or inhibit Hg
methylation. Given our understanding of the microbial mechanisms of Hg transport and
methylation, and the complexity and diversity of aquatic environments, chemical controls
on MeHg processes are not fully established. This section is meant to lend context to the
description of how S deposition affects methylation, but it is not a complete review of the
chemical controls on biological Hg methylation in natural environments.
12.3.3.1. Sulfur and Methylation Potential
Sulfate in soil pore water, surface water, or in sediment pore waters can stimulate
mercury methylation. Sulfate is utilized as a terminal electron acceptor in heterotrophic
metabolism by some classes of bacteria, and in the course of this dissimilatory sulfur
reduction, these bacteria transform sulfate to sulfide and expel the reduced sulfur into the
environment as a waste product. Sulfate-reducing bacteria can also form syntrophies,
symbiotic exchanges of metabolic substrates and products, with methanogens and other
archaea, so that sulfate stimulates metabolism even in organisms that do not utilize it
directly as an electron acceptor. The current scientific understanding is that MeHg is
produced as an accidental byproduct in both sulfate-reducing bacteria and syntrophic
archaea. Conditions optimal for these organisms and which favor their activity over
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competing anaerobic heterotrophs probably occur in short time frames and in shifting
locations in many ecosystems. Sulfate stimulation of mercury methylation will probably
occur at an annual time scale at very low or undetectable background levels, with
occasional peaks of microbial methylation (hotspots) when other controls (temperature,
organic matter, anoxic microsites, or hypoxic zones) on microbial metabolism reach their
optima. At temporal and spatial hotspots of mercury methylation, sulfate concentrations
may drop as sulfate is converted to sulfide, particularly in shallow water bodies with
longer water residence times. As a result, concentrations of sulfate in surface water may
not correlate with sulfate inputs to the ecosystem or with MeHg produced by sulfate
reduction, particularly in aquatic or wetland ecosystems where there are active SRPs.
The byproduct of microbial sulfate reduction, sulfide, is also an important regulator of
mercury methylation. Sulfide dissolved in surface water or pore water inhibits microbial
methylation, and interferes with plant nutrient uptake (see Appendix 12.2.3). Sulfide can
also produce a negative feedback on mercury methylation by binding with inorganic
mercury in the water column and precipitating into cinnabar in the sediment, lowering the
bioavailability of Hg to microbial methylators. However, in water bodies with a high
fraction of mineral sediments or dissolved minerals, particularly iron, iron will bind to
sulfide and precipitate out of solution, preventing the formation of a negative feedback to
sulfur reduction and mercury methylation. In water bodies with longer residence times,
sulfide concentrations in sediments may correlate with microbial sulfate reduction, and if
SRPs are the primary methylators in the system, with MeHg concentrations.
Laboratory studies support field research in demonstrating that under controlled
conditions, sulfate additions increased Hg methylation in environmental samples. The
2008 ISA reported evidence of Hg methylation in lake water and sediment samples in
response to experimental sulfate addition. In slurries of sediment collected from Quabbin
Reservoir, MA, adding 4.8 mg/L (reported as 50 (j,M) sulfate to the slurry (ambient lake
water sulfate = 5.8-7.7 mg/L or 60-80 (iM) increased potential MeHg production 150%,
adding 9.6 mg/L sulfate (100 (j,M) increased MeHg production 160%, and adding
19.2 mg/L sulfate (200 (j,M) sulfate increased MeHg 410% above ambient lake water
MeHg (Gilmour et al.. 1992). Slurries can overestimate the potential rates of MeHg
production, so this study also incubated intact sediment cores with sulfate concentrations
from 0.3-110 mg/L (reported as 3-1,140 (j,M). Peak potential MeHg production was at
11 mg/L sulfate (110 (j,M), where MeHg was 8.6 ng MeHg/g, a 200% increase from
MeHg production at 0.3 mg/L sulfate [3 (j,M (Gilmour et al.. 1992)1. In longer
incubations of sediment cores from two lakes in Wisconsin, including Little Rock Lake,
adding 5.8-23 mg/L sulfate (60-240 (iM) changed sediments from net demethylating
(-100 ng MeHg/m2/day or -0.1 |ig MeHg/m2/day) to net methylating
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(5,500 ng MeHg/m2/day or 5.5 |ig MeHg/m2/day) environments (Gilmour and Ricdcl.
1995V
Since the 2008 ISA, several new studies have provided experimental evidence of sulfate
stimulation of Hg methylation in a broader range of aquatic freshwater environments
(Table 12-6). Laboratory experiments using samples from wetlands at Sunday Lake in the
Adirondack Mountains of New York were conducted to determine potential Hg
methylation rates. When slurried peat-and-water samples from a floating fen (bog mat
composed of Sphagnum spp. and ericaceous shrubs) received an addition of sulfate to
raise the concentration of sulfate by 190 mg/L (initial sulfate concentration not given),
Hg methylation rates (%MeHg/day) were 2.1 times higher than unamended samples (Yu
ct al.. 2010). This result is consistent with evidence from observational (Appendix 12.3.5)
and S addition studies (Appendix 12.3.4) in peat bogs that sulfate stimulates Hg
methylation in these wetland systems.
There are three new studies that demonstrate sulfate stimulation of water and sediment
Hg methylation in samples taken from rivers. Slurried sediments collected in 2010 from
the South River, VA, were spiked with sulfate to raise sulfate levels from average
ambient levels of 19.2 mg/L to 38.2 or 96.1 mg sulfate/L. Sulfate addition significantly
increased the potential methylation rate of Hg (%/day), with methylation rates
1.6-2.6 times higher in sulfate-spiked sediments than in sediments with ambient sulfate
concentrations (Yu et al.. 2012). There was a significant difference in methylation rates
between the low and high S addition treatments in only one of three sampled river
sediments; this sample showed methylation was higher in the sediment when sulfate
levels were 96.1 mg/L (Yu et al.. 2012). In microcosm slurries composed of water and
sediment samples from the Wupper River in Germany, higher MeHg in water correlated
weakly (If = 0.28, both variables natural log-transformed) with higher sulfate
concentrations, with mean pore water sulfate of 32.2 mg/L and range of 2-223 mg/L
(Frohne et al.. 2012). Higher total Hg concentrations in water correlated weakly
(If = 0.20) with higher sulfate, although DOC was a stronger predictor of total Hg
[i?2 = 0.53 (Frohne et al.. 2012)1. Tsui et al. (2008) constructed mesocosms to mimic leaf
decomposition in the hyporheic zone of Minnesota streams and rivers, where conditions
are hypothesized to be favorable for the SRB that methylate inorganic Hg. Water samples
were collected from seven streams or rivers in Minnesota, with sulfate levels at collection
of 4.0-26.2 mg sulfate-S/L (with an ultrapure water sample as a control, 0.0 mg S/L), and
incubated w ith Acer sp. (maple) leaves. After 66 days of decomposition, the total Hg
released from the decomposing litter into solution as dissolved Hg increased with initial
sulfate concentration (R2 = 0.541,/? = 0.038), and the fraction of total Hg that had been
transformed to MeHg increased in a power law dependence upon initial sulfate
concentration [i?2 = 0.572,p = 0.030, see Figure 12-8. (Tsui et al.. 2008)1.
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o>
O)
X
35
30
25 ¦
20 ¦
15 ¦
10 ¦
5
0
100
80
60
40
20
0
10	20
DSO, (mg S L"1)
30
cn
X
CD
%MeHg = methylmercury fraction; DHg = mercury dissolved in mesocosm water; DS042 = sulfate dissolved in mesocosm water;
L = liter; mg = milligram; ng = nanogram; S = sulfur; THg = total mercury.
Source: Figure 2A in Tsui et al. (2008).
Figure 12-8 The relationship between surface water sulfate and total mercury
or methylmercury fraction in river/leaf litter mesocosms.
This evidence of sulfate stimulation of Hg methylation in river ecosystems is consistent
with observational studies of positive correlations between sulfate and Hg methylation in
rivers and streams.
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Table 12-6 New mesocosm or incubation studies on sulfur addition effects on
methylmercury.
Type of Additions or Load Biological and Chemical	Study
Ecosystem (kg S/ha/yr)	Effects	Study Site Species
Reference
River	S deposition not
reported
Soil:
2.0-2.6 mg S/g soil
[SO42"] pore water
in microcosm:
2.2-223 mg/L,
mean 32.0 mg/L
Total Hg and MeHg were
positively correlated with
sulfate concentrations:
ln(THg) = 2.956 + 0.697 x In
(S042"),
ln(MeHg) = 4.962 + 0.473 x In
(S042").
Wupper River,
Germany
Sediment
pore water
Frohne et al.
(2012)
Floating
bog
2.0 mM (190 mg/L)
sulfate added to
slurried samples
Potential Hg methylation rates
(%MeHg/day) were 2.1 *
higher with SO42"
amendment.
Sunday Lake,
Adirondack
Mountains, NY
Sphagnum
spp.
Yu et al. (2010)
River
Ambient
Sediment MeHg (ng/g)
South River,
In-channel
Yu et al. (2012)

concentration in
increased linearly with pore
VA
surface


water is 200 |jM
water sulfate (|jM)

sediments


(19.2 mg/L) sulfate
concentrations in the river,





MeHg = 0.61 x [SO42"] - 0.08.



Addition of 400 |jM
(38.4 mg/L) or
1,000 |jM
(96.1 mg/L) sulfate
in anoxic sediment
slurries
SO42" addition in slurries
increased potential
methylation rates by 1.6-2.6x.
Fluvial
zones
(streams
and rivers)
S deposition not
reported
Initial water SO42"
concentrations
(mg/L): range
4.0-26.2, mean
9.2, median 6.1
Following aquatic anoxic
decomposition of leaf litter,
total dissolved Hg increases
in decreasing proportion to
initial SO42" concentration.
Dissolved %MeHg increases
in decreasing proportion to
initial SO42" concentration.
Mesocosms
constructed
with samples
from Cedar
Creek LTER
(leaf litter) and
seven streams
or rivers
(water), MN
/Acersp. and
stream
microbial
communities
Tsui et al. (2008)
g = gram; ha = hectare; Hg = mercury; kg = kilogram; L = liter; MeHg = methylmercury; mg = milligram; mM = millimolar;
|jM = micromolar; ng = nanogram; S = sulfur; S042" = sulfate; THg = total mercury; yr = year.
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12.3.3.2. Total Mercury Concentration and Methylation Potential
The 2008 ISA did not present information on how total Hg concentration affects rates of
MeHg production, in part because there is no consensus across studies as to whether there
is a linear relationship between total Hg and MeHg in water or soil. In part this is because
most reported MeHg rates are potential rates determined by experimental incubations
with added Hg, and methylation rates have been shown to be dependent on initial
concentrations of Hg (King et al.. 1999; Gilmour et al.. 1992). Additionally, different
chemical forms of inorganic Hg vary in bioavailability and affect methylation rates
(Kucharzvk et al.. 2015; Graham et al.. 2012).
Two studies that sampled MeHg fractions at different locations within the U.S. provide
information at the landscape level about factors that increase probability of detrimental
effects upon wildlife due to Hg availability. There is a large variability across systems in
MeHg fraction. In a study of five forested watersheds distributed across the U.S. (in
Vermont, Wisconsin, Colorado, Georgia, and Puerto Rico), total Hg flux out of the
watershed in the streams was highest at Rio Icacos, Puerto Rico (54,400 ng Hg/m2/yr,
reported as 54.4 |ig Hg/m2/yr) and lowest at Allequash Creek, WI [250 ng Hg/m2/yr or
0.25 |ig Hg/m2/yr (Shanlev et al.. 2008)1. MeHg fraction was highest at Allequash Creek
(14.8 % of total Hg) and lowest at Rio Icacos (0.7%), indicating that watershed flux of
total Hg does not directly indicate risk to organisms (Shanlev et al.. 2008). In a study of
eight streams in Oregon, Wisconsin, and Florida, MeHg fractions and concentrations
(range: <0.01-17.8 ng/g sediment) in streambed sediments were higher in nonurban than
in urban streams (Marvin-Dipasquale et al.. 2009).
12.3.3.3. Temperature and Methylation Potential
The 2008 ISA did not describe temperature or seasonal effects on Hg methylation. This
section presents information from key older papers and recent research on temperature
effects in lakes, wetlands, and watersheds. Seasonal patterns of microbial methylation are
directly controlled by temperature effects on microbial metabolism, as well as by indirect
temperature effects, such as oxygen depletion and stratification and mixing of water
bodies.
As a microbial process, Hg methylation is temperature dependent. In Lake Clara, WI,
incubation at in situ temperatures of seasonally sampled sediment cores from a lake depth
of 10.5 m showed that Hg methylation increased slowly from near 0% in April to peaks
of 1% methylation in August and September, with temperature accounting for 30% of
methylation rate variation (Korthals and Winfrey. 1987). Demethylation was also
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measured in incubations and displayed slightly different seasonal variation; it increased
sharply from 1% in May to 4% in early July, then fell sharply to 1% in late July, staying
low for the rest of the sampling year. As a result, the methylation:demethylation ratio (a
predictor of MeHg concentrations in water) rose above 1 in late July through August, and
then again in late September (Korthals and Winfrey. 1987). Timing of methylation and
demethylation peaks were similar in sediments sampled from a lake depth of 1.0 m,
although gross methylation and demethylation rates were slightly higher (Korthals and
Winfrey. 1987). Sampling across the ELA in Ontario suggests that seasonal patterns of
methylation in lakes are due to seasonal shifts in oxygen availability, as locations and
depths within the lake where methylation was documented in the summer had no
methylation potential when resampled in the fall following turnover (Ecklcv and
Hint.e1ma.nn. 2006).
Fall mixing of stratified lakes can have significant effects on total MeHg in lake food
chains. In all years of sampling at Onondaga Lake, NY, there was a significant peak in
MeHg concentrations and in MeHg:total Hg ratios in the epilimnion in the fall following
turnover (Todorova et al.. 2014). In Little Rock Lake, WI, researchers inferred that
considerable demethylation occurred in the fall following mixing because high MeHg in
the summer water column did not accumulate in sediments and was not reflected in low
spring MeHg water column concentrations (Watras et al.. 2006). Short weather events
can also affect Hg exchange. During the fall of 2006, regular weekly sampling in
Onondaga Lake, NY, captured the effects of strong winds and heavy rains mixing the
stratified layers of the lake upon MeHg dynamics. Dissolved oxygen values dropped in
the sampled surface waters, and MeHg concentrations rose 23% to 0.236 ng/L in surface
waters, even as MeHg concentrations dropped in the hypolimnion (Todorova et al..
2014).
There are seasonal patterns of Hg methylation in wetlands as well. In peatlands of
Allequash Creek Watershed, WI, MeHg concentrations and MeHg fraction (percentage of
total Hg) peak in early fall (Creswell et al.. 2008). In the Adirondack Mountains, NY,
sampling at the inlet stream of Arbutus Lake showed a pulse of MeHg entering the lake
in summer months from the upstream wetlands (Gerson and Driscoll. 2016). Sampling of
the streams in the lake watershed showed no seasonal patterns of MeHg in streams
draining forest; but in streams draining wetlands, MeHg increased 300% during the
growing season compared to winter MeHg concentrations (Selvendiran et al.. 2008a'). In
freshwater marshes in the Yolo Bypass, CA, seasonal wetland export of MeHg occurs
during the winter as well as the summer. In the summer, newly methylated MeHg is
drawn into the root zone of the sediment by plant transpiration, and in winter, diffusion
releases MeHg from the sediment into the water column (Bachand et al.. 2014). Seasonal
changes have also been documented in the salt and freshwater marshes at Kirkpatrick
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Marsh on the Chesapeake Bay, MD, with higher MeHg concentrations and higher
methylation rates in summer than in fall (Mitchell and Gilmour. 2008). and temperature
effects have been experimentally demonstrated in Georgia coastal marshes, where the
methylation rate was 2 times higher at 25°C than at 4°C, and was 30 times higher at 35°C
than at 4°C (King ct al.. 1999). Seasonal patterns are even observed in the Florida
Everglades, where December potential methylation rates were lower than March or July
rates, although water temperatures at midday in December were 18°C (Gilmour ct al..
1998V
In addition to affecting Hg methylation rates, seasonal changes in hydrology can alter
fluxes of Hg in watersheds. In the Marcell Experimental Forest, MN, water flow during
snowmelt over 2 weeks in spring is 30 or 41% of total annual discharge from two wetland
catchments (Mitchell et al.. 2008c'). and 26 or 39% of total annual Hg export from the
catchment occurred during that time. The majority of MeHg within the catchments (79 or
95% of total catchment MeHg) was in the wetlands and exported in high amounts during
snowmelt, about 22-23% of annual MeHg export from the catchments (Mitchell et al..
2008c). In the Adirondacks, there were increases in stream MeHg concentrations and
%MeHg in summer months, when stream discharge was low and water residence time
was long in the wetlands where methylators were active (Gerson and Driscoll. 2016;
Selvendiran et al.. 2008a). High flow events can be important for transporting MeHg
through ecosystems but do not necessarily cause methylation of mercury in aquatic
ecosystems.
Climate Modification of S Effects on Ecosystems
Climate factors that will modify ecosystem response to S include changes in temperature
and intensification of hydrologic cycling (Paraniape et al.. 2017). Recent evidence from
experiments in lakes and bogs allows evaluation of how climate change will affect the
relationship between SOx deposition and MeHg production. Climate change will affect
aquatic and wetland ecosystems via warming temperatures and increasing fluctuations in
water tables due to changes in the timing and magnitude of precipitation (IPCC. 2013).
Generally, microbial activity increases with temperature. It is biologically plausible that
warming could increase MeHg via stimulation of both the decomposer guilds of microbes
producing DOC and of SRPs actively methylating Hg; however, it is also plausible that
warming will increase the activity of microbes that demethylate MeHg. IPCC (2013)
projected that heavy precipitation events will increase in number and strength due to
climate change, which can lead to rapid changes in water table level in wetlands and
lakes. These changes in water level can move pulses of labile carbon and sulfate into the
anoxic zones of the water, sediment, and periphyton where SRPs are active, and flooding
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or rewetting of previously dry zones can create new environments favorable for SRPs.
The following studies suggest that drought and warming increase S cycling and Hg
cycling. Drought, but not warming, increases MeHg production in lakes and wetlands.
As reviewed in the 2008 ISA, recovery in Little Rock Lake, WI, has been monitored
because the lake was experimentally acidified for 5 years (1984-1989). Drought affected
sulfate and MeHg concentrations in the lake, which is rain-fed. During the drought years
of 1998-2006, MeHg increased 0.004 ng/L/yr, although high variation among samples
made this trend marginally significant \p = 0.06 ("Watras and Morrison. 2008)1. MeHg
increased from 0.04 ng/L in 1997 to 0.07 ng/L in 2006, a 75% increase, which is more
than can be attributed to evapoconcentration from the lake's 30% decrease in volume
(Watras and Morrison. 2008). During this time period, there were no significant changes
in total Hg deposition; although between 1988 and 2006, the total Hg concentration in
Little Rock Lake has declined 0.04 ng/L/yr (Watras and Morrison. 2008). During the
drought, between 2000 and 2004, total MeHg mass in the north basin of the lake was
positively correlated with higher Hg and S deposition in precipitation (Figure 12-9).
Wetlands are characterized by fluctuating water tables, and microbial communities in
wetlands are adapted to respond rapidly to changes in substrate and oxygen availability.
This is particularly true for SRPs, which methylate Hg in geographic and temporal
pulses, resulting in hotspots of high MeHg concentrations within wetlands where Hg is
actively methylated (Supplemental Material). Fluctuations in water tables caused by
periods of drought and periods of flooding can create pulses of SRP activity and hotspots
of Hg methylation, raising MeHg concentrations in the environment. The following
studies present evidence that climate change (through drought and warming) will interact
with S addition to change MeHg and Hg dynamics in wetlands.
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in

ro
2000 2001 2002 2003 2004
Year
70 80 90 100110 4
HgT (mg- ha"1)
Atmospheric deposition rate
6 8 10
S04 (kg-ha"1)
ha = hectare; HgT = total mercury deposition kg = kilogram; MeHg = methylmercury; mg = milligram; S042- = sulfate.
Source: adapted from Watras and Morrison (2008).
Figure 12-9 Total methylmercury mass in water at Little Rock Lake, Wl,
annually (a), and in relationship to annual mercury (b) or sulfur (c)
deposition.
1	Drought and rewetting increased MeHg in Bog Lake Fen, and S addition treatments
2	enhanced this effect. A 9-month drought from the summer of 2006 to the spring of 2007
3	resulted in a pulse of reduced S when the water levels rose again, and the sulfate
4	concentrations were 147% higher in the S addition treatment, and 78% higher in the
5	recovery treatment, than in the control wetland (Wasik et al.. 2015). Sulfate pulses
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resulted in proportional increases in Hg methylation, with peak MeHg fraction 150%
higher in the S addition and 60% higher in the recovery treatment ("Wasik et al.. 2015).
The peat experienced drought conditions in summer 2007 as well, and pulses of sulfate in
the experimental treatments that fall (165% higher in S addition, 14% higher in recovery
wetland than in control wetlands) resulted in higher peak MeHg fractions in fall and the
following spring in both the S addition and recovery treatments. Experimental
manipulation of water levels in the different wetland treatments showed that higher water
levels raised sulfate, total Hg, and MeHg in S addition wetlands, but did not affect these
parameters in control and recovery wetlands (Wasik et al.. 2015). In the recovery
treatment, S and MeHg pulses were higher than in control plots, indicating persistent
effects of S addition several years after addition ceased. However, because
experimentally raising water levels did not alter S or Hg chemistry in control or recovery
wetlands, the recovery wetland was returning to S and Hg processes more similar to the
control plots.
Warming may ameliorate the effect of S upon MeHg production in wetlands, but
experimental evidence suggests that it will also increase sulfate and MeHg export from
peatlands into downstream aquatic ecosystems. Degero Stormyr in Sweden is the site of a
wetland experiment that addresses the effects of S and N deposition as well as warming
due to climate change. Warming altered the S and Hg dynamics under S deposition at
Degero Stormyr. Total Hg (per mass of peat) was 19% lower in 20 kg S/ha/yr + warming
plots than in S only plots, indicating increased mobilization or volatilization of Hg from
the peat. The Hg methylation constant in S + warming was similar to the rate in
unamended control plots, much lower than the rate in S only plots (Akerblom et al..
2013). MeHg concentrations in the S and warming treatment were lower than in any
other treatment, 73% lower in the 20 kg S/ha/yr + warming treatment than in the 20 kg
S/ha/yr treatment, and 54% lower than in the control treatment (Akerblom et al.. 2013).
This last result suggests that the combination of warming and S addition increases MeHg
mobility beyond current ambient rates. Given the decreases in both total Hg and MeHg, it
appears that all forms of Hg were mobilized and either exported or volatilized by the
warming + S addition treatment, despite decreases in Hg methylation rate.
12.3.3.4. pH and Methylation Potential
The 2008 ISA described the state of knowledge current at the time that lower pH waters
had higher MeHg production. This relationship was assumed to be causal. However, the
laboratory body of evidence suggests that if all other factors are held equal, methylation
is highest at near neutral pH and drops with increasing acidity. The apparent discrepancy
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is due to the fact that SOx deposition causes both elevated sulfate concentrations and
acidification in aquatic ecosystems.
Studies from the early 1980s assumed that acidification, not sulfate concentration, was
the causal factor in observed increases of MeHg in water and increased Hg burdens in
fish in response to SOx deposition. For example, samples from circumneutral lakes at the
ELA in Ontario that were experimentally acidified with H2SO4 had increased methylation
and decreased demethylation, so that methylation:demethylation rates were 2.9 to 4 times
higher at pH 5.1 than at pH 7.1 (Xun et al.. 1987; Ramial et al.. 1985).
Recent research on the role of S reducing bacteria in methylating Hg (Appendix 12.3.2)
suggests reinterpretation of an acid pH-MeHg relationship. Controlled laboratory
experiments showed that it was the added sulfate, not pH, that increased Hg methylation
rates in these lakes. Slurries with sediments from Lake Clara, WI, were made with added
H2SO4 or a molar equivalent amount of sulfate as Na2S04. Acidification from neutral
slurries to pH 4.5 with H2SO4 decreased methylation by 65%, while an equivalent
addition of Na2S04 did not alter methylation rates; acidification to pH 3.0 almost entirely
inhibited methylation (Steffan et al.. 1988). This study also measured demethylation over
a range of pH 2.0-8.0. The results showed that demethylation was higher than 2% from
pH 4.4-8, while methylation rose above 2% from pH 6.0-8.0, and both processes slowed
with decreasing pH (Steffan et al.. 1988). The 2008 ISA correctly identified the
correlation between acidic surface water and high rates of MeHg, but the two factors
share a common cause, which is increased surface water sulfate concentrations, rather
than a direct causal relationship. In the Adirondacks, which have historically experienced
high S deposition (see Appendix 16.2). surface water total Hg, MeHg, and %MeHg were
unrelated to pH when all variables were measured across 44 lakes in 2003-2004 (Yuet
al.. 2011). However, there were effects of pH on Hg in biota (see Appendix 12.4). In
regions like the Adirondacks, which historically received high S deposition and have not
fully recovered from this historical acidification, pH remains an important predictor of
Hg in biota.
12.3.3.5. Organic Matter and Methylation Potential
Organic matter is the metabolic substrate of all the microbial methylators, including both
the group of organisms responsive to sulfate and the microbial methylators that depend
on other metabolic pathways. Organic matter in environments where carbon affects
mercury cycling may be quantified as soil organic matter (SOM; as percentage of soil or
in g OM/g sediment); total organic carbon (TOC, g carbon/mL pore or surface water) or
its components; particulate organic carbon (POC, g C/mL water); and dissolved organic
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carbon (DOC, g C/mL water). The 2008 ISA described some of the complex qualitative
relationships (stimulatory and/or inhibitory) between DOC and rates of Hg methylation.
Organic matter influences the mercury cycle in a number of important ways. Organic
matter affects MeHg production through effects on inorganic Hg ion chemistry and
availability to microbes, and by serving as a metabolic substrate for SRPs. New studies
from lakes, rivers, wetlands, and estuaries have found positive relationships between a
range of lower DOC concentrations and MeHg production or concentrations. However,
the chemical form of DOC or the salinity of the water body can alter DOC and MeHg
relationships.
Certain forms of DOC may serve as a metabolic substrate that increases SRP activity, and
by extension, Hg methylation. In wetlands in the Arbutus Lake Watershed, NY, there
were strong correlations (r2 = 0.58 or 0.60) between Hg and total carbon in the upper
30 cm of peat (Selvendiran et al.. 2008b'). In northern Wisconsin lakes, the quantity of
DOC derived from wetlands is the strongest (>80% variance) predictor of MeHg (Watras
et al.. 2006). In Canadian boreal lakes, there was a positive relationship between
methylation rates and DOC (r2 = 0.62). In Little Rock Lake, WI, DOC and S reduction
had a complex relationship: in the range of 3.6-7.6 mg C/L (reported as
300-600 (imol/L), sulfide (a proxy for S reduction) increased linearly with DOC
concentration, while S reduction did not occur at DOC less than 3.6 mg C/L, and S
reduction did not increase at DOC concentrations above 7.6 mg C/L (Watras et al.. 2006).
In Lake Geneva, Switzerland, mercury methylation rates were an order of magnitude
higher in settling particles than in sediments, and analysis of the particles suggested that
they were derived from algal biomass (Gascon Diez et al.. 2016).
Sediment organic matter may also increase microbial Hg methylation. At another
watershed in the Adirondacks, Sunday Lake, NY, the highest MeHg concentrations were
in samples that were more than 20% organic material (Yu et al.. 2010). In prairie potholes
in Saskatchewan, surface water [MeHg] increased in a linear relationship with the
sediment percentage of organic matter (OM), with a 2.65 ng/L increase in MeHg for
every 10% increase in the percentage of OM (Hoggarth et al.. 2015).
In wetlands, rivers, and lakes of the Mississippi River Delta, LA, total Hg dissolved in
surface water was positively correlated with DOC concentrations, but negatively
correlated with sulfate concentrations. However, this correlation is confounded by
ecosystem salinity, because mean sulfate concentrations in surface water were 16 mg/L in
freshwater wetlands, 63 mg/L in brackish wetlands, and 909 mg/L in marine wetlands
(Hall et al.. 2008). Also, dissolved MeHg in surface water increased linearly with the
hydrophobic organic acid fraction of DOC, with dissolved MeHg increasing 1 ng/L
(0.000001 mg/L) for each increase of 0.048 mg organic acid/L (Hall et al.. 2008). The
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hydrophobic organic acid fraction represents a recalcitrant source of aromatic carbon
such as that found in peat, which would account for the high amounts of MeHg in these
wetlands. The hydrophobic organic acid fraction of terrestrially derived C also controls
Hg export (see paragraph below). Another recent study used controlled incubations of an
SRP strain, varying concentrations of Hg, and two well-characterized forms of DOC to
quantify the effects of DOC on methylation. DOC typical of terrestrial sources (high
aromatic fraction, high molecular weights) increased MeHg production 10-40 times
above unamended controls, while aquatic DOC (high aliphatic fraction, low molecular
weights) doubled MeHg production (Graham et al.. 2012). The authors hypothesize that
DOC increases MeHg production by slowing the growth of Hg-S particles, increasing
Hg-S residence time in the water, and enhancing Hg bioavailability (Graham et al..
2012). The stronger effects of terrestrial DOC on MeHg may explain why MeHg
production is particularly high in wetlands, recently flooded reservoirs, and periodically
flooded river plains (Benoit et al.. 2003). In estuarine and marine sediments of the
Chesapeake Bay, MD, organic C in the sediments had a strong positive correlation
(r = 0.96) with sediment MeHg concentrations (Hollweg et al.. 2010). In coastal marshes
of the Chesapeake such as Kirkpatrick Marsh, MD, there was also a statistically
significant positive relationship (r2 = 0.478) between Hg methylation rate and
mineralization of organic C, a metric of cellular metabolism within the sediment
(Mitchell and Gilmour. 2008).
Organic matter in surface water can increase the amount of Hg transported throughout a
watershed, as well as MeHg production at the watershed scale. DOC can form complexes
with Hg ions that facilitate their transport. In lakes and rivers of the St. Louis River, DOC
and dissolved Hg were correlated (r2=0.69), as were aromatic DOC and dissolved Hg
(r2 = 0.70), indicating that aromatic DOC transports Hg through aquatic ecosystems in
this watershed (Jcrcmiason et al.. 2016). In riparian forest soils in Oak Ridge, TN,
flooding resulted in a pulse of both aromatic DOC and MeHg in the soil solution (Poulin
et al.. 2016). In the Arbutus Lake watershed in the Adirondack Mountains, NY, the fluxes
of Hg and DOC in streams were correlated (r2 = 0.80) across seasons, and the highest
fluxes were in June and July (Selvendiran et al.. 2008a). In a study of five watersheds
across the U.S., MeHg concentrations in catchment outflows correlated positively with
particulate organic matter (POC) in Sleepers River, VT (r2 = 0.93), and in Rio Icacos,
Puerto Rico [r2 = 0.83 (Shanlev et al.. 2008)1. In three Northeast watersheds, stream
export of dissolved Hg correlated positively with DOC, with 0.34 ng Hg/L increase for
each 1 mg C/L increase (r2 = 0.87), and hydrophobic organic acid fraction of DOC
explained 91% of variation in dissolved Hg (Dittman et al.. 2010). Storm events, which
increased water flow through surface soil layers and stream discharger, also increased the
export of organic matter and associated Hg (Dittman et al.. 2010). Thus, the evidence
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indicates that the amount and form of organic matter will affect Hg transport in
watersheds.
DOC affects the partitioning of inorganic Hg between sediment and water, and hence the
bioavailability of inorganic Hg to methylating organisms (Hsu-Kim et al.. 2013; Marvin-
Dipasquale et al.. 2009; Benoit et al.. 2003). In Kirkpatrick Marsh on the Chesapeake
Bay, MD, linear relationships between methylation rates and absorbance values indicated
that Hg methylation rates were higher when the DOC was composed of high molecular
weight, aromatic organic compounds (Mitchell and Gilmour. 2008). There were no
correlations between DOC characterization and measures of microbial activity, so the
authors suggested that the form of DOC (size and aromaticity of compounds) controls Hg
bioavailability rather than serving as a metabolic substrate powering MeHg production
(Mitchell and Gilmour. 2008). However, the relationship between DOC and Hg
partitioning is complex [reviewed in Hsu-Kim et al. (2013)1. and in an extensive survey
of the Alabama River basin around Mobile, AL, there was no relationship across
52 sampling sites between DOC and Hg fractions [total, aqueous, particulate (Warneret
al.. 2005)1.
12.3.3.6. Iron and Methylation Potential
The 2008 ISA briefly identified iron (Fe) as a surface water component that can alter the
relationship between SOx deposition and MeHg production (as it also interacts with
sulfide phytotoxicity and internal eutrophication, see Appendix 12.2.3 and
Appendix 12.2.4). Iron can alter the rates of Hg methylation via abiotic and biotic
mechanisms. Iron binds with sulfide produced by SRPs during methylation to precipitate
out of solution (Hellal et al.. 2015). preventing the negative feedback of sulfide on sulfate
reduction. Iron oxides form surface complexes with inorganic Hg that decrease Hg
bioavailability. However, sulfide released by SRPs can react with Fe-Hg complexes to
release Hg into the water column (Hellal et al.. 2015). Iron can increase methylation rates
directly, by stimulating the activity of iron-reducing bacteria capable of Hg methylation,
including Geobacter spp. In laboratory incubations mimicking freshwater sediments, iron
reduction and sulfate reduction co-occurred in time but at different zones in the sediment
(Hellal et al.. 2015). Iron and S reduction also co-occurred in wetlands of the Yolo
Bypass, CA (see Appendix 12.3.4.3).
In Kirkpatrick Marsh on the Chesapeake Bay, MD, there was a positive relationship
(r2 = 0.478) between Hg methylation rate and mineralization of organic C and an equally
strong positive relationship between Hg methylation rate and the pool of Fe2+ [r2 = 0.478
(Mitchell and Gilmour. 2008)1. Although Fe precipitation with Hg-S has been posited as
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one of the mechanisms by which Hg is buried in sediments, there was no relationship
between iron concentration and partitioning of Hg between water and soil.
12.3.3.7. Nitrate and Methylation Potential
The 2008 ISA did not address the effect of nitrate upon MeHg production. Nitrate ions
are energetically favored over sulfate in redox reactions, so N addition can depress Hg
methylation by shifting the energetic advantage from SRPs to denitrifiers (Dcv ct al..
2015). In Florida Everglades sediment incubations, addition of 1.4 mg N/L as NOs to
incubations decreased MeHg production by 70% (Gilmour et al.. 1998). In Onondaga
Lake, NY, the Syracuse water treatment plant incorporated year-round nitrification in
2004, doubling the nitrate concentrations measured annually in May before the lake
stratified (Todorova et al.. 2009). MeHg concentrations decreased exponentially with
increasing NO3 concentrations (estimated pseudo-r2 = 0.61) in the 2-3 years following
the change in water treatment. This change in N load reduced MeHg accumulation in the
hypolimnion by 50% in 2006, and by 93% in 2007 compared to MeHg measured earlier,
in 1982, 1992, and 2002 (Todorova et al.. 2009). In summer 2011, a calcium nitrate
solution was pumped directly into the hypolminion of Onondaga Lake thrice weekly,
adding 84 MT of N over the course of the season (Matthews et al.. 2013). As a result,
hypolimnion N-NO3 remained above 1.0 mg/L during summer stratification, and
hypolimnion maximum MeHg concentration decreased 94% compared to 2009.
(Matthews et al.. 2013) attributed this to two effects of N: a shift in competitive
advantage from SRP to denitrifying bacteria, and increased sorption of MeHg to lake
sediments.
12.3.3.8. Salinity and Methylation Potential
Mercury methylation occurs in estuarine and marine ecosystems (Lchnhcrr et al.. 2011;
Benoitetal.. 1998). but sulfate concentrations in saline water are generally so high that
direct deposition of SOx is unlikely to alter existing MeHg dynamics. An early study on
coastal marshes in Georgia showed that both sulfate reduction and MeHg production
were at their highest rates in the top 4 cm of the marsh soil (King et al.. 1999). An older
study of Hg dynamics in the Patuxent River Estuary, MD, where sulfate concentrations
range from 150 to 1,200 mg/L, found no correlation between sediment pore water sulfate
concentrations and MeHg fraction in sediments (Benoitetal.. 1998). There was a peak in
surface water MeHg fraction of 25% at the mouth of the river, which authors speculated
was due to mixing by wind of stratified anoxic water from the Chesapeake Bay with the
river water (Benoitetal.. 1998). MeHg sediment fraction across the Patuxent River, MD,
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was negatively correlated with sulfide sediment concentrations, presumably because Hg
bound to sulfide is unavailable to methylating bacteria (Benoit et al.. 1998).
12.3.4. Mercury Methylation in Sulfur Addition Field Studies
There are a number of ecosystems in which sulfate is elevated by anthropogenic
activities. Scientists have added sulfate to ecosystems in order to study S and Hg cycling
(see Appendix 12.3.4. IV S is used as an agricultural amendment in peat soils, and so Hg
cycling has been extensively studied in the Everglades Water Conservation Area (see
Appendix 12.3.4.2) and the San Joaquin Delta (see Appendix 12.3.4.3V
12.3.4.1. Ecosystem Scale and Field Mesocosm Studies
The 2008 ISA found that experimental S addition to aquatic and wetland ecosystems
increased MeHg concentrations in surface water (and Hg concentrations in biota, see
Appendix 12.4). Key studies supporting this finding focused on the long-term S
acidification (1984-1989) and recovery (1990-present) experiment at Little Rock Lake,
WI and the relatively recent (initiated in 2001) experimental S addition experiment at
Bog Lake Fen in the Marcell Experimental Forest, MN. This section integrates older
published reports and recent papers from S addition experiments that presented
quantitative relationships between S addition and Hg mobilization, methylation, and
MeHg concentrations in surface water. There is new evidence from the S addition
experiment at Bog Lake Fen and from an S addition by warming experiment in a Swedish
bog that confirms the relationship reported in the 2008 ISA between increased S loading
and environmental increases in MeHg. There is also new evidence that decreases in S
addition allow recovery of MeHg toward control conditions, as well as new evidence that
climate change (drought, warming) will increase S cycling and increase MeHg
concentrations (Table 12-7).
As reported in the 2008 ISA, experimental S addition increases MeHg production in lake
ecosystems, and cessation of S addition results in reductions in MeHg production in
lakes. Little Rock Lake, WI was an S addition experiment that ran from 1985-1991. In
this experiment, the north and south basins of the lake were artificially isolated by a
plastic barrier, and then annual S additions were used to acidify the northern basin, while
the unamended south basin served as a control. In 1993, after S addition had ceased but
sulfate concentrations in the north treatment basin (4.8 mg/L or 50 (j,M) were still twice
that of the reference basin (2.4 mg/L or 25 (j,M), the potential methylation rate was 220%
higher in the treatment basin (Gilmour and Riedel. 1995). In June 1993, when MeHg
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fractions were at their highest, the MeHg fraction of total Hg was 30% higher in the
treatment basin (Gilmour and Riedel. 1995). Mercury and S cycling were monitored in
the lake through 2003, and elevated sulfate concentrations and acidification persisted in
the lake through 1996; the lake was considered recovered from 1998-2003. During
acidified years, 25% more of annual Hg2+ inputs were transformed into water column
MeHg than in recovery years ("Watras et al.. 2006). Net methylation during the summer
was 70% higher in acidified than in recovery years, and summertime MeHg accumulation
was 75% higher in the whole lake in acidified years (Watras et al.. 2006). There was a
strong positive correlation between summer MeHg accumulation (1990-2003) and the
interaction of epilimnion sulfate concentration and epilimnion Hg2+ concentration. The
hypolimnion or the anoxic-oxic boundary were likely the sites of Hg methylation,
because the hypolimnion, which was less than 5% of the lake volume, accumulated more
than 70% of the total MeHg mass of the lake in acidified years (Watras et al.. 2006).
Hypolimnion MeHg was associated with hypolimnion S reduction, as there was a
10.8 ng MeHg/L (0.0000108 mg MeHg/L) increase for each 1 mg/L increase in hydrogen
sulfide (H2S; reported as 1.71 pmol/L increase in MeHg for each 1 (imol/L increase in
H2S; r2 = 0.65). Total MeHg in Little Rock Lake, WI increased 14.4 mg for every 1 kg/ha
increase in sulfate deposition (Watras and Morrison. 2008).
Table 12-7 New studies on mercury methylation in sulfur-amended ecosystems.
Type of Additions or Load
Ecosystem (kg S/ha/yr)
Biological and Chemical Effects Study Site
Study
Species Reference
Peat fen Ambient S
deposition:
3 kg S/ha/yr
S addition: low S
(7 kg S/ha/yr
added for total load
of 10 kg S/ha/yr) or
high S
(17 kg S/ha/yr
added for total load
of 20 kg S/ha/yr)
as Na2SC>4, with 0
or 30 kg N/ha/yr as
NH4NO3
At 30-cm depth, total Hg (per peat
mass) in warming + high S was 19%
lower than in high S plots.
Low S decreased total Hg in peat
methylation hotspots by 25 or 31% on a
per-mass or per-volume basis.
The Hg methylation rate constant
increased by 71% in low S and by 400%
in high S.
Peat MeHg concentrations decreased
by 22% in low S and increased by 74%
in high S.
Degero
Stormyr,
Sweden (64°
¦ 09' N, 20°
22' E),
measured in
2007 and
,2008
Peat
samples
Akerblom
et al.
(2013)
Peat fen
Ambient S
deposition:
3 kg S/ha/yr
S addition: high S
(17 kg S/ha/yr
S addition increased pore water MeHg	Degero
concentrations by 117%. S addition	Stormyr,
increased the range of MeHg	Sweden (64°
concentrations by 4.3x, and the	11' N, 19°
variance by 22.2x the control variance.	33' E),
Pore water
in peat
Bergman et
al. (2012)
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Table 12-7 (Continued): New studies on mercury methylation in sulfur-amended
ecosystems.
Type of Additions or Load	Study
Ecosystem (kg S/ha/yr)	Biological and Chemical Effects Study Site Species Reference
added for total load With s additioni MeHg increased with measured in
of 20 kg S/ha/yr) higher groundwater levels, although ^
there was no relationship in ambient
plots.
Total Hg in the top 5 cm of peat
increased by 133% with S addition.
Peatland
(poor fen)
Experimental
treatment, addition
of 32 kg/ha/yr in
three simulated
rainfall events,
2001-2008
Recovery
treatment,
32 kg/ha/yr
2001-2006, no
added S
2006-2008
Deposition = 5.5 kg
S/ha/yr (NADP site
MN16)
Mesocosms
installed in
ambient, added S,
and recovery
zones of marsh,
pulse of 130 mg
Na2SC>4 added
In spring 2005, S addition increased
%MeHg 387% over control wetland. In
fall 2005, S addition increased %MeHg
275%.
Following a 9-mo drought, spring 2007
rewetting released internally stored S,
the pulse of SO42" was 78% higher in
recovery and 147% higher in
experimental than in control wetlands.
Following a 9-mo drought and rewetting
in spring 2007, peak [MeHg] was 75%
higher in recovery and 120% in
experimental wetland. Peak %MeHg
was 60% higher in recovery and 150%
in experimental wetland.
Following summer-long drought and
rewetting in fall 2007, [SO42"] was 14%
higher in recovery and 165% higher in
experimental wetland.
Following summer-long drought, the fall
2007 peak [MeHg] was 102% higher in
recovery and 301% higher in the
experimental wetland. Peak %MeHg
was 150% higher in recovery and 350%
higher in the experimental wetland.
Following summer-long drought, the
spring 2008 peak [MeHg] was 62%
higher in recovery and 133% higher in
the experimental wetland. Peak %MeHg
was 146% higher in recovery and 262%
higher in the experimental wetland.
Experimental water level rise in
mesocosms does not affect chemical
constituents in control and recovery
wetlands, but increased total Hg, SO42",
MeHg, and %MeHg in the experimental
wetland. DOC decreases in the
experimental wetland mesocosm.
S6 peatland
in Marcell
Experimental
Forest, MN
Sphagnum
spp.,
herbaceous
forbs,
ericaceous
shrubs,
spruce, and
tamarack
Wasik et al.
(2015)
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Table 12-7 (Continued): New studies on mercury methylation in sulfur-amended
ecosystems.
Type of Additions or Load
Ecosystem (kg S/ha/yr)
Biological and Chemical Effects Study Site
Study
Species Reference
Bog	Ambient S load:
2.07 kg S/ha/yr as
quantified by
NADP
S addition:
8.28 kg S/ha/yr
(low), or
20.7 kg S/ha/yr
(high) as SO42"
solution
S addition increased pore water MeHg
concentrations 219-246%.
High S increased MeHg fraction
(MeHg/total Hg) by 117%.
Labile organic C and S additions
increased methylation rates (pg
MeHg/L/day) by 163% and rates of
increasing MeHg fraction (%MeHg/day)
by 152% over rates for S-only addition.
Bog Lake Peat pore Mitchell et
Fen, Marcell water	al. (2008a)
Experimental
Forest, MN
Bog	Ambient S
deposition:
5.5 kg/ha/yr (mid
2000s) as
quantified by
NADP site MN16
S addition:
32 kg S/ha/yr
dissolved in pond
water and
delivered by
sprinklers (mimics
4x the 1990s
deposition rate)
In pore water, MeHg (ng/L) was
increased by 8.8-17.9x the control
levels in 2006 and 3.4-11.7* the control
levels in 2008. MeHg fraction
(MeHg/total Hg) was increased
6.1-13.4x in 2006 and 3.9-11.6x in
2008.
In solid peat, MeHg concentrations and
MeHg fraction were 4-9* higher than in
controls or5-6x higher when
accounting for annual variability.
3 yr after SO42" treatment cessation
(recovery treatment), pore water MeHg
declined 32%, peat MeHg declined
(ng/L) by 62%, and peat MeHg fraction
(%) declined 76%.
S6 peatland, Pore water, Wasik et al.
bog section,
Marcell
Experimental
Forest, MN
peat, and
Culex spp.
(mosquito)
larvae
(2012)
Lakes: rain-
fed LRL,
DL drains
stream fed
by wetland
LRL: [SO42"] =
2.5 mg/L, DL:
[SO42"] = 2.8 mg/L,
stream feeding DL
[SO42"] = 1.6 mg/L
S deposition not
reported
In LRL during drought years, lake SO42"
concentrations increased 0.18 mg/L/yr,
as MeHg concentrations increased
0.004 ng/L/yr.
In stratified basin of LRL, total MeHg
increased with increasing S deposition:
mg MeHg = 14.368 * (kg
S/ha) - 18.494.
In stream draining wetland, SO42"
decreased 80% as MeHg increased
69% over the course of spring and
summer.
Little Rock
Lake (LRL)
and Devils
Lake (DL),
Wl
Lake water
Watras and
Morrison
(2008)
C = carbon; cm = centimeter; DL = Devils Lake; Ha = hectare; Hg = mercury; kg = kilogram; L = liter; LRL = Little Rock Lake;
MeHg = methylmercury; mg = milligram Na2S04 = sodium sulfate; NADP = National Atmospheric Deposition Program;
ng = nanogram; NH4NO3 = ammonium nitrate; S = sulfur; S042" = sulfate; yr = year.
DL = Devils Lake; ha = hectare; kg = kilogram; L = liter; LRL = Little Rock Lake; LTER = Long Term Ecological Research;
MeHg = methylmercury; mg = milligram; ng = nanogram; S = sulfur; S042" = sulfate; yr = year.
Experimental S addition substantially increases MeHg in wetlands. The ELA in Ontario
has been the subject of considerable research in Hg cycling (see Supplemental Material).
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An early experiment added pulses of sulfate to peat inside collars installed in a poor fen,
and monitored changes in MeHg concentrations over 4 days. In one experiment, addition
of 14 kg sulfate/ha increased MeHg pore water concentrations (ng/L) at the water table
surface 30% above the highest measured MeHg concentrations in the control plot, with a
100% increase in MeHg above control plot concentrations at a 10-cm depth in the water
table (Branfircun ct al.. 1999). Following a second 14 kg sulfate/ha addition, MeHg at
5-cm depth in the peat water table increased to 130% above maximum MeHg in the
control plot. Four days after the first S addition, MeHg at the 5-cm depth remained
elevated at 250% above maximum measured MeHg in control plots (Branfircun et al..
1999). In an experiment the following year, one sulfate addition equivalent to
2.8 kg sulfate/ha was added. After 24 hours, this S addition had increased MeHg pore
water concentrations at the water table surface 100% above maximum measured MeHg in
the control plot, and at the 10-cm depth below the water table, MeHg increased 20%.
This MeHg increase did not persist to 48 hours after addition (Branfireun et al.. 1999).
There is new evidence that addition of labile DOC typical of leachate from a forested
watershed enhanced the stimulatory effect of S addition upon MeHg concentrations in
wetlands. At the Marcell Experimental Forest, MN, field mesocosms were initiated in
Bog Lake Fen to test the effects of S and carbon addition on Hg cycling in an
ombrotrophic peat bog. Mesocosms received experimentally added loads of
8.3 kg S/ha/yr or 20.7 kg S/ha/yr, alone or in combination with additions of organic
carbon in different forms (Mitchell et al.. 2008a'). S addition increased peat pore water
concentrations of MeHg 219-246% above MeHg levels in unamended control plots, with
no significant difference between net methylation under 8.3 versus 20.7 kg S/ha addition
(Mitchell et al.. 2008a'). Adding 20.7 kg S/ha/yr to mesocosms increased the fraction of
the total Hg pool that took the form of MeHg, with MeHg 117% higher than in control
plots (Mitchell et al.. 2008a'). The combined addition of sulfate and organic carbon (as
glucose, acetate, or leachate from coniferous leaf litter) increased methylation rates
(reported in pg MeHg/L/day) by 163% and increased the fraction of MeHg (%MeHg) by
152% over rates measured in mesocosms that received only S addition (Mitchell et al..
2008a'). This study showed that S addition will increase MeHg in surface water in bogs,
particularly when combined with labile C inputs. These results coincide with findings
presented elsewhere in the ISA that DOC concentrations are increasing in surface waters
(Appendix 7) as terrestrial ecosystems recover from acidification.
A decade of research at Bog Lake Fen in Minnesota shows that additional S increases
MeHg production in this wetland, with consequences for water quality and also for
resident organisms (see Appendix 12.4). Bog Lake Fen is the site of a long-term S
addition experiment in which entire sections of the Sphagnum peat wetland have received
additional S loading in simulated wet deposition. As reported in the 2008 ISA, S addition
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began in 2001 at a rate of 32 kg S/ha/yr, and increased the levels of MeHg in the bog by
40-190% in the first summer and increased the MeHg export from the bog by 144% in
2002 (Jcrcmiason et al.. 2006). More recent sampling has confirmed the relationship
between S addition and elevated MeHg production and concentrations in the wetland. In
peat pore water sampled 5 cm below the water table for 2 weeks after S addition events, S
addition increased MeHg concentrations by 880-1,790% in 2006 (0.2-0.3 ng MeHg/L in
control, 2.9-4.2 ng/L in S addition plots), and increased MeHg concentrations
340-1,170% in 2008 ("Wasik et al.. 2012). Pore water MeHg fraction (percentage of total
Hg in MeHg form) increased 610-1,340% over control levels in 2006 and 390-1,160%
over control levels in 2008 (Wasik etal.. 2012). Researchers also sampled surface peat
and found that S addition altered MeHg concentrations in peat. MeHg concentrations
(ng/L) and MeHg fraction (percentage of total Hg as MeHg) were 4-9 times higher in S
addition plots than in control plots, although there was no effect of S addition on total Hg
in plots (Wasik et al.. 2012). The elevated MeHg in water and peat had consequences for
the food chain (see Appendix 12.4).
There is also evidence of a link between S addition and MeHg production from a
European wetland experiment. Degero Stormyr in Sweden is the site of an experiment
that addresses the effects of S and N deposition as well as warming due to climate
change. Plots received S loads of 10 kg S/ha/yr or 20 kg S/ha/yr, and a subset of plots
was enclosed by greenhouses to raise air temperature by about 4°C and peat temperature
(at the 20-cm depth) by about 2°C. Four years after the 1995 initiation of the experiment,
Bergman et al. (2012) measured MeHg production in the control and 20-kg-S/ha/yr plots.
S addition increased mean growing season peat pore water MeHg concentrations 117%
over MeHg concentrations in control plots (Bergman et al.. 2012). There was higher
variation in MeHg in S addition plots (20 kg S/ha/yr), as the range of measured
concentrations was 4.3 times the range of measured MeHg in control plots (Bergman et
al.. 2012). The experiment was resampled 12 years after initiation, and the Hg
methylation constant was 71% higher in methylation hotspots (a hotspot was
experimentally defined as the point in each sampled peat profile with the highest Hg
methylation rate constant) under addition of 10 kg S/ha/yr compared to control plots,
while MeHg concentrations were 22% lower than in control hotspots (Akerblom et al..
2013). These seemingly contradictory results (higher production, lower concentration)
may indicate increased mobilization of MeHg from methylation hotspots (see below) or
increased microbial demethylation rates, which were not tested. Under higher S addition
of 20 kg S/ha/yr, the Hg methylation rate constant was 400% higher than in the control
plots, while the MeHg concentrations were 74% higher than in control plots (Akerblom et
al.. 2013). Together, these results indicate that in wetlands an S addition of 10 or
20 kg/ha/yr increases microbial MeHg production and that an S addition of 20 kg/ha/yr
increases MeHg concentrations.
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S addition also increases the mobilization and transport of Hg and MeHg from peat
methylation hotspots into the water column, increasing bioavailability. After 4 years of S
addition at Degero Stormyr in Sweden, plots that received 20 kg S/ha/yr had 133% higher
total Hg levels in the top 5 cm of peat than control plots (Bergman et al.. 2012). After
12 years, sampling of Hg methylation hotspots in plots with 10-kg-S/ha/yr addition found
total Hg was 25% lower per peat mass or 31% per peat volume than in control plots
(Akerblom et al.. 2013). These results indicate that S addition mobilizes Hg stored in
deeper parts of the peat profile, resulting in Hg transport to more accessible parts of the
peat profile.
Recent work from the experiment at Bog Lake Fen also provides evidence that reduction
of experimental S addition to wetlands results in reductions of MeHg concentrations in
wetlands. In 2006, after 5 years of elevated S levels, a recovery treatment was initiated in
the S6 bog experiment, where S addition was halted in recovery plots. Following the
cessation of S addition, pore water MeHg (ng/L) declined by 32% between 2006 and
2008, and peat MeHg concentration and peat MeHg fraction declined by 62 and 76%,
respectively ("Wasik et al.. 2012).
12.3.4.2. Agricultural Systems: The Everglades Water Conservation Area, Florida
There are multiple new studies that relate Hg concentrations in water, periphyton, soil, or
fish to sulfate concentrations in the Everglades Water Conservation Area (see
Table 12-8). In the Water Conservation Areas outside Everglades National Park, FL
(Class I area), sulfate concentrations were as high as 48 mg/L near the EAA, whereas in
pristine areas distant from the EAA, sulfate concentrations were less than 0.5 mg/L.
Previous studies noted that Hg methylation was highest in areas of the wetland where
sulfate concentrations ranged from 0.7 to 1.9 mg sulfate/L (Bates et al.. 2002; Gilmour et
al.. 1998). More recently, a study by Corrales et al. (2011). quantified the S budget and
determined that atmospheric deposition of S is 15 kg S/ha/yr, 4% of the total S load in the
WCAs. MeHg concentrations in surface water in the Florida Everglades increase when
sulfate concentrations are above 1 mg/L, which Corrales et al. (2011) recommended as a
target level for sulfate reduction efforts (Figure 12-10).
Table 12-8 New studies on mercury (Hg) and sulfur (S) cycling in the Everglades
Water Conservation Areas (WCA).
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Table 12-8 (Continued): New studies on mercury (Hg) and sulfur (S) cycling in the
Everglades Water Conservation Areas (WCA).
Additions	Biological
Type of or Load	and
Ecosys (kg S/ha/y	Chemical Study Study
tem r)	Effects Site Species Reference
Freshw
ater
marsh
W3 pore
water
[S042"]
<4 |jM
U3 pore
water
[S042"]
39 |jM
F4 pore
water
[S042"]
74 |jM
Deposition
= not
measured
I is
I is
SRB
abundance
was 6.9x
higher in U3
than W3,
16.8x higher
in F4.
In W3,
abundance of
SRB and
methanogens
are not
significantly
different. In
U3, SRB
abundance is
80% higher
than
methanogen
abundance. In
F4, SRB
abundance is
60% higher
than
methanogens.
mcrA copy
number
correlates
positively with
mRNA and
methane
production
rates.
Acetotrophic
methanogens
are dominant
at W3, while
at high S U3
and F4 sites,
hydrogenotrop
hie
methanogens
are dominant.
F4, U3,
and W3
sites in
Water
Conserv
ation
Area,
Everglad
es, FL
Methanog
ens as
quantified
by mcrA
copies
Sulfur
reducing
bacteria
as
quantified
by dsrB
copies
Bae etal. (2015)
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Table 12-8 (Continued): New studies on mercury (Hg) and sulfur (S) cycling in the
Everglades Water Conservation Areas (WCA).

Additions
Biological


Type of
or Load
and


Ecosys
(kg S/ha/y
Chemical
Study
Study
tern
r)
Effects
Site
Species Reference
Wetland
Total S
MeHg surface
Everglad
Surface Corrales et al. (2011)

load is
water
es
and

110,303
concentration
agricultu
groundwa

mtons/yr,
rises above
ral area
ter

atm dep is
suggested
(EAA),


4% of
threshold level
FL


total, or
of 1 mg/L



15 kg
sulfate in



S/ha/yr as
groundwater.



measured




by US




EPA




CASNET




(FL11 and




FL99)



Constru
Deposition
Fish Hg
Western
Gambusi Fena et al. (2014)
cted
not
(Y= log mg
Palm
a
wetland
reported
Hg/kg G.
Beach
holbrooki
s
S load as
holbrooki) was
County,
(eastern

[S042"] in
negatively
FL
mosquitof

wetland
correlated

ish) and

inflow:
with water

surface

39-110 m
sulfate

water

g/L
(X = log mg



S042"/L)
concentration
s in three
treatment


wetlands:
Y	= -2.09X +
2.60;
Y	= -3.17X
+3.80;
Y= -1.09X
-0.01.
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Table 12-8 (Continued): New studies on mercury (Hg) and sulfur (S) cycling in the
Everglades Water Conservation Areas (WCA).
Additions	Biological
Type of or Load	and
Ecosys (kg S/ha/y	Chemical Study Study
tem r)	Effects Site Species Reference
Wetland Deposition Fish Hg (mg Everglad Gambusi Gabriel et al. (2014)
s and
canals
not
reported
S load as
[SO42"] in
surface
water:
0-60 mg/L
Hg/kg or mg
Hg/kg/mm for
Lepomis)
increased
over 0-1 mg/L
SO42",
reached peak
levels at
1-12 mg/L
SO42", and
decreased
over
12-25 mg/L
SO42", to
remain flat at
SO42"
concentration
s 25-60 mg/L.
es
protectio
n area,
FL
a spp.
(mosquito
fish),
Lepomis
spp.
(sunfish),
and
Micropter
us
salmoide
s
(largemo
uth bass)
Constru
cted
wetland
s
Deposition
not
reported
S load in
surface
water as
[S042"]:
39.3-110.
1 mg/L
(mean
59.5 mg/L)
Between
2000-2004,
total Hg and
MeHg wetland
export were
negatively
correlated to
inflow [SO42"].
tHg: (log ng
tHg/L) =
-1.22*(log mg
S0427L) + 2.4
0; MeHg: (log
ng
MeHg/L) = -1.
91*(log
[S042"])
+3.02.
Between
2000-2011,
sulfate was
negatively
correlated
with tHg and
MeHg as one
variable in
multivariate
models.
Western
Palm
Beach
County,
FL
Surface Zheng et al.
water (2013)http://hero.ero.gov/index.cfm?action=search.vie
w&reference id=2044818 - ENREF 2954
atm = atmospheric; CASTNET = Clean Air Status and Trends Network; dep = deposition; EEA = Everglades agricultural area;
ha = hectare; Hg = mercury; kg = kilogram; L = liter; MeHg = methylmercury; mg = milligram; ng = nanogram; S = sulfur;
S042" = sulfate; THg = total mercury; yr = year.
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6
cn
c
o
(C
at
o
c
o
o
3
o
a_
o
E
4-1
3
2
Threshold

~ ~
a-	4	*	~
0	** I
^ 1	* * *1^ ~~ * #~ «*
1	0 Jfffe
2	4	6	8
Sulfate concentration (mg L-i)
10
L = liter; mg = milligram; ng = nanogram.
Source: adapted from Corrales et al. (2011).
Figure 12-10 The relationship between surface water sulfate and
methylmercury concentrations in the Florida Everglades.
Orem et al. (2011) suggested that at high surface water sulfate (>20 mg/L) concentrations
in the Florida Everglades, sulfide accumulates and inhibits methylation. In treatment
wetlands that collect stormwater runoff from the EAA, the concentrations of sulfate are
so high that sulfide inhibition of MeHg may occur at the higher end of the sulfate range.
Zheng et al. (2013) reported on the water quality parameters, total Hg, and MeHg in the
inflow and outflow waters of a constructed treatment wetland in the Everglades WCAs.
The wetland was a net source of MeHg to surface water between 2000 and 2004, and
during this time, total Hg and MeHg concentrations in water exiting the wetland were
negatively correlated with sulfate concentrations of water flowing into the wetland
(Zheng et al.. 2013). The range of sulfate concentrations of inflow was 39-110 mg/L
sulfate, with mean sulfate of 59.5 mg/L. Zheng et al. (2013) also constructed multivariate
models for total Hg and MeHg, and even with other factors such as DOC, pH, and
chloride in the model, inflow sulfate remained a strong and negative correlate of outflow
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Hg concentrations. The positive relationship between sulfate and MeHg in the Florida
Everglades is significant only at low sulfate concentrations (in this study,
39 mg S0427L).
A gradient of nutrient eutrophication (N, P, and S) from agricultural runoff runs from
northeast to southwest across the WCA, with sulfate concentrations around 4.8 mg/L
(reported as 50 |iM) in southern, less disturbed sites (Gilmour et al.. 1998). Sediment Hg
increased from north to south by a factor of 3-4, and MeHg sediment concentrations
increased across the same gradient by a factor of 25 (Gilmour et al.. 1998). A more recent
study in the Everglades suggested that mercury hotspots (elevated Hg in mosquitofish)
were correlated with relatively oligotrophic conditions (low soil total phosphorus, TP)
and native sawgrass sloughs, while fish Hg was lower in sites with higher soil TP with
invasive cattail stands (Julian et al.. 2016a). Surface water sulfate concentrations
measured at the same sites ranged from 0.2-35 mg SC>427L in the Hg hotspots and
6.5-70 mg SC>427L at the nonhotspots (Julian et al.. 2016b). Surface water sulfate and
pore water sulfate were correlated, and the ratio of pore water sulfide:pore water sulfate
correlated with pore water DOC (Julian et al.. 2016b). suggesting that dissimilatory
sulfate reduction (and by extension, a portion of mercury methylation) is controlled by
DOC as well as sulfate in these marshes. In addition to the studies described in this
section, studies in the Everglades WCA have focused on the microbial mechanisms of
mercury methylation (see Appendix 12.3.2) and Hg loads in wildlife (see
Appendix 12.4).
12.3.4.3. Agricultural Systems: Rice in the San Joaquin Delta, California
There is evidence from the California Central Valley that microbial Hg methylation
occurs in managed agricultural wetlands that produce rice, which suggests a route of
MeHg exposure to humans and wildlife. California (particularly the Central Valley where
these studies were conducted) is the second largest producer of rice in the U.S., as well as
an important stop along a major bird migratory route. The USGS coordinated an intensive
research effort on Hg cycling and methylation of Hg in wetlands of the Central Valley of
California, focusing on the Yolo Bypass Wildlife Area, located in the northwestern
region of the San Francisco Bay delta where there are permanent wetlands dominated by
Typha spp. (cattails) and Scirpus spp. (tule) and also agricultural wetlands in which
Oryza sativa (white rice) and Zizania palustris (wild rice) are cultivated. Rice fields go
through annual drying and rewetting cycles. They are flooded in the summer, drained for
fall harvesting, and flooded in the winter to speed decomposition of rice straw and
organic residue.
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In the Yolo Bypass, CA, permanent and agricultural wetlands were sampled over the
course of a growing season, June 2007-April 2008. About 10% of the 105 water samples
taken over that time period exceeded the U.S. EPA Hg water quality criterion of
50 ng total Hg/L, and total Hg was higher in the center and outflow of agricultural
wetlands than in permanent wetlands (Alpcrs ct al.. 2014).
Across the wetlands, all of the water samples taken for MeHg determination exceeded the
San Joaquin Delta regulatory total maximum daily load (TMDL) limit of
0.06 ng MeHg/L. In fact, MeHg concentrations in all water samples exceeded 1 ng/L,
with maximum measured water MeHg of 37 ng/L at the outlet of wild rice fields during
harvest. Water MeHg concentrations were significantly higher, about twice as high, in the
agricultural fields as in the permanent wetlands (Alpcrs et al.. 2014). and sediment MeHg
concentrations were also higher in the agricultural wetlands (Marvin- D i pasq ualc et al..
2014). The water MeHg fraction ranged from 1-80%, with a median value of 6.4%, and
there were seasonal variations in agricultural but not permanent wetlands. In agricultural
wetlands, the MeHg fraction increased 20 times between June and August in fields where
white or wild rice was grown, and the MeHg fraction increased 5 times in fallow fields
over the same time period (Alpcrs et al.. 2014).
Hg methylation in the permanent and agricultural wetlands of the Yolo Bypass, CA
correlated with S, Fe, and manganese (Mn) reduction. The strongest correlations of
MeHg fraction in the wetlands were with metrics of Mn reduction, not reduction of Fe or
S (Alpcrs et al.. 2014). Sulfate was not a limiting factor of Hg methylation in the
wetlands, although S34S and sulfate:CI in outlet waters indicated that sulfate reduction
occurred in the wetlands, with at least 20% of the sulfate in the water column reduced
(Alpcrs et al.. 2014). In agricultural wetland sediments, mineralization of C by iron
reduction was 2.4 times the mineralization of C by sulfate reduction, indicating that Fe
reduction was more important than S reduction in terms of anaerobic decomposition
(Marvin-Dipasquale et al.. 2014). Furthermore, addition of sulfate fertilizer in rice fields
did not raise measured sulfate reduction rates to the higher levels observed in fallow,
unfertilized fields, indicating that sulfate was not a limiting factor on anaerobic microbial
activity in the agricultural wetlands (Alpcrs et al.. 2014; Marvin-Dipasquale et al.. 2014).
In the Yolo Bypass, CA, seasonal wetland export of MeHg occurs during the winter as
well as the summer. In the summer, newly methylated MeHg is drawn into the root zone
of the sediment by plant transpiration, and in winter, diffusion releases MeHg from the
sediment into the water column (Bachand et al.. 2014). The short-term storage of MeHg
in wetland soils decreases MeHg exports from the wetland, but increases exposure of
wetland fish to MeHg (Windham-Mvcrs et al.. 2014b). Also, MeHg in rice seeds were
correlated with root MeHg (r = 0.90), suggesting that MeHg from the root zone may be
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transported through the plant to tissues that are consumed by migrating waterfowl
(Windham-Myers et al.. 2014a'). Research in other systems has shown that MeHg
produced in sediments is transported to rice grains (Paranjapc et al.. 2017). In a study of
rice paddies surrounding industrial emitters of SOx and Hg in Hunan Province, China
(Xu et al.. 2017a). MeHg and %MeHg in rice grains were positively correlated with soil
S, soil MeHg, and atmospheric concentrations of gaseous elemental Hg. MeHg in rice
fields and in rice grains may be an important pathway to human and wildlife Hg exposure
via rice consumption.
Evidence from fish Hg levels in different wetland types suggests that fluctuating water
levels in the agricultural fields, as well as high levels of organic carbon from winter
decomposition of rice straw, raise Hg methylation rates in the rice fields above rates in
the permanent wetlands. Mercury concentrations of penned Gambusia affinis
(mosquitofish) rose 12 times over initial concentration in white rice fields, 5.8 times in
wild rice fields, and 2.9 over initial fish Hg concentration in the permanent wetlands
(Ackcrman and Eagles-Smith. 2010). and these trends were similar to Hg concentration
trends in wild mosquitofish and Menidia audens (Mississippi silversides) caught in the
wetlands. However, trends of Hg concentrations in macroinvertebrates did not follow
trends in fish. There was no difference by wetland type (permanent, white rice, wild rice)
in Hg concentrations of collected water boatmen, Corisella spp., which feed on plant and
algal biomass. Hg concentrations in collected backswimmers, Notonecta spp., which are
predatory, were higher in collections from permanent wetlands than in collections from
cultivated rice wetlands (Ackcrman et al.. 2010).
Drivers of Mercury Methylation under Ambient Conditions
The 2008 ISA presented evidence that elevated sulfate concentrations increased MeHg
concentrations in aquatic and wetland ecosystems. At the time, most studies documenting
this relationship were conducted in lakes and wetlands, and sulfate was often added
experimentally to test the mechanisms of the relationship. Since then, a number of new
studies have shown that increases in ambient sulfate concentrations can produce
increased MeHg concentrations in surface water and sediments in lakes and wetlands (as
well as increased Hg in wildlife, see Appendix 12.4). In addition, there is new evidence
that increased sulfate in rivers, streams, and watersheds increase Hg methylation. Not all
studies include information about the form or relative contribution of S sources in these
ecosystems, but the studies provide evidence that current concentrations of surface water
sulfate contribute to Hg methylation in a broad range of American aquatic and wetland
ecosystems.
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The 2008 ISA described watersheds with low pH waters and high wetland cover as
sensitive to Hg methylation, but did not describe MeHg fraction in the environment as an
indicator of Hg methylation. MeHg is bioavailable and bioaccumulates within organisms
due to the strong complexes it forms with thiosulfate and sulfhydryl groups in organic
molecules. The fraction of total Hg that is in the form of MeHg (or %MeHg) is used as an
indicator of bioavailable Hg within an ecosystem. New studies (as well as key older
studies) of MeHg fractions in watersheds, streams, lakes, and wetlands show the link
between S deposition and MeHg fraction in water and sediments, and confirm the
importance of wetlands and nonurban watersheds as sources of MeHg to aquatic
ecosystems.
12.3.5.1. Sulfur in Ambient Conditions
Aquatic and wetland ecosystems which have received high loads of S generally have
higher MeHg fractions. In sampling of Little Rock Lake and four nearby lakes in
Wisconsin, MeHg fractions in lake water ranged from 6-25% MeHg, with no
relationship to pH (Bloom et al.. 1991). In the Adirondacks, which have historically
experienced high S deposition (see case study), surface water total Hg, MeHg, and
%MeHg were unrelated to pH when all variables were measured across 44 lakes in
2003-2004. Surface water MeHg ranged from 0-48% MeHg, with a mean of 6% (Yuet
al.. 201IV
At the time of the 2008 ISA, only one paper had considered the links between measures
of SOx deposition and MeHg concentrations in ecosystems. Since then, two additional
papers have been published documenting the effects of SOx deposition on Hg levels in
water or fish (Table 12-9). The large scale and scope of these studies include
confounding and interacting environmental factors, such as correlated Hg deposition and
variations in wetland cover, the full effects of which researchers are unable to fully
quantify. Despite this limitation, each of these papers show evidence of temporal or
spatial correlations between SOx deposition and MeHg concentrations in fish or water.
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Table 12-9 New studies on correlations between sulfur (S) and methylmercury
(MeHg) in ecosystems.
Type of
Ecosystem
Additions or Load
(kg S/ha/yr)
Biological and
Chemical Effects
Study Site
Study
Species
Reference
River and reservoir
S addition or
deposition not
reported
[SO42"] in Rio
Grande: pore water:
(means)
74-230 mg/L; deep
water: (range)
152-342 mg/L
SO42"] in Devils
River: pore water:
(means)
74-230 mg/L; deep
water: (range)
15-168 mg/L
Sediment MeHg
(ng/g) correlated
negatively with pore
water sulfate
concentrations.
RG had 40% higher
THg than DR.
DR sediment MeHg
was 65% higher than
in RG sediment, and
DR sediment %MeHg
was 65% than in RG
sediment.
Rio Grande
(RG) and
Devils River
(DR) branches
of Amistad
International
Reservoir, TX
Sediment,
pore water,
and deep
water;
Micropterus
salmoides
(largemouth
bass)
Becker et al.
(2011)
Prairie potholes
(shallow freshwater
marsh)
Lowland ponds,
average [SO42"]
14 mg/L
Upland ponds,
average [SO42"]
1,490 mg/L
Deposition not
reported
In July, surface water
[MeHg] was 290%
higher in upland
ponds than in lowland
ponds. In August,
surface water [MeHg]
was 233% higher in
upland ponds.
Surface water %MeHg
was 185% higher in
upland ponds, and
sediment %MeHg was
205% higher in upland
ponds.
Surface water [MeHg]
increases with
sediment organic
matter: MeHg
(ng/L) = -2.26 + 0.265
(percentage OM).
St Denis
National
Wildlife Area,
Saskatchewan,
Canada
Four lowland
ponds with
high SO42"
Five upland
ponds with
low SO42"
Hoqqarth et al.
(2015)
Wetlands
Deposition not
reported
Ambient wetland
[SC>42"]<5 mg/L
Mine-impacted
wetlands,
[S042~]>500 mg/L
Peat %MeHg
increased with soil
acid volatile sulfide in
ambient wetland.
In mine-impacted
wetlands, peat
%MeHg decreased
with increasing S
reduction.
Two wetlands,
St Louis River
Watershed,
MN
One wetland is
downstream of
a mine and
receives high
SO42" effluent
Peat
samples
Johnson et al.
(2016b)
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Table 12-9 (Continued): New studies on correlations between sulfur (S) and
methylmercury (MeHg) in ecosystems.
Type of
Ecosystem
Additions or Load
(kg S/ha/yr)
Biological and
Chemical Effects
Study Site
Study
Species
Reference
Bogs
S addition or
deposition not
reported
Peatland pore water
median [SO42"]: 0.1,
0.4, 0.6, 3.0 mg/L
Pore water [SO42"]
(mg/L) was positively
correlated with pore
water MeHg
(Spearman R = 0.237)
and total Hg
(Spearman
R = 0.576).
two
Four
peatlands
in Marcell
Experimental
Forest, MN;
two in
Experimental
Lakes Area,
Ontario
Pore water
in peat
Mitchell et al.
(2008b)
Upland-peatland
interface
S deposition not
reported
Peat pore water
[SO42"] = 0-20 mg/L
MeHg was positively
correlated with sulfate
(Spearman R = 0.39):
ng
MeHg/L = 0.037 * (mg
S0427L) + 0.58.
S2 and S6
peatlands,
Marcell
Experimental
Forest, MN
Peat pore
water
Mitchell et al.
(2009)
Streams
S deposition not
During the growing
Archer Creek,
Stream
Selvendiran et

reported
season, MeHg
Adirondacks,
water
al. (2008a)

Monthly [S042"]
increased with
NY



means in streams:
decreasing sulfate




4-8 mg/L
concentrations: ng




MeHg/L = -0.11 x (mg
S042"/L) + 0.88.



Streams and lakes Not reported
MeHg was negatively	Arbutus Lake
correlated with water	and Sunday
sulfate concentrations:	Lake,
ng	Adirondacks,
MeHg/L = 6.67 x (mg	NY
S042"/L) - 2.40.
Stream and
lake surface
water
Selvendiran et
al. (2009)
DOM = dissolved organic matter; DR = Devils River; g = gram; ha = hectare; kg = kilogram; L = liter; MeHg = methylmercury;
mg = milligram; ng = nanogram; OM = organic matter; RG = Rio Grande; S = sulfur; S042" = sulfate; THg = total mercury;
yr = year.
A study of fish and SOx deposition sediment records at Isle Royale in Lake Superior
indicated a temporal relationship in this remote location between S deposition and Hg
loads in fish. Isle Royale is a Class I area and contains several lakes impacted by S and
Hg deposition. Between 1981 and 2006, S deposition in the region fell 40-60% (from
about 10 kg S/ha/yr in 1981 to about 6 kg S/ha/yr in 2006 atNADP site MN18 in
mainland Minnesota), while Hg deposition remained stable. S deposition on the island
between 1985 and 2006 was in the range of 2.3 to 7 kg S/ha/yr with a trend (p = 0.09)
toward a decline in deposition over that time period (Drevnick et al.. 2007). Fish
collections were conducted in eight Isle Royale lakes in 1995-1996 and again in
2004-2006. In six lakes, 1995 collections of northern pike (Esox lucius) exceeded the
U.S. EPA fish tissue Hg limit (0.3 |ig Hg/g wet weight). When these lakes were
resampled in 2004-2006, northern pike mean Hg levels were below the U.S. EPA limit,
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which the authors attributed to a decline in SOx deposition despite steady Hg deposition.
This study also used museum collections of fish from Isle Royale (collected in 1905,
1929, 1966) and an analysis of lake sediment cores to test the relationship between
historical S loads and fish Hg burdens. The chromium-reducible sulfur (CRS) fraction in
sediment cores was determined to closely track known patterns of S deposition since
monitoring began in 1985, allowing the inference of S loads to island lakes since the
Industrial Revolution in the late nineteenth century. The historical S load data were
compared to Hg loads in both the historical and recent fish collections from Isle Royale
lakes. The CRS fraction explained 79% of the variation in Hg concentration of northern
pike fish fillets. There were similar correlations between sediment reduced S and Hg
burden in other fish species, but sample sizes for those species were low and correlations
were not statistically significant (Drevnick et al.. 2007). DOC and pH were not directly
tested for their relationship with Hg levels in fish, but the authors note that there are no
significant trends in DOC or pH in these lakes from 1980-2006. This study showed that
Hg fish burdens increased in Isle Royale lakes with increasing SOx deposition in the
twentieth century, and that Hg fish burdens declined in response to reductions in SOx
deposition since the 1990s.
The other two studies that find effects of SOx deposition are summarized in the following
sections. A 12-year study of MeHg and fish Hg in Voyageurs National Park found that
declines in SOx deposition have resulted in MeHg declines only in lakes where DOC has
remained stable (see Appendix 12.3.5.3). A study that compiled data from reservoir fish
sampling in Texas found correlations between geographic variations in SOx and Hg
deposition and average fish Hg burden over 25 years (see Appendix 12.3.5.2).
Sampling at small scales within peatlands has shown positive relationships between
sulfate concentrations and MeHg concentrations (Table 12-9). In two peatlands at the
Marcell Experimental Forest, MN, researchers monitored water quality at the
upland-peatland interface, hypothesized to be an important zone for methylation due to
labile carbon inputs in runoff from terrestrial communities to the peatland microbial
community. Sulfate peat pore water concentrations ranged from 0-20 mg sulfate/L, and
MeHg concentrations increased 0.037 ng/L for each 1 mg/L increase in sulfate (Mitchell
et al.. 2009). The relationship between sulfate and MeHg is also significant at a larger
spatial scale in the same region. Across four peatlands at the temperate-boreal boundary
in Minnesota and Ontario, MeHg was positively correlated with sulfate in peat pore water
(Spearman's R = 0.237) as well as with pH during the 2005 growing season (Mitchell et
al.. 2008b). The range of median sulfate levels in these peat bogs during the course of the
study was 0.01-3 mg/L, while the range of MeHg concentrations was 0.35-0.62 ng
MeHg/L (Mitchell et al.. 2008b).
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The South River in Waynesboro, VA, was contaminated by historical (1929-1950)
textile industrial use and release of Hg. In 2008, sediments were collected within the
channel at 10 study sites at increasing distances downstream of the historic Hg source.
MeHg concentrations in the sediment increased linearly with increasing sediment pore
water sulfate concentrations, with a 0.06 ng increase in MeHg per gram of sediment for
every 1 mg/L increase in sulfate pore water concentration, with mean pore water sulfate
concentrations ranging from 2-16 mg/L (Yu et al.. 2012).
Chemical constituents measured at watershed outflows represent the cumulative
microbial and abiotic chemical transformations within the watershed. A watershed in
which SRPs are present will therefore have higher MeHg and lower sulfate
concentrations in its outflow than an adjacent watershed without active SRPs that
experiences the same S and Hg deposition loads. Many of the following studies document
inverse relationships between sulfate concentrations and MeHg concentrations in
outflows of trunk streams (those that drain an entire watershed) where sulfate and Hg
concentrations reflect upstream sulfate stimulation of SRPs and Hg methylation. In a
tributary stream draining a wetland in Wisconsin, there were seasonal trends in sulfate
and MeHg concentrations indicative of wetland Hg methylation. Sulfate in the stream
declined 80% between April and June 2003, indicating S immobilization or reduction in
the wetland during the early summer, and MeHg increased 69% over the same time
period ("Watras and Morrison. 2008). This increase in MeHg and simultaneous decrease
in sulfate in stream water concentrations suggests upstream activity of SRB in Hg
methylation.
In streams draining wetlands in the Archer Creek watershed in the Adirondack
Mountains, NY, water sulfate concentrations were lower than in streams upstream of the
wetlands (Sclvcndiran et al.. 2008a'). Wetland microbial communities reduced sulfate
while methylating Hg, which would account for the elevated MeHg concentrations in
wetland-draining streams compared to upland streams. There was a negative correlation
between sulfate concentrations and MeHg concentrations in streams draining the
wetlands where Hg methylation occurred, with a 0.11 ng/L increase in MeHg
concentrations in the stream for each mg/L sulfate removed from surface water
(Sclvcndiran et al.. 2008a'). Selvendiran et al. (2009) measured and calculated Hg budgets
for the adjacent watersheds of Arbutus Lake and Sunday Lake in the Adirondacks, NY,
measuring inlet and outlet streams at both water bodies as well as lake surface water at
Arbutus. Across both watersheds, MeHg concentrations were inversely correlated with
sulfate concentrations. This pattern was driven by the concentrations measured in the
streams in Sunday Lake, which drains a watershed that is approximately 20% wetlands
(Selvendiran et al.. 2009). indicating that SRB in upstream wetlands may have been
reducing sulfate and methylating Hg to impact the measured stream surface water.
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Arbutus Lake drains a watershed of which wetlands comprise only 4% of surface area,
which may explain the relatively higher sulfate concentrations and the lower mean and
variability of MeHg (Sclvcndiran et al.. 2009). In watersheds where S and Hg deposition
rates do not differ spatially, negative correlations between sulfate concentrations and
MeHg show the link between sulfate reduction and Hg methylation.
Highly elevated sulfate concentrations, as can result from agricultural runoff, can depress
SRP activity and Hg methylation, as sulfide produced by SRPs accumulates and
downregulates SRP activity. In a sampling study of the Amistad International Reservoir
on the Texas-Mexico border, Hg pools in sediment, water, and largemouth bass (for more
on this endpoint, see Appendix 12.4) were quantified and compared between the Rio
Grande and Devils River branches of the reservoir. The Rio Grande branch had sulfate
concentrations an order of magnitude higher (site with highest sulfate concentration:
230 mg/L sediment pore water), as well as 40% higher total Hg pools in sediments, than
did the Devils River branch [site with highest sulfate concentrations: 23 mg/L (Becker et
al.. 2011)1. However, total Hg and sulfate were negatively correlated with MeHg, and
SRB (researchers used techniques to detect bacterial sulfur reducers) were present at all
sites. Hg methylation in this system was not limited by Hg supply or biota, and was not
stimulated by the Rio Grande's high sulfate levels, originating from agricultural land use
in the watershed. The Devils River branch had 65% higher MeHg concentrations in
sediments, and a 230% higher proportion of total sediment Hg in the form of MeHg
(Becker et al.. 2011). Becker etal. (2011) posited that the higher MeHg in Devils River
sediments was due to higher DOC in this area of the reservoir and to sulfate
concentrations (in sediment pore water, 2.5-23 mg/L) being optimal for the metabolism
of SRB.
Two wetlands in the St Louis River Watershed, MN, were the subject of a study of Hg
and S because one wetland regularly receives high-sulfate mine effluent (Johnson et al..
2016b). Peat %MeHg increased with S reduction (quantified as pore water sulfide or soil
acid volatile sulfide) in the wetland that did not receive mine effluent, and in which
surface water sulfate concentrations averaged less than 5 mg/L. In wetlands that received
sulfate mine effluent of >500 mg SO42 /L, peat %MeHg decreased with increasing S
reduction. Together, these results support the model of methylation first proposed for the
Everglades [see Figure 12-7. and Glimour (2011)1. in which there is a net methylation
optimum at low moderate sulfate concentrations (1—10s of mg SCV/L) and a
suppression of net methylation at high sulfate (100s of mg SCV/L) and sulfide
(0.6-0.7 mg S/L in pore water) concentrations (Johnson et al.. 2016b).
In contrast, in the St. Denis National Wildlife Area in Saskatchewan, Canada, prairie
potholes vary in sulfate concentrations in relation to geology and position in the
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landscape. In a study of water MeHg, five upland ponds had average sulfate
concentrations of 1,490 mg/L (S sources are geologic), and four lowland ponds had
average sulfate concentrations of 14 mg/L (Hoggarth ct al.. 2015). MeHg fraction was
185% higher in surface water and 205% higher in sediments in the upland ponds than in
the lowland ponds, which had lower sulfate concentrations (Hoggarth et al.. 2015).
12.3.5.2. Mercury and Sulfur Interactions in Ambient Conditions
It can be difficult to tease apart the effects of Hg and SOx deposition upon mercury
methylation rates or concentrations of MeHg in biota, because at some scales, Hg and
SOx emissions are correlated (Figure 12-11. Figure 12-12. and Table 12-10). In a
small-scale (15 km, single sampling event) gradient study of Hg, MeHg, and S downwind
of a coal plant and other industrial sources in China (Xu et al.. 2017a). soil S and gaseous
elemental Hg (GEM) were positively correlated, with effects upon cultivated rice. In
surrounding rice paddies, MeHg and %MeHg in rice grains were positively correlated
with soil S, soil MeHg, and atmospheric concentrations of gaseous elemental Hg
(Figure 12-11).
Regionally, Hg and S depositions were correlated in Texas (Drenner et al.. 2011) and
atmospheric reactive Hg and SO2 were correlated in the Ohio River valley (Yatavelli et
al.. 2006). At a national or international spatial scale, different factors govern long-range
transport and deposition of SOx, PM, and Hg, so S deposition and Hg deposition may not
be spatially correlated. However, at a longer temporal scale, Hg and S deposition have
declined in concert over the past decades in the U.S. [U.S. emissions of Hg declined 79%
and SO2 emissions declined 73% between 1990 and 2011 (U.S. EPA. 2017a)l. Temporal
and spatial correlations between Hg and S emissions and deposition make it difficult to
distinguish between Hg and SO42 as the limiting factor in ecosystems receiving high
deposition.
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5.0
4.0-
3.0-
SjQ
bO
i
tsfi
~ 2.0
O
o
2
1.0-
0.0
0.0
r=0.56, p<0.05
.-'-'Q °o
r-0.68, p<0.01
O
o
• Rice McHg
O Rice MeHg ratio
-80
60 o
«d
t-H
bX)
40 K
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Table 12-10 New studies on deposition of sulfur (S) and mercury (Hg) and their
effect on methylmercury (MeHg).
Type of
Ecosystem
Additions or
Load (kg
S/ha/yr)
Biological and Chemical	Study
Effects	Study Site Species
Reference
Reservoir NADP quantifies Mean Hg fish levels were
S deposition in
2008 at CT:
7.7 kg S0427ha,
TBP: 8.1 kg
S0427ha, ECTP:
8.1 kg S0427ha,
SCTP: 11.9 kg
S0427ha
61-89% higher in the SCP
region (highest S and Hg
deposition) than in other
regions.
145 reservoirs
in four Level 3
ecoregions of
eastern
Texas: CT,
TBP, ECTP,
SCTP
Micropterus
salmoides
(largemouth
bass)
Drenner et al.
(2011)
Lake	S deposition not Between 2004-2015, Hg°
reported	deposition decreased 25%,
while Hg2 deposition remained
constant.
Sulfate concentrations in
Arbutus Lake declined
approximately 40% between
2004-2016. In 2004, sulfate
concentrations ranged
4.3-7.0	mg/L, and in 2016,
sulfate concentrations ranged
2.4-4.3	mg/L.
Between 2004-2015, lake
MeHg and THg concentrations
and fluxes decreased
statistically significantly (mean
changes not reported due to
high variability).
Arbutus Lake
watershed,
Adirondacks,
NY
Water Gerson and Driscoll
samples (2016)
Agricultural S and Hg
wetland deposition not
recorded
MeHg and %MeHg in rice grains Gradient of S Oryza
were positively correlated with
soil S, soil MeHg, and
atmospheric concentrations of
gaseous elemental Hg.
and Hg
deposition
2-15 km
downwind of
Yueyang
Coal-fired
Power Plant,
Hunan
Province,
China
Xu et al. (2017a)
sativa (rice)
grown in
paddies
CT = Cross Timber; ECTP = East Central Texas Plains; ha = hectare; kg = kilogram; MeHg = methylmercury; NADP = National
Atmospheric Deposition Program; S = sulfur; SCTP = South Central Texas Plains; S042" = sulfate; SOx = sulfur oxides;
TBP = Texas Blackland Prairie; yr = year.
1	A study that compiled data from reservoir fish sampling in Texas found correlations
2	between geographic variations in SOx deposition and average fish Hg burden over
3	25 years. Largemouth bass (Micropterus salmoides) were collected between 1985 and
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2009 with heavy sampling between 2004 and 2008 (45% of the samples) from
145 reservoirs in eastern Texas. The samples were grouped and analyzed by Level 3
ecoregions which received varying amounts of Hg and S deposition, increasing from the
westernmost Cross Timber region where S deposition was 7.7 kg/ha to the easternmost
South Central Plains region where S deposition was 11.9 kg/ha (Premier et al.. 2011).
The South Central Plains region received high deposition of SOx and Hg, and also had
the highest percentage of land area as wetlands (12%) and the lowest percentage of land
area in agriculture of the four regions, all of which are conducive to high methylation
rates in aquatic ecosystems. Mean Hg levels in sampled largemouth bass were 61-89%
higher in the South Central Plains than in other eastern Texas regions (Premier et al..
2011). These results have regional implications because the South Central Plains
ecoregion extends into Arkansas, Louisiana, and Oklahoma. This study showed that
within Texas, the geographic region with the most wetlands and the highest SOx
deposition also had fish with the highest Hg burden.
In the Adirondacks, a temporal study of water MeHg, water sulfate concentrations, and
mercury deposition reported no correlation between stream sulfate and MeHg (as
concentration or as %MeHg) in the summer when SRB are active, in lake data collected
in 2004-2005 (Gcrson and Driscoll. 2016). Over the decade in which these data were
collected, Hg° atmospheric concentration and deposition in the watershed, litter Hg
concentration, and sulfate concentrations in streams all declined in this system; Gerson
and Driscoll (2016) suggested that decreases in surface water MeHg in the watershed
were attributable to decreased Hg deposition, and that surface water sulfate concentration
over the entire decade was so high in the lake that it did not control Hg methylation.
12.3.5.3. Organic Matter and Sulfur Interactions in Ambient Conditions
There is evidence from recent research of interactions between organic matter and sulfur
in controlling mercury methylation (Table 12-11). A 12-year study of MeHg and fish Hg
in Voyageurs National Park found that declines in SOx deposition have resulted in MeHg
declines only in lakes where DOC has remained stable. In lakes in Voyageurs National
Park, MN, a Class I area, researchers measured MeHg in the epilimnion of four lakes
between 2000 and 2012 and compared trends in lake water Hg concentrations with trends
of measured Hg and S deposition (Brigham et al.. 2014). S deposition to the region
decreased 48% between 1998 and 2012 based on regression analysis of data measured at
National Atmospheric Deposition Program (NADP) stations MN16 and MN18, and Hg
deposition decreased 32%. In two lakes, MeHg surface water concentrations decreased
over 12 years of measurement, 50% in Peary Lake and 43% in Ryan Lake (Brigham et
al.. 2014). and fish Hg concentrations followed suit (see Appendix 12.4). In contrast,
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Brown Lake MeHg concentrations increased 85% over the same time period, which the
authors ascribed to increasing inputs from upstream methylating aquatic environments.
An alternative explanation may be in the differences between lakes in dissolved organic
carbon (DOC) concentrations; in Peary and Ryan Lakes, DOC concentrations remained
flat over the decade in which MeHg concentrations decreased, while in Brown Lake,
DOC concentrations increased 30% (Brigham et al.. 2014). Percaflavescens (yellow
perch) were collected between 2000 and 2012 to determine fish Hg levels in comparison
with Hg lake water trends and S deposition. Between 2000 and 2012, P. flavescens Hg
concentrations decreased 37% in Peary Lake and 32% in Ryan Lake, just as S deposition
and lake MeHg declined (Brigham et al.. 2014). In Brown Lake, where DOC increased
30% over the same time period, P. flavescens Hg increased 80% (Brigham et al.. 2014).
In Voyageurs National Park, lakes with stable DOC responded to decreased S deposition
with decreasing MeHg in water and Hg concentrations in fish.
There is evidence from recent research that higher sulfate concentrations correspond to
higher MeHg concentrations and sulfate reduction rates in prairie potholes . Prairie
potholes are shallow wetlands and lakes in depressions created by glaciation, which
depend upon snowmelt and precipitation for their water supply, and tend to be
hydrologically isolated. A survey of prairie pothole wetlands and lakes in Saskatchewan
measured Hg levels in water. Surface water MeHg correlated positively with sulfate
concentrations and negatively with conductivity and specific ultraviolet absorbance
(SUVA; an indicator of aromatic fraction of DOM), such that surface water MeHg
increased 0.03 ng/L for each mg/L increase in sulfate concentrations, when all other
factors remained constant (Hall et al.. 2009a).
In wetlands in the Allequash Watershed, WI, potential methylation rates in streambed
sediments were positively correlated with microbial sulfate reduction (quantified as
sediment acid volatile sulfide, r2 = 0.258), DOC, and reduced iron (Creswell et al.. 2017).
These correlations reflect the activity of methylating SRB, IRB, and methanogens, which
shift in dominance spatially and seasonally across the wetland streams (see
Appendix 12.3.2). Hg concentrations did not correlate with mercury methylation in this
system (Creswell et al.. 2017). The total dissolved S and sulfide fraction did not correlate
with methylation rate, suggesting that sulfate was not a limiting factor, and that DOC was
probably the main driver of methylation in this system.
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Table 12-11 New studies of Interactions between organic matter and sulfur (S) in
Controlling methylmercury (MeHg).

Additions or




Type of
Load (kg
Biological and Chemical

Study

Ecosystem
S/ha/yr)
Effects
Study Site
Species
Reference
Lakes
6.44 kg S/ha in
As SOx deposition decreased
Voyageurs
Surface
Briqham et al.

1998, 3.35 kg
48% in 1998-2012, water MeHg
National Park,
water and
(2014)

S/ha in 2012
decreased 50% in Peary Lake
MN
Perca


as quantified
and 43% in Ryan Lake. In Brown

flavescens


by NADP sites
Lake, water MeHg increased

(yellow


MN16 and
85% in 1998-2012.

perch)


MN18
Between 2000 and 2012, Hg
concentration in P. flavescens
decreased 37% in Peary Lake
and 32% in Ryan Lake, and
increased 80% in Brown Lake.



Wetland
Pore water
Methylation was correlated with
Allequash
Pore water
Creswell et al.

concentrations
acid volatile sulfide (AVS)
Creek, Wl

(2017)

0.0 to 59.2 |jM
(r2 = 0.258, p = 0.001), and




S042"
stepwise with DOC, and
porewater Fe(ll), THg, MeHg and
dissolved Fe (r2 = 0.5096,
p = 0.004).



Reservoir
Deposition not
reported
Permanently
inundated
sediment
porewater:
3.8 mg S042"/L
Seasonally
inundated:
8.2 mg S042"/L
Seasonal inundation increased
sediment MeHg and MeHg
fraction of total Hg in comparison
to permanently inundated
sections:
Seasonal: 2.70 ng MeHg/g
sediment, 0.55% MeHg
Permanent: 0.71 ng MeHg/g soil,
0.18% MeHg
Cottage Grove
Reservoir, OR.
Sediment,
pore water,
and surface
water
Ecklev etal. (2017)
Lakes and
Not reported
Surface water MeHg was
49 lakes and
Lake water Hall et al. (2009a)
wetlands

positively correlated with sulfate
wetlands,



concentrations: ng
prairie pothole



MeHg/L = -2.94 - 0.44
region,



(conductivity) + 0.303(S042") -
Saskatchewan,



0.268 (aromatic fraction of DOM).
Canada

Artificial
Concentrations
Strong correlation noted between
Reservoirs in
Surface Noh et al. (2016)
reservoir
ranged from
sulfate and chlorophyll a
South Korea
water

3-12 mg/L
(r = 0.73), DOC (r = 0.92), and



S042"
conductivity (r= 0.83)


Reservoirs have been identified as sources of MeHg. Cottage Grove Reservoir, OR, was
built for flood control and experiences large, seasonal fluctuations in water levels. Hg and
MeHg dynamics differ between permanently flooded portions of the reservoir and areas
that are inundated during summer months (Ecklev et al.. 2017). When the reservoir was
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flooded in the spring, sulfate, inorganic Hg, and DOC were released in high
concentrations from previously exposed sediments, creating zones of high mercury
methylation in seasonally inundated sediments. Across nine reservoirs in South Korea,
algal blooms stimulated methylation of mercury. Algal blooms released large pulses of
labile C, as demonstrated by a positive correlation of Chi a and DOC (r = 0.73) (Nohet
al.. 2016). Sulfate concentrations ranged from 3 to 12 mg/L, and also correlated with
DOC (r = 0.92). Concentrations of MeHg and %MeHg were positively correlated with
Chi a, DOC, and sulfate across reservoirs.
12.3.5.4. pH and Sulfur Interactions in Ambient Conditions
Sulfate deposition is not the primary driver of MeHg concentrations in every ecosystem
with elevated Hg concentrations in its biota. In Wallowa-Whitman National Forest, OR
and ID, fish Hg sampled in high-elevation lakes in 2011 correlated positively with basal
area of conifers in the lake catchment and correlated negatively with lake area, surface
water SO42 . and DOC (Eagles-Smith et al.. 2016). This study did not test microbial
methylation rates or community composition, but lakes with low SO42 and low DOC
were acidified, as quantified by pH measurements. Eagles-Smith et al. (2016) interpreted
model results to show that acidification of lakes had negative effects upon fish condition
and resulted in high Hg bioaccumulation, with conifer-forested catchments most at risk.
In Kejimkujik National Park, Nova Scotia, regressions and model selection of Hg in biota
against S34S, S13C, and S15N suggested that lake productivity was a stronger driver of Hg
in biota than sulfur reduction (indicated by enrichment of S34S), which was negatively
correlated with Hg in lake biota (Clavden et al.. 2017). The lakes in this study are
unstratified, shallow and acidic, with S deposition enriched in S34S from nearby marine
waters, all conditions unfavorable to sulfate-reducing bacteria. Clavden et al. (2017)
suggested that microbial methylators not utilizing sulfate are responsible for producing
MeHg in these lakes, and that trophic interactions are responsible for variations in Hg in
biota across lakes.
12.3.5.5. Nutrient Enrichment Effects on Methylmercury (MeHg)
Nutrient availability—the relative availability of nitrogen (N) and P as well as of S—can
impact MeHg fraction, with low or intermediate levels of nutrients most conducive to
higher MeHg across two very different nutrient gradients. In seven wetlands across a
nutrient gradient in Sweden, total Hg was 113-287 ng/g in soils, MeHg was 3.5-21 ng/g
in soils, and the highest MeHg fraction was 16%. Potential methylation rate constants and
MeHg fractions were highest in wetlands with an intermediate nutrient status (Tjerngren
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ct al.. 2012). N and P concentrations have also been implicated as contributing to
variation in Hg methylation rate in the Everglades (see Appendix 12.3.4.3).
Sulfur Impacts on Mercury in Wildlife
Mercury has no known beneficial property in living cells. MeHg is bioconcentrated by
living cells, in which it forms strong bonds with thiosulfate and sulfhydryl groups in
organic compounds. It is common for MeHg to be present at progressively higher
concentrations up the trophic level of food chains, as MeHg consumed is remarkably
persistent (Figure 12-13). A recent report by the USGS reviews the known effects of Hg
on wildlife. Fish and birds experience negative physiological and reproductive effects at
Hg levels below the 0.3 ppm tissue level set by the U.S. EPA to protect human health
(Wentz et al.. 2014). There are a number of new studies on the effects of sulfur upon Hg
in wildlife (Figure 12-12).
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r Top
predator fish
rtebrates
Algae and other
inicroorganisnrs
Water
USEPA fish tissue mercury criterion =0,3 parts per million
0.001
Łr 0.0001
E
H 0.00001
0.000001
0.0000001
Note logarithmic scale of y-axis. The top and bottom of each circle represent the range of measured methylmercury concentrations
from streams in Oregon, Wisconsin, and Florida sampled during 2002-2006, and in New York and South Carolina sampled during
2007-2009.
Source: from Wentz et al. (2014).
Figure 12-13 Bioconcentration and biomagnification result in methylmercury
concentrations about 1 million times higher in predator fish than
in stream water.
0.00000001
Increasing trophic level
Forage itsh
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Table 12-12 New studies on sulfur (S) impacts upon mercury (Hg) in wildlife.
Type of
Ecosystem
Additions or Load
(kg S/ha/yr)
Biological and
Chemical Effects
Study Site Study Species Reference
River	S addition or
deposition not
reported
[SO42"] in Rio Grande:
pore water: (means)
74-230 mg/L; deep
water: (range)
152-342 mg/L
[SO42"] in Devils
River: pore water:
(means)
74-230 mg/L; deep
water: (range)
15-168 mg/L
M. salmoides muscle Hg
(mg/fish kg) was 38%
higher in DR than RG.
Rio Grande
(RG) and Devils
River (DR)
branches of
Amistad
International
Reservoir, TX
Sediment, pore
water, and
deep water;
Micropterus
salmoides
(largemouth
bass)
Becker et al.
(2011)
Constructed
S load as [SO42 ] in
Fish Hg (Y = log mg
Western Palm
Gambusia
Fena et al.
wetlands
wetland inflow:
Hg/kg G. holbrooki) was
Beach County,
holbrooki
(2014)

39-110 mg/L
negatively correlated with
FL
(eastern



water sulfate (X = log mg

mosquitofish)



SC>42"/L) concentrations

and surface



in three treatment

water



wetlands:





Y = -2.09X + 2.60;





Y = -3.17X + 3.80;





Y = -1.09X- 0.01.



Wetlands
S load as [SO42"] in
Fish Hg (mg Hg/kg or mg
Everglades
Gambusia spp.
Gabriel et al.
and canals
surface water:
Hg/kg/mm for Lepomis)
protection area,
(mosquitofish),
(2014)

0-60 mg/L
increased over 0-1 mg/L
FL
Lepomis spp.



SO42", reached peak

(sunfish), and



levels at 1-12 mg/L

Micropterus



SO42", and decreased

salmoides



over 12-25 mg/L SO42",

(largemouth



to remain flat at SO42"

bass)



concentrations





25-60 mg/L.



Lakes
Not reported
Fish Hg is positively
One
Sander vitreus
Stone et al.


correlated (r2 = 0.928)
impoundment
(walleye)
(2011)


with lake water SO42"
and five natural




(mg/L).
lakes, SD


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Table 12-12 (Continued): New studies on sulfur (S) impacts upon mercury (Hg) in
wildlife.
Type of Additions or Load	Biological and
Ecosystem (kg S/ha/yr)	Chemical Effects	Study Site Study Species Reference
Bog
Ambient S deposition:
5.5 kg/ha/yr (mid
2000s) as quantified
by NADP Station
MN16
Hg levels in Culex spp.
increased 126%.
Culex spp. Hg was 41%
lower in recovery
treatment than under
SO42" treatment, but still
34% higher than control.
S6 peatland,	Pore water, Wasik et al.
bog section,	peat, and	(2012)
Marcell	Culex
Experimental	spp.(mosquito)
Forest, MN	larvae
S addition:
32 kg S/ha/yr
dissolved in pond
water and delivered
by sprinklers (mimics
4x the 1990s
deposition rate)
DR = Devils River; ha = hectare; Hg = mercury; kg = kilogram; L = liter; mg = milligram; NADP = National Atmospheric Deposition
Program; RG = Rio Grande; S = sulfur; S042" = sulfate; yr = year.
S deposition affects the production and net amount of MeHg in water bodies, increasing
accumulation of Hg up the food chain (Table 12-12) but not affecting the rates at which
Hg is transferred between trophic levels. Early work by Bloom et al. (1991) in seepage
lakes of varying pH in the Northern Highland Lake District, WI, considered whether
varying deposition would affect the rates at which MeHg accumulates in biota. The
authors sampled five lakes with common geology but a range of pH (5.1-7.2, including
the experimentally acidified north basin of Little Rock Lake), which the authors ascribed
to varying acidic deposition loads. The fraction of Hg was higher in Perca flavescens,
yellow perch, than in seston, reflecting bioaccumulation of MeHg at higher trophic
levels, but there was no detected effect of pH (and by extension, S deposition load) on
relative bioaccumulation across lakes (Bloom et al.. 1991). In Dickie Lake, Ontario, early
studies documented the biomagnification of Hg in higher trophic levels of the lake.
MeHg concentrations were 0.2 ng/g in sediments; Hg concentrations were 56.5 ng/g in
periphyton, 143 ng/g in P. flavescens, and 703 ng/g inMicropterus spp. [largemouth and
smallmouth bass, (Kerry etal.. 1991)1. Sulfate effects on biomagnification were not
documented in these studies, although sulfate was measured at 6-8 mg/L in lake water,
and microbial sulfate reduction and Hg methylation within the lake sediments were
demonstrated with sediment slurry incubations (Kerry et al.. 1991). There is evidence
from 44 Adirondack lakes sampled in 2003-2004 that the legacy of acidic deposition,
low surface water pH, low ANC, and high aluminum concentrations, all correlated with
increasing Hg concentrations in zooplankton and Hg body burdens in fish and loons (Yu
et al.. 2011). Because these factors did not affect Hg or MeHg concentrations in water,
Yu et al. (2011) concluded that acidic lake conditions enhanced bioaccumulation of
MeHg.
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Experimental S addition studies demonstrate that S addition increases Hg burdens in
different trophic levels. Little Rock Lake, WI was the site of a multiyear (1985-1990),
progressively acidic S addition study, in which the basins of this lake were separated by a
plastic barrier, allowing the northern half to be experimentally acidified while the
southern half served as a control. Abiotic compartments and biota of the lake were
intensively sampled during the course of the experiment. Aqueous MeHg concentrations
were 2 times higher in the treatment basin, while zooplankton MeHg concentrations
(ng/g) were 1.8 times higher and microseston MeHg concentrations were 2.75 times
higher in the treatment than the control basin (Frost etal.. 1999; Watras and Bloom.
1992). S amendment also affected MeHg in Perca flavescens (yellow perch); MeHg
concentrations were 12-100% higher in 1-year-old yellow perch from the treatment
basin, and the annual mean burden (fig MeHg/fish) was 16-214% higher in the treatment
basin over 5 of the 6 years of the experiment (Frost etal.. 1999; Wiener etal.. 1990).
Decreasing S loads result in decreased Hg burdens in biota. The S addition experiment at
the Marcell peat bog also shows the effects of S addition on Hg burdens in an animal on a
lower trophic level. In the S addition and S recovery experiments at the S6 peat bog,
Marcell Experimental Forest, MN, elevated S increased Hg levels in sampled Culex spp.
(mosquito) larvae. Mosquito larvae are important to consider as a food source because
they are widely consumed by invertebrate predators, fish, amphibians, and birds. In
spring 2009, mosquito larvae Hg was 126% higher in plots that had received
32 kg S/ha/yr addition for 9 years than in mosquito larvae from the control plots (Wasik
ct al.. 2012). The recovery plots received 32 kg S/ha/yr from 2001-2006 and no S
addition after 2006, and when these plots were sampled in 2009, Culex spp. (mosquito)
larvae had Hg levels 34% higher in recovery plots than in control plot mosquitoes,
indicating legacy effects of past S addition. However, recovery plot mosquito Hg
concentrations were 41% lower than in mosquitoes from plots still receiving S additions,
indicating that decreases in S addition will decrease biotic Hg concentrations (Wasik ct
al.. 2012).
In stormwater treatment wetlands draining the EAA in Florida, sulfate concentrations
have been elevated by agricultural runoff, and concentrations ranged from 39-110 mg/L
between 2000 and 2007. Under these very high sulfate surface water concentrations, Hg
levels (mg Hg/kg fish) in bioindicator fish species Gambusia holbrooki (mosquitofish)
were negatively correlated with inflow water sulfate concentrations (Feng et al.. 2014)
because the higher end of the sulfate concentration range reached levels inhibitory to
SRPs. In the Everglades Protection Area, where water sulfate concentrations were lower
and ranged from 0-60 mg/L in surface water, relationships between sulfate and fish Hg
levels were more complicated. Gambusia spp. (mosquitofish), Lepomis spp. (sunfish),
and largemouth bass were sampled annually between 1998 and 2009 (Gabriel et al..
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2014).	In all three fishes, Hg tissue concentrations increased between 0-1 mg sulfate/L in
the water, and reached their highest tissue levels between 1 and 12 mg sulfate/L (see
Figure 12-14). Based on these relationships, Gabriel et al. (2014) recommended a sulfate
water standard of 1 mg/L. Two papers challenged this analysis and interpretation of the
sulfate and fish Hg data. In a commentary, Julian et al. (2015) suggested that the
variability in fish Hg levels was too high to correlate with sulfate concentrations, and in
analyses of largemouth bass collected between 2005-2011, Hg concentrations in fish
correlated with alkalinity, pH, and specific conductance as well as sulfate (Julian and Gu.
2015).	Gabriel et al. (2015) responded to Julian's commentary by highlighting the
similarity of fish Hg patterns to documented MeHg water column concentrations in the
Florida Everglades and suggested that sulfate reductions are the most efficient way to
reduce Hg loads in fish because about 60% of the Everglades area has sulfate
concentrations exceeding 1 mg/L (Orem etal.. 2011). Using structural equation
modeling, Pollman (2014) showed that sulfate was second only to periphyton MeHg
concentrations in predicting Hg burdens in mosquitofish in the Everglades, although the
relationships among predictive factors were complex and involved both direct and
indirect effects.
A number of studies document Hg burdens in fish without also quantifying contributions
of S to methylation, but other relationships between S and MeHg documented in the 2008
ISA suggest that the findings of these studies may be relevant (U.S. EPA. 2008a). In the
Alabama River basin around Mobile, AL, there was no relationship across 52 sites in the
watershed between Hg burdens in largemouth bass and MeHg fractions in sediments or
water (Warner et al.. 2005). However, a subset of seven of the sites were downstream of
wetlands, and in these sites the bass Hg burden was positively correlated with watershed
area (r2 = 0.71), indicating that wetlands may contribute to higher Hg burdens in fish as
well as to higher MeHg concentrations in water (Warner et al.. 2005). In a study of Hg
burdens in South Dakota, Stone etal. (2011) analyzed water quality and fish samples
from one impoundment and five natural lakes, and found a positive correlation
(If = 0.928) between water sulfate measurements and Hg in piscivorous Sander vitreus
(walleye). However, sulfate concentrations were not reported, and sulfate measurements
and fish samples were not collected simultaneously, with a gap of up to 3 years between
collection of fish and water samples. In the Amistad International Reservoir, Texas,
elevated MeHg in sediments of one branch of the reservoir correlated with Hg
accumulation in young (age: 0-3 years) largemouth bass. The Devils River area had
elevated sediment MeHg compared to the Rio Grande area of the reservoir (see
Appendix 12.3.5.1). and the length-standardized mean of muscle Hg was 38% higher in
Devils River than in Rio Grande bass (Becker etal.. 2011). Of 138 largemouth bass (age
>3 years) line-caught at the reservoir in April 2007, 77% exceeded the U.S. EPA
recommended human consumption value of 0.3 mg Hg/kg (or 0.3 ppm), and mean fish
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Hg was 0.51 mg/kg fish, although fishing site data for these samples were unavailable,
preventing assessment of connections between sediment MeHg and Hg levels in these
large fish (Becker et al.. 2011).
0.35
0.30 -
0.26 -
¦ running average
D
i J~T~H m
• I	I
i iij i	•
_	I J ~* ~ W\ I
li l» • I ml	rnJ=i
I	®H	• ^ I

-•	hH
rm	1—i—i i i i 11111111 ii 11	1—i—i i i i i 1111111111'	1—i—i i i i 1111111
0.006
e
E
o>
E
0.005
0.004
0.003
0.002
0.001 -
0.000
E
w-
ft.#"
*•%:<
"3i
Ł

o>
<
Surface water sulfate (mg/L)
Surface water sulfate (mg/L)
L = liter; kg = kilogram; mg = milligram; mm = millimeter; THg = total mercury.
Panels A-C show the running average of mercury concentrations in (A) Gambusia spp. (n = 484), (B) Lepomis spp. (n = 2,559), and
(C) Micropterus salmoides (n = 679). Panels D-F show the same data with sulfate transformed to a log scale to better illustrate the
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peak in fish Hg concentrations that approximately corresponds to sulfate concentrations of 1-10 mg/L. Samples were collected from
12 fish and 12 sulfate stations within the Everglades, between 1998-2009.
Source: Figure 3 in Gabriel et al. (2014).
Figure 12-14 Tissue mercury concentrations as a function of surface water
sulfate concentrations (n = 2,360 surface water samples) in the
Everglades Protection Area.
12.5. Extent and Distribution of Sensitive Ecosystems
Reservoirs in which water levels fluctuate enough to expose sediments may be
particularly susceptible to S deposition effects on MeHg production because mud flats are
directly exposed to the atmosphere and then flooded to create conditions favorable to Hg
methylation by SRPs. Older work in the ELA, Ontario, showed that reservoir creation
leads to a pulse of MeHg production. A small watershed was flooded in 1993, and Hg
and MeHg fluxes were measured over the next decade (St. Louis et al.. 2004). Before
flooding, 90% of the annual total Hg input and 130% of the annual MeHg was retained
by the wetland. In the first 3 years following flooding, 100-170% of the annual total Hg
input was exported annually; and in Years 5-9, 69-79% of the annual total Hg input was
exported annually. MeHg levels were high following flooding, 60-80% in water, and
300-860% of the annual MeHg input was exported annually from the reservoir (St. Louis
et al.. 2004). The reservoir creation also affected Hg in the food web: the level of MeHg
in zooplankton was 7 to 10 times higher in the first 3 years of flooding, and 14 to
30 times higher Years 4-9, than MeHg in preflood zooplankton (St.Louis et al.. 2004).
Researchers stocked finescale dace (current scientific name, Chrosomus neogaeus)
annually and found that the Hg body burden of the fish increased 306% following a year
in the reservoir (St.Louis et al.. 2004). More recently, at the Cottage Grove Reservoir,
OR, elevated MeHg concentrations in the fall outflow (4-13% MeHg fraction compared
to 0.5-7% MeHg in the river flowing into the reservoir) were linked to seasonal water
fluctuations (Ecklev et al.. 2015). The reservoir water level is lowered from September to
February to allow the dam to catch and regulate spring flooding, so 50% of the reservoir
area is exposed mud flats in the winter. Sulfate was below detection limits (0.005 mg
sulfate/g sediment, reported as 5 jxg/g) in permanently inundated reservoir sediments, but
was 0.0011 mg sulfate/g sediment (11 jxg/g) in mudflat sediment. MeHg concentrations
were higher at sediment surfaces than in the sediment profile at seasonally inundated
sites, indicating that MeHg production was occurring there; in contrast, [MeHg] was
constant across sediment depths in the permanently inundated sites, indicating that
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settling of MeHg into sediments was the only process occurring there (Ecklev et al..
2015).
The National-Scale Assessment of Mercury Risk to Populations with High Consumption
of Self-Caught Freshwater Fish of 2011 (U.S. EPA. 201 lb) compiled fish Hg load data
on a national scale, which can be considered a county-wide sampling of ecosystems
where Hg loads in fish are high. The assessment did not directly quantify or analyze
sulfate effects, but it considered patterns of Hg deposition and Hg concentrations in
freshwater fish caught between 2005-2009 by state regulatory agencies, and modeled the
extent to which reductions in Hg deposition would affect subsidence-level human
consumers of fish from U.S. water bodies. Fish collections showed that there were
elevated fish Hg levels in water bodies across the country, although high fish Hg
concentrations were more common in the eastern U.S. [see Figure 12-15 (U.S. EPA.
2011b)l. In this study, 5% of total Hg deposition was attributed to coal and oil-fired
electric utility steam generating units (EGUs), but EGUs contributed 9% to fish tissue
levels of Hg.
The 2014 USGS's report on mercury in streams (Wentz et al.. 2014) suggests that eastern
water bodies, watersheds with high wetland cover, and watersheds with low population
densities are linked to higher MeHg concentrations in streams and in stream biota.
Fish Tissue Mercury Data
(HUC12-level 75th percentile values, ppm)
~	0.000000 - 0.109143
o	0.109144 - 0.195000
o	0.195001 - 0.300000
o	0.300001 - 0.479000
*	0.479001 - 6.605000
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HUC12 = 12-digit hydrologic unit code.
Source: U.S. EPA (2011b).
Figure 12-15 Fish mercury concentrations across the U.S.
12.6. Critical Loads
There are currently no critical loads for SOx deposition related to the nonacidifying
effects described in this appendix. There are European critical loads for Hg deposition
and Hg effects upon human health and ecosystems. A critical load of 1 |_ig Hg/L in soil
drainage water was set to protect drinking water (Hcttclingh et al.. 2015). A critical load
of 0.3 mg Hg/kg fresh weight fish tissue was set to protect human health. A critical load
to protect soil invertebtates and soil microbial process of 0.5 mg Hg/kg organic material
was set for the humus layer of forest soils (Hcttclingh et al.. 2015). This source did not
report critical loads in terms of areal deposition rates, but did map exceedances based on
soil, water, and fish tissue samples, and found that about 59% of European natural areas
received Hg deposition in exceedance of these critical loads (Hcttclingh et al.. 2015).
12.7. Summary of Nonacidifying Sulfur Effects
Sulfate deposition has a number of effects beyond acidification in ecosystems. In the
2008 ISA, the evidence was sufficient to infer a causal relationship between S deposition
and increased methylation of Hg in aquatic environments where the value of other factors
was within adequate range for methylation. The 2008 ISA surveyed literature from
approximately 1980 to 2008 and described the qualitative relationships between sulfate
deposition and a number of ecological endpoints, including altered S cycling, sulfide
phytotoxicity, internal eutrophication of aquatic systems, altered methane emissions,
increased Hg methylation, and increased Hg loading in animals, particularly fish. New
evidence from wetland and freshwater aquatic ecosystems strengthens and extends the
causal findings of the 2008 ISA regarding nonacidifying sulfur effects, and provides the
basis for a new causal determination. Table 12-13 summarizes recent thresholds and
quantitative relationships describing effects of sulfate deposition. This new information
strengthens the weight of evidence presented in the 2008 ISA, and consistent with the
2008 causal statement, the body of evidence is sufficient to infer a causal relationship
between S deposition and the alteration of Hg methylation in surface water,
sediment, and soils in wetland and freshwater ecosystems. In addition, recent research
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1	demonstrates sulfide phytotoxicity under current conditions in North American wetlands,
2	consistent with the 2008 ISA's summary of sulfide phytotoxicity in European wetland
3	and freshwater ecosystems. The body of evidence is sufficient to infer a causal
4	relationship between S deposition and changes in biota due to sulfide phytotoxicity
5	including alteration of growth and productivity, species physiology, species richness,
6	community composition, and biodiversity in wetland and freshwater ecosystems.
7
Table 12-13 Summary of quantitative effects of nonacidifying sulfur enrichment.
Section of
Nonacidifying Sulfur
Effects Appendix
Threshold, Critical Level, or Quantitative Relationship
Reference
12.2.1
2.5-7.9 times higher sulfate reduction in lake with 22.7 mg
S042"/L than in lake with 9.0 mg S042"/L
Kleeberq etal. (2016)
12.2.3
34.1 mg S2"/L for plant root nutrient uptake
Koch etal. (1990)
12.2.3
48-50 mg S042"/L as a threshold to protect Potamogeton spp.
and Utricularia vulgaris
Vermaat et al. (2016):
Smolders et al. (2003)
12.2.3
0.3-29.5 mg S2"/L for altered growth, productivity, physiology,
or mortality of 16 freshwater wetland emergent plant and
aquatic submerged macrophyte species native to North
America
Lamers et al. (2013)
12.2.3
250 mg S042"/L U.S. EPA aesthetic secondary water quality
standard for human consumption
Pastor et al. (2017)
12.2.3
<7.5 mg/L sulfide to preserve growth rate of Cladium
jamaicense, sawgrass, in Everglades
Li et al. (2009)
12.2.3
0.165 mg sulfide/L to protect Zizania palustris, wild rice
MPCA (2015a. 2015b)
12.3.2
Sulfate reduction rates increase 0.35 mg (sulfate
reduced)/L/day with a 1 mg/L increase in sulfate addition
Kerrvetal. (1991)
12.3.3.1
11.1 mg SC>42"/L added for peak MeHg production in incubation
of sediments from Quabbin Reservoir, MA
Gilmour et al. (1992)
12.3.3.1
5.8-23 mg SC>42"/L added to change incubation of sediments
from net demethylating to net demethylating, from Little Rock
Lake, Wl
Gilmour and Riedel
(1995)
12.3.3.1
190 mg SC>42"/L added to increase MeHg production in
incubation of peat from Sunday Lake, NY
Yu etal. (2010)
12.3.3.1
96.1	mg SC>42"/L increases methylation rates in incubations of
sediment from South River, VA (ambient concentrations of
19.2	mg S042"/L)
Yu etal. (2012)
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Table 12-13 (Continued): Summary of quantitative effects of nonacidifying sulfur
enrichment.
Section of
Nonacidifying Sulfur
Effects Appendix
Threshold, Critical Level, or Quantitative Relationship
Reference
12.3.3.5
<3.6 mg C/L as dissolved organic carbon, microbial sulfate
reduction did not occur
3.6-7.6 mg C/L of DOC, microbial sulfate reduction increases
linearly with C increase
>7.6 mg C/L, microbial sulfate reduction rate does not change
Watras et al. (2006)
12.3.3.5
20% organic material in sediments for peak MeHg
concentrations in Adirondacks, NY watersheds
Yu et al. (2010)
12.3.3.5
2.65 ng/L increase in MeHg for every 10% increase in
percentage organic matter
Hoaaarth et al. (2015)
12.3.3.5
1 ng MeHg/L increase for each increase of 0.048 mg
hydrophobic organic acid fraction of DOC/L
Hall et al. (2008)
12.3.3.5
0.34 ng Hg/L increase in Hg exported in streams for each 1 mg
C/L increase in DOC from Northeast watersheds
Dittman etal. (2010)
12.3.3.7
1.0 mg N-N03"/L in hypolimnion to prevent mercury methylation
in Onondaga Lake, NY
Matthews etal. (2013)
12.3.4.1
Addition of 4.8 mg S042"/L increases methylation rate in Little
Rock Lake, Wl (ambient water concentration, 2.4 mg S042"/L)
Gilmour and Riedel
(1995)
12.3.4.1
10.8 ng MeHg/L increase for each 1 mg/L increase in H2S in
lake water
Watras et al. (2006)
12.3.4.1
14.4 mg total lake MeHg increase for every 1 kg/ha increase in
SOx deposition
Watras and Morrison
(2008)
12.3.4.1
14 kg S/ha increased MeHg concentrations in peat mesocosm
Branfireun etal. (1999)
12.3.4.1
<8.3 kg S/ha addition for pore water MeHg in peatlands
Mitchell etal. (2008a)
12.3.4.1
<32 kg S/ha/yrto control pore water MeHg concentrations and
fraction of total Hg in peatlands
Jeremiason et al. (2006)
12.3.4.1
<20 kg S/ha/yr to control pore water total Hg concentrations and
MeHg concentrations in peatlands
Berqman et al. (2012)
12.3.4.3
1 mg/L sulfate in surface water to keep MeHg concentrations
low in Everglades surface water
Corrales et al. (2011)
12.3.4.3
>20 mg S042"/L inhibits Hg methylation in Everglades due to
sulfide accumulation
Orem et al. (2011)
12.3.4.3
>39 mg S042"/L inhibits Hg methylation in Everglades due to
sulfide accumulation
Zhenq etal. (2013)
12.3.4.3
50 ng total Hg/L, the U.S. EPA water quality criterion
Atoers et al. (2014)
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Table 12-13 (Continued): Summary of quantitative effects of nonacidifying sulfur
enrichment.
Section of
Nonacidifying Sulfur
Effects Appendix
Threshold, Critical Level, or Quantitative Relationship
Reference
12.3.5.1
<10 kg S/ha/yr, decreases in S deposition result in declining fish
Hg concentrations on a decadal time scale in Isle Royale lakes
Drevnick et al. (2007)
12.3.5.1
MeHg concentrations increased 0.037 ng/L for each 1 mg/L
increase in sulfate in pore water in peatlands
Mitchell et al. (2009)
12.3.5.1
0.06 ng increase in MeHg per gram of sediment for every
1 mg/L increase in sulfate pore water concentration in South
River, VA
Yu et al. (2012)
12.3.5.1
0.11 ng/L increase in stream MeHg concentration for each mg/L
sulfate decrease in Archer Creek, NY
Selvendiran et al.
(2008a)
12.3.5.1
65% higher MeHg in Devils River branch (SO42" 223 mg/L) than
in Rio Grande branch (SC>42~2230 mg/L) of Amistad Reservoir,
TX
Becker et al. (2011)
12.3.5.1
MeHg increases with S reduction in wetland where average
SO42" 25 mg/L, but not in wetland where SO42" 2500 mg/L
Johnson et al. (2016b)
12.3.5.2
11.9 kg S deposition/ha/yr increases Hg levels in largemouth
bass compared to ecoregions where deposition was 7.7-8.1 kg
S/ha/yr in Texas
Drenneretal. (2011)
12.3.5.2
MeHg decreased in surface water and in Perca flavescens
when S deposition decreased from 6.44 kg S/ha/yr (1998) to
3.35 kg S/ha/yr (2012) in two lakes in Voyageurs NP
Briqham et al. (2014)
12.3.5.3
When conductivity and aromatic DOC remain constant, surface
water MeHg increases 0.3 ng/L for each mg/L increase in
sulfate concentrations
Hall et al. (2009a)
12.4
<32 kg S/ha/yr, for Hg load in larval Culex spp. (mosquitoes) in
peatlands
Wasiketal. (2012)
12.4
1 mg/L sulfate in surface water to keep Hg concentrations low in
fish: Gambusia spp., Lepomis spp., and Micropterus salmoides,
in the Everglades
Gabriel et al. (2014)
C = carbon; DOC = dissolved organic carbon; H2S = hydrogen sulfide; ha = hectare; Hg = mercury; kg = kilogram; L = liter;
MeHg = methylmercury; mg = milligram; ng = nanogram; S = sulfur; yr = year.
1
2	12.7.1. Terrestrial Sulfur Cycling
3	The 2008 ISA noted that S is an essential plant nutrient and that S deposition can affect
4	plant protein synthesis by affecting S availability for S-containing amino acids as well by
5	affecting N uptake. Sulfate can be taken up directly by plant roots as well as by soil
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microbes, or it can adsorb to soil particles. The 2008 ISA reported that watersheds in the
Southeast, which retained more S than was deposited during high S loading in the 1980s,
received decreasing S deposition over the 1990s but continued to export historically
deposited sulfate in streams. In the Northeast, budgets of sites from the 1980s indicated
that high S deposition caused high sulfate export in streams. On a national level, S
deposition resulted in increased S content in organic matter in soils, and studies suggested
that mineralization of this stored S may contribute to elevated sulfate leaching for
decades after reductions in S deposition. More recent research confirms that watersheds
in the Northeast continue to leach more S than they currently receive from S deposition,
as much as 24-45% more S than the S deposition load (Mitchell et al.. 2011).
Aquatic Sulfur Cycling
The 2008 ISA reported that nonacidifying effects of S in freshwater systems will be
affected by water residence times, with larger lakes and beaver ponds providing ample
opportunity for SRPs to reduce sulfate and generate acid-neutralizing capacity (ANC).
Early work showed that a 1 mg/L increase in sulfate concentration in Dickie Lake,
Ontario, increased S reduction rates 0.35 mg S/L/day (Kerry et al.. 1991).
The 2008 ISA reported the role of wetlands in removing sulfate from surface water and
storing it in sediments or biomass, but it also highlighted the way drought reverses this
capacity to make wetlands sources of S to downstream waters. Specifically, studies
reviewed in the 2008 ISA demonstrated that under drought conditions, sulfide in
previously saturated moss or sediments is exposed to atmospheric oxygen and reoxidized
to sulfate. This resulted in high sulfate concentrations in surface water when the water
table rose to its former level, and MAGIC modeling showed that variable water levels
due to climate change-induced droughts would delay ANC recovery in the Plastic Lake
Watershed, NY (Aherne et al.. 2004). In Little Rock Lake, WI, water level fluctuation
due to drought added an additional 5 kg S/ha/yr from internal ecosystem S stores to the
lake (Watras and Morrison. 2008). Recent work confirmed that periods of drought result
in elevated water sulfate concentrations during the recovery period following the drought
(Wasiketal.. 2012).
Sulfide Toxicity
In aquatic systems or in saturated soils, SRPs convert sulfate into sulfide, which inhibits
nutrient uptake in freshwater aquatic and wetland plant species. The 2008 ISA showed
that sulfide toxicity reduced biomass of wetland plants and aquatic macrophytes in
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mesocosms under aquatic S concentrations higher than those that occur in U.S. regions
with high S deposition. Recent research has shown sulfide phytotoxicity occurs at
ambient aquatic S concentrations within multiple ecosystems in the U.S. (Figure 12-16).
Sulfide decreases total plant cover and cover of dominant species in a New York fen
(Simkin et al.. 2013) and decreases the growth rate of Everglades, FL, keystone species
Cladium jamcticense (sawgrass) at surface water concentrations of 7.5 mg sulfide/L (Li et
al.. 2009). The state of Minnesota is also working on a sulfide standard to protect the
economically and culturally important Zizania palustris (wild rice) and has proposed a
water standard of 0.165 mg sulfide/L to protect the species (MPCA. 2015a). Additionally,
a review by Lamers et al. (2013) reported experimentally-determined sulfide toxicity
values between 0.3-29.5 mg S27L for 16 North American species (see Table 12-2). and
there are sulfate tolerance values for North American fish and macroinvertebrates based
on USGS's NAWQA. This new information shows that sulfide toxicity occurs in North
American wetlands under current deposition conditions, and the body of evidence is
sufficient to infer a causal relationship between S deposition and changes in biota
due to sulfide phytotoxicity including alteration of growth and productivity, species
physiology, species richness, community composition, and biodiversity in wetland
and freshwater ecosystems.
Water quality thresholds for non-acidifying S effects
Increase in MeHg concentrations in surface water of Everglades
1 mg sulfate/L |ncrease in mosquitofish, sunfish, and largemouth bass Hg load in Everglades
7.5 mg sulfide/L Decrease in sawgrass
growth rate in Everglades
| 0.165 mg sulfide/L Decline in populations of wild rice in MN
0	2	4	6	8
mg/L
L = liter; Hg = mercury; MeHg = methylmercury; mg = milligram; S = sulfide.
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Figure 12-16 Thresholds of sulfate or sulfide concentrations in water, which
cause biological and chemical effects in ecosystems.
12.7.4. Internal Eutrophication
The 2008 ISA described the contribution of S deposition to internal eutrophication in
aquatic systems. In wetland and lake waters, sulfate is reduced to sulfide, which reacts
with Fe to form insoluble iron sulfide complexes. In many ecosystems, the iron in this
reaction is provided by FePC>4, which releases phosphorus as it forms FeSxthus
contributing to downstream eutrophication. More recently, internal eutrophication caused
by sulfate addition was observed in mesocosms of samples collected from Lake Moshui,
China (Yu et al.. 2015).
12.7.5. Effects on Methane Production
Sulfate deposition can shift microbial community interactions, resulting in lower methane
emissions. The 2008 ISA documented the suppression of methane emissions in wetland
soils by sulfate addition in several studies and noted that 15 kg S/ha/yr (the lowest S load
in experimental treatments) suppressed methane emissions to the same extent as higher S
loads (Gauci et al.. 2004). There is no recent research that relates SOx deposition to
wetland methane emissions, but recent research has confirmed the underlying microbial
mechanisms of this process. Sulfate deposition increases the abundance or metabolic
activity of SRPs. In many freshwater water bodies, including wetlands, low primary
productivity limits microbially available, labile C during the growing season when higher
temperatures favor microbial activity. Recent research documents competition among
microbial guilds for labile C, and when SRPs are favored, competing methanogens
decline, resulting in suppressed methane emissions (Bae et al.. 2015; He et al.. 2015;
Paulo et al.. 2015). However, there is also evidence of methanogens forming syntrophic
associations with SRPs (Paraniapc et al.. 2017). suggesting that these groups may not
always directly compete in all ecosystems.
12.7.6. The Role of Microbes in Mercury Methylation
The 2008 ISA identified sulfur-reducing bacteria as responsible for Hg methylation, and
identified anoxic wetland and lake bottom sediments as their location within watersheds.
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More recent research confirms that SRB in these locations methylate Hg, but also shows
that the ability to methylate Hg is more broadly distributed in the prokaryotic taxa and
across the environment.
Recent research suggests that the hgcAB gene pair confers the ability to methylate
inorganic Hg. HgcAB has been sequenced in lineages within bacteria and archaea, which
is why this document refers to mercury methylators as SRPs rather than the SRB
described in the previous ISA (Gilmour et al.. 2013). Experimental work has shown that
MeHg production rates vary among prokaryotic strains (Shao et al.. 2012). but also has
confirmed that sulfate addition causes Hg methylation in the Everglades (Goni-Urriza et
al.. 2015). as abundant syntrophs (heterotrophs dependent on the metabolic byproducts of
other microbes) were responsible for both Hg methylation and sulfate reduction (Bac et
al.. 2014).
Recent research shows that the microbial communities responsible for Hg methylation
are more widely distributed than in the freshwater lake or wetland bottom sediments
described in the 2008 ISA (Figure 12-17). There is evidence from Sweden, Finland, and
the U.S. that disturbances in terrestrial forests stimulate mercury methylation in soils and
MeHg export to surfacewaters (Kronbcrg et al.. 2016; Poulin et al.. 2016; Ukonmaanaho
et al.. 2016). The Sphagnum (moss) mat of wetlands was the most efficient retention site
of deposited Hg in a forested watershed in the Adirondacks (Selvendiran et al.. 2008b). as
well as an important location of mercury methylation within peat wetlands in New York
and Wisconsin (Yu et al.. 2010; Creswell et al.. 2008; Selvendiran et al.. 2008b).
Experiments using samples from rivers or streams in Virginia, Minnesota, or Germany
showed that higher sulfate concentrations increased methylation rates or [MeHg] (Frohnc
et al.. 2012; Yu et al.. 2012; Tsui et al.. 2008). with a 0.06 ng/L increase in MeHg per
gram of sediment for every 1 mg/L increase in sulfate pore water concentration in South
River, VA (Yu et al.. 2012). New research shows that SRPs actively methylate Hg within
periphyton, the aquatic biofilms attached to substrate or macrophytes in oxic freshwater
environments (Correia etal.. 2012; Achaetal.. 2011). MeHg production has also been
documented in estuarine and marine sediments of the Chesapeake Bay (Hollweg et al..
2009) and in the marine water columns sampled at 20 to 327 m depths in the Canadian
Arctic Archipelago (Lehnherr et al.. 2011).
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SK lakes (ng/L)*
5 -2
o
"SB 2
c
4
MN peatlands
~ (ng/L)
-4
-6
-8
-10
0
5
10	15
Surface water sulfate (mg/L)
20
25
C = carbon; cm = centimeter; DOM = dissolved organic matter; g = gram; L = liter; MeHg = methylmercury; mg = milligram;
mS = millisiemens; ng = nanogram; SK = Saskatchewan; S042- = sulfate.
Note: A study of prairie lakes and wetlands in Saskatchewan found that conductivity, aromatic DOM, and sulfate concentrations
correlate with water MeHg (ng MeHg/L) = -2.94 - 0.44 (conductivity, mS/cm) + 0.303 (mg S0427L) - 0.268 (aromatic fraction of
DOM, L/mg C/cm). Simple linear regressions relate MeHg to water sulfate concentrations in Minnesota peatlands (water
ng MeHg/L = 0.58 + 0.037 mg S042"/L) and in South River, VA (riverbed ng MeHg/g sediment = -0.08 + 0.059 mg S0427L).
Figure 12-17 Linear relationships between sulfate and methylmercury
concentrations in published studies.
As a microbial process, Hg methylation is determined not just by sulfate and Hg
concentrations, but by other environmental and nutritional requirements of SRPs: pH,
temperature, and water quality parameters, and carbon supply. The 2008 ISA identified
pH and dissolved organic carbon (DOC) to be major controls on Hg production, with low
pH and moderately high DOC correlating with high fish MeHg levels in lakes. The 2008
ISA also identified wetland area or density as an important determinant of MeHg
bioaccumulation and export in watersheds. Recent research presented in Appendix 12.3.3
evaluates temperature, total Hg concentration, pH, organic matter in water and sediments,
iron, and nitrate for their influence on Hg methylation rates.
The 2008 ISA identified organic C as an important control on microbial sulfate reduction
and Hg methylation, and recent research has quantified this relationship, hi Little Rock
Lake, WI, DOC concentrations had to be at least 3.6 mg C/L to allow microbial sulfate
reduction, and microbial sulfate reduction increased linearly with an increase of C only
when DOC was in the range of 3.6-7.6 mg C/L (Watras et al.. 2006). Across prairie
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potholes in Saskatchewan, surface water [MeHg] increased 2.65 ng/L for every 10%
increase in percentage organic matter in underlying sediment (Hoggarth et al.. 2015). In
wetlands, rivers, and lakes of the Mississippi River delta, LA, there was a 1 ng MeHg/L
increase in surface water for each additional 0.048 mg hydrophobic organic acid fraction
(aromatic carbon typical of peat) of DOC/L (Hall et al.. 2008).
Impacts of Sulfur upon Mercury Cycling
Based on experimental evidence, the 2008 ISA identified maximum MeHg production
occurring at sulfate concentrations of 200-400 (j,eq sulfate/L (9.6-19.2 mg SO42 /L) and
noted that waters under S deposition in the U.S. have sulfate concentrations in the range
of 60-200 (j,eq sulfate/L (2.9-9.6 mg SC>427L). More recent research shows that
detectable mercury methylation and bioaccumulation occurs at water sulfate
concentrations >1 mg/L in the Everglades (Gabriel et al.. 2014; Corrales et al.. 2011) and
at ambient S deposition and water sulfate concentrations within the U.S. There was a
positive relationship between sulfate reduction and concentrations of MeHg in the
hypolimnion of Little Rock Lake, WI, with a 10.8 ng MeHg/L increase for each mg S/L
reduced ("Watras et al.. 2006). There was also a positive relationship at this site between
total lake MeHg and annual SOx deposition, with a 14.4 mg MeHg/lake increase for
every 1 kg S/ha/yr increase in SOx deposition (Watras and Morrison. 2008).
There are multiple lines of evidence of a positive relationship between sulfate surface
water concentrations and MeHg concentration or production in multiple freshwater
systems. Both experimental manipulations and observational studies show that higher
sulfate concentrations in peat wetlands increase MeHg concentrations (Figure 12-18).
Experimental addition of sulfate to Bog Lake Fen in Minnesota showed that 8.3 kg S/ha
increased pore water [MeHg] (Mitchell et al.. 2008a). and 32 kg S/ha/yr increased pore
water [MeHg] as well as MeHg fraction of total Hg, or %MeHg (Jeremiason et al.. 2006).
In other bogs sampled in the same region of Minnesota and Ontario, [MeHg] in water
increased 0.037 ng/L for each 1 mg sulfate/L increase (Mitchell et al.. 2009). These
results were confirmed in a peatland S addition experiment in Sweden, which found that
20 kg S/ha/yr addition to the peat increased [MeHg] (Akerblom et al.. 2013; Bergman et
al.. 2012).
Recent work also established positive relationships between sulfate and MeHg in
ecosystems other than peat wetlands. In prairie potholes in the West (Hoggarth et al..
2015). with surface water MeHg increasing 0.3 ng/L for each additional mg sulfate/L
where conductivity and aromatic DOC remain constant (Hall et al.. 2009a). In a
freshwater marsh in the Everglades, water [MeHg] concentrations increased when sulfate
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concentrations exceeded 1 mg sulfate/L (Corrales et al.. 2011). which is also an important
threshold value for fish in the Everglades (see below). There are also recent
ecosystem-scale studies that show that sulfate reduction and MeHg can be correlated at
the landscape level. At the watershed level in the Adirondacks, [MeHg] in stream
outflows was negatively correlated with sulfate concentration (Selvendiran et al.. 2009).
with a 0.11 ng/L increase in stream [MeHg] for each mg sulfate/L removed from surface
water (Selvendiran et al.. 2008a).
Deposition thresholds for non-acidifying S effects
I 32 Increase in porewater %MeHg in MN bog
Increase in mosquito Hg load in MN bog
20 Increase in porewater [Hg] in MN bog
11.7 Increase in largemouth bass Hg load in TX reservoirs
5.3 Increase in porewater [MeHg] in MN bog
10	15	20	25	30	35
kg S/ha/yr
ha = hectare; Hg = mercury; kg = kilogram; MeHg = methylmercury; S = sulfur; yr = year.
Figure 12-18 Thresholds of sulfate addition or deposition from published
studies which affect chemical or biological changes in
ecosystems.
12.7.8. Sensitive Ecosystems
The 2008 ISA identified ecosystems in the Northeast as particularly sensitive to Hg
methylation in response to S deposition, because many watersheds there have abundant
wetlands and freshwater water bodies with high DOC and low pH. Recent analyses
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confirm that ecosystems east of the Mississippi tend to have higher Hg concentrations in
fish ("Wentz et al.. 2014; U.S. EPA. 2011b'). Recent papers suggest that reservoirs with
high water-level fluctuation or high algal productivity also have high methylmercury
production or bioaccumulation within fishes (Ecklev et al.. 2017; Noh et al.. 2016;
Ecklev et al.. 2015; Becker etal.. 2011). and that at the ecoregion level, reservoir fish Hg
concentrations correlate with both Hg and S deposition (Drenner et al.. 201IV
Recent research in the Yolo Bypass Wildlife Area, CA, indicates that Hg methylation
occurs in both permanent and agricultural wetlands and was higher in the agricultural
wetlands where rice is grown (Alpers et al.. 2014; Marvin-Dipasquale et al.. 2014). Hg
methylation in this system was associated with Mn reduction, and then to a lesser extent
with Fe reduction, and was only marginally associated with S reduction; sulfate addition
did not affect Hg methylation (Alpers et al.. 2014; Marvin-Dipasquale et al.. 2014).
However, MeHg produced in the sediments where rice grew was highly correlated with
MeHg concentrations in rice seeds ("Windham-Mvers et al.. 2014a). suggesting a route to
human exposure, and there were higher MeHg burdens in several animal species
collected from agricultural wetlands than from permanent marshes (see below).
Mercury Effects on Animal Species
Mercury is a developmental, neurological, endocrine, and reproductive toxin across
animal species. It is transformed by SRPs from inorganic Hg into methylmercury
(MeHg), which forms strong complexes with thiosulfate and sulfhydryl groups of organic
molecules. This bound MeHg accumulates in biota in successively higher concentrations
at ascending trophic levels. The 2008 ISA documented Hg accumulation in four turtle
species, songbirds, insectivorous passerines, and the common loon (Gavia immer). A
recent report by the USGS notes that Hg accumulation has also been documented in
insectivorous songbirds and bats (Wentz et al.. 2014). In recent research conducted in the
Yolo Bypass, CA, higher water [MeHg] in agricultural rice-producing wetlands than in
permanent marshes led to higher MeHg concentrations in mosquitofish (Gambusia
affinis) and Mississippi silversides (Menidia audens) from the agricultural wetland than
from the permanent marshes (Ackerman and Eagles-Smith. 2010).
The 2008 ISA reported that 23 states had issued fish advisories by 2007 in response to the
U.S. EPA's fish tissue criterion set to protect human health of 0.3 |ig MeHg/g fish, or
0.3 ppm. The 2008 ISA reported on the negative impacts of Hg on development,
morphology, survival, or reproduction in the following fish species: walleye (Stizostedion
vitreum), grayling (Thymallus thymallus), mummichog (Fundulus heteroclitus), rainbow
trout (Oncorhynchus mykiss), fathead minnows (Pimephales promelas), and zebrafish
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(Danio rerio). However, a recent report on Hg in streams of the U.S. by the USGS
summarizes current research that birds, fish, and fish-eating wildlife experience negative
effects of Hg at lower concentrations than the 0.3 ppm criteria ("Wentz et al.. 2014).
The 2008 ISA reviewed two studies that considered the link between S deposition and Hg
levels in fish (Drevnick et al.. 2007; Hrabik and Watras. 2002); both studies showed that
decreases in S deposition resulted in decreases in fish MeHg levels. Recent research
supports this finding in Voyageurs National Park (a Class I area) in Ryan Lake between
1998-2012 (Brigham et al.. 2014). when decreasing S deposition correlated with
decreasing fish Hg. In addition, a survey of fish caught in Texas reservoirs found that fish
in the highest deposition region (11.7 kg S/ha/yr) had significantly higher mean Hg levels
than fish from other regions (Drenner et al.. 2011). Experimental S addition to the
Marcell peat bog in Minnesota demonstrated that 32 kg S/ha/yr increased the Hg
concentrations in larval Culex spp. (mosquitoes), which are an important food source for
both aquatic and terrestrial species (Wasik et al.. 2012).
In addition to the studies that consider S deposition, there are recent studies that consider
sulfate concentrations in water in relation to fish Hg concentrations. A study of fish from
six lakes in South Dakota found a positive correlation between sulfate water
concentrations and walleye Hg concentrations (Stone etal.. 2011). The marshes of the
Everglades receive high S loading as agricultural runoff, and recent analysis of Hg loads
in mosquitofish, sunfish (Lepomis spp.), and largemouth bass (Micropterus salmoides)
collected from 1998-2009 showed that Hg levels in fish were highest when sulfate
concentrations were between 1 and 12 mg sulfate/L; the researchers proposed 1 mg
sulfate/L as a water standard (Gabriel et al.. 2014).
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SUPPLEMENTAL MATERIAL
MERCURY CYCLING
This supplemental material describes the deposition and cycling of mercury within North
American ecosystems. Connections to sulfur cycling are emphasized where such
information is available. This supplemental material provides background information
and context for the discussion of nonacidifying sulfur stimulation of mercury methylation
in wetland and aquatic ecosystems (Appendix 12).
Transfer of Mercury from the Atmosphere to Terrestrial Ecosystems
Mercury enters most natural terrestrial and aquatic ecosystems in the U.S. via
atmospheric deposition, although both active and historic mining and industrial sites have
contaminated soil and water in many ecosystems. Sources and atmospheric processes
governing Hg deposition are covered more exhaustively elsewhere ("Wentz et al.. 2014;
Pirrone et al.. 2010). For the purposes of this section, we will focus on the deposition of
reactive Hg, which is deposited into ecosystems in the form of free Hg(II),
particulate-bound Hg, or as MeHg. In a jack pine/birch forest in the ELA sampled in
1998-1999, wet deposition was 71,000,000 ng/ha/yr (reported as 71 mg total Hg/ha/yr),
with dry deposition estimated at 9,000,000 ng/ha/yr [reported as 9 mg/ha/yr in (St Louis
etal.. 2001)1.
Foliage is an important Hg pool in both forested and wetland ecosystems because it can
directly absorb Hg deposition as well as store Hg taken up by plant roots. In the Arbutus
Lake watershed in the Adirondack Mountains, Hg in the AI mis and Betula overstory was
3,080 ng Hg/m2 (3.08 |ig Hg/m2), and was 9,480 ng/m2 (9.48 (ig/m2) in the Sphagnum
mat of the wetlands (Selvendiran et al.. 2008b). In a 3-year-long Hg tracer isotope annual
addition experiment at ELA, upland canopy and ground vegetation retained 40% of
added Hg. In a wetland ecosystem, vegetation retained 80% of added Hg (Harris et al..
2007). Although the wetland ecosystems exported Hg and MeHg to an adjacent lake in
the watershed during the course of the study, the Hg was from sources existing before the
start of the experiment, indicating that Hg storage in wetland vegetation may have a
residence time greater than 1 year (Harris et al.. 2007).
Foliage also represents an important flux of Hg; following senescence, leaf litter becomes
the substrate for microbial decomposition, and Hg moves out of plant biomass and into
microbial biomass, soil, or water. Flux of total Hg from litter in the ELA, Ontario, was
twice that of wet deposition in litter from upland trees and was of a similar magnitude to
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deposition in litter from wetland shrubs and trees [72,000,000 ng/ha/yr or 72 mg
Hg/ha/yr, and 88,000,000 ng/ha/yr or 88 mg/ha/yr; (St Louis et al.. 2001)1. The organic
horizon of the soil, where litter and microbial communities are concentrated, was a major
sink of Hg within the forested and wetland ecosystems for the 20 years preceding this
study, with annual Hg accumulation rates of 130,000,000 ng Hg/ha/yr (reported as
130 mg Hg/ha/yr) for ridgetop soils (160% of total annual deposition),
200,000,000 ng/ha/yr (200 mg/ha/yr) for upland forest (250% of deposition), and
590,000,000 ng/ha/yr (590 mg/ha/yr) in wet soils supporting Sphagnum mosses
[7.4 times annual deposition; (St Louis et al.. 2001)1. In forest catchments, much of the
Hg deposition is retained in soil (St Louis et al.. 2001; St. Louis et al.. 1996). as
confirmed by the Hg tracer isotope addition experiment in ELA, which found that after
3 years of Hg addition, 58% of the added Hg was retained in soil in the upland forest
catchment (Harris et al.. 2007). The length of residence within the soil for Hg likely
depends on the same factors that control the storage of organic matter, such as soil
structure, temperature, precipitation, and timing of disturbances, particularly of wildfire.
Transfer of Mercury from Terrestrial to Aquatic Ecosystems
When Hg is released from plant tissue, detritus, or soil particles into surface water or the
soil solution, it becomes available for transport into aquatic systems. In five boreal forest
catchments in the ELA, Ontario, measured during the period 1990-1993, the flux of total
Hg from the forest into water bodies was estimated to be
14,000,000-21,000,000 ng/ha/yr (reported as 14-21 mg/ha/yr 28-42% of annual Hg
input), in response to approximately 47,000,000 ng/ha/yr (47 mg/ha/yr) wet + dry
deposition of Hg, and estimated contributions from weathering of granite of as much as
3,000,000 ng/ha/yr [reported as 3 mg/ha/yr in St. Louis et al. (1996)1. On Isle Royale, a
large island in Lake Superior and a Class I Area, researchers estimated that lakes received
one-third of their annual Hg load directly from atmospheric deposition, and the remaining
two-thirds of annual Hg load was mobilized from soils in the watersheds (Drevnick et al..
2007). In aquatic environments, specialized microbes (SRPs; as well as iron-reducing
bacteria) in anoxic zones can methylate inorganic Hg, transforming it into the toxic,
persistent MeHg.
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Transfer of Mercury from Atmosphere to Aquatic Ecosystems
Total Hg deposited directly to lakes is rapidly incorporated into Hg cycling in abiotic and
biotic pools. In a Hg isotope addition experiment at the ELA, Ontario, a spike of isotope
202Hg was added directly to the lake annually over the course of 3 years. The 202Hg was
detected in the anoxic hypolimnion within days of addition, and in bottom sediments of
the lake within 4 weeks (Harris et al.. 2007). and methylated 202Hg was detected in
zooplankton, benthic amphipod Hyallela Azteca, and fish species within 2 months of
addition (Harris et al.. 2007). indicating that Hg methylation and movement of MeHg
within the food web can occur within a single season in aquatic ecosystems. Estimated
residence time of atmospherically deposited Hg2+ in the water column of Little Rock
Lake, WI, was 150 days ("Watras et al.. 2006).
Methylmercury Cycling
Methylmercury (MeHg) is produced by sulfur-reducing prokaryotes (SRPs) in the course
of sulfate reduction. It is toxic to humans and vertebrate animals, particularly because it
biomagnifies up the food chain and bioaccumulates within organisms. The following
sections describe transfer of MeHg between ecosystem compartments, as well as
characteristics of zones of high MeHg production within the landscape.
Transfer of Methylmercury from Terrestrial to Aquatic Ecosystems
MeHg is a very small component (1%) of Hg atmospheric deposition to ecosystems
(Wcntz et al.. 2014); most MeHg in ecosystems is microbially produced from inorganic
Hg. Much of our knowledge about Hg cycling in natural ecosystems comes from the
Experimental Lakes Area (ELA) in northwestern Ontario, Canada, which receives low
Hg deposition [47,000,000 ng/ha/yr or 47 mg/ha/yr in 1990-1993; 80,000,000 ng/ha/yr
or 80 mg/ha/yr in 1998-1999 (St Louis et al.. 2001; St. Louis et al.. 1996)1. A small
portion of total Hg deposition in the ELA is in the form of MeHg, 900,000 ng/ha/yr
(reported as 0.9 mg/ha/yr, or <1.5% of total Hg deposition). MeHg cycles through the
terrestrial ecosystem as total Hg does. Leaf litter represents a significant flux of MeHg
from the living canopy in upland forest to the detrital food web of 800,000 ng
MeHg/ha/yr (reported as 0.8 mg MeHg/ha/yr), a flux of a similar magnitude to
atmospheric deposition (St Louis et al.. 2001). In the 20 years preceding the study,
600,000 ng MeHg/ha/yr (reported as 0.6 mg/ha/yr of MeHg) were annually stored in the
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organic soil pool, and on average 100,000 ng MeHg/ha/yr (reported as 0.1 mg
MeHg/ha/yr) was exported from the catchment in stream flow (St Louis et al.. 2001).
In five boreal forest catchments at the ELA, St. Louis et al. (1996) assumed a background
MeHg deposition rate of <500,000 ng MeHg/ha/yr (reported as <0.5 mg MeHg/ha/yr). A
catchment consisting only of upland forests exported 18% of estimated annual MeHg
deposition, indicating that terrestrial ecosystems retain MeHg (St. Louis et al.. 1996). and
that aquatic MeHg is produced endogenously.
Transfer of Methylmercury from Wetlands to Aquatic Ecosystems
At the landscape level, MeHg export from wetlands in streams and rivers is correlated
with wetland cover in watersheds. In the ELA in Ontario, a catchment with valley bottom
wetlands exported 73% of deposited MeHg in a dry year, but 158-200% of deposited
MeHg in a wet year, indicating that the wetlands produced MeHg in wet years. MeHg
export was high in catchments with other types of wetlands as well. A catchment with
riverine wetland (peatland fed by surface water flow) exported 120-130% of annually
deposited MeHg in dry and wet years, and a catchment containing a basin wetland
exported 213% MeHg in a dry year and 500-719% annually deposited MeHg in wet
years, indicating a strong source of MeHg within the watershed, presumably in the
wetlands (St. Louis et al.. 1996). More recent work in the watershed of Arbutus Lake, NY
measured Hg and MeHg in streams draining upland and wetland areas; MeHg was
1.7-3.0% of total Hg in upland streams and 7.8% of total Hg in a similarly low-order
stream that drained a wetland. In downstream peatlands, MeHg was 5.4-9.5% of total Hg
(Selvendiran et al.. 2008a). and the beaver meadow peatland exported 623% of total
MeHg input annually, while the riparian wetland exported 8% of total MeHg input
(Selvendiran et al.. 2008b). Wetlands with high dissolved organic carbon and fluctuating
water levels and anoxic zones are areas of high MeHg production within watersheds, and
such wetlands will affect MeHg concentrations in water and biota downstream.
Methylmercury Cycling in Aquatic Ecosystems—Lake Onondaga, New York
Lake Onondaga, NY was the subject of a Hg budget project in 1992 because the
eutrophic lake received high Hg loads through the late 1970s from industry
[Supplemental Figure 1; (Henry etal.. 1995)1.
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Net Methylalion
(0, 0.63)
Net Uptake by Fishes
(0.20, 0.20)
Gross
Sedimentation
(13.7, 0.73)
Dissolved Flux
(0.056, 0.017)
Remineralization (2.6, 0.25)
Net Sedimentation
(11.3, 0.62)
Tributaries Atmospheric
and Metro Deposition Volatilization Outlet
(13 6, 0,26) (0,44. 0.006) (0.016, 0) (2,8, 0.24)
^ Groundwater
Inflow
(0.02. 0.001)
Hg = mercury; kg = kilogram: MeHg = methylmercury; yr = year.
Source: from Henry et al. (1995).
Supplemental Figure 1 Mercury and methylmercury mass balance cycle in Lake
Onondaga in 1992. The quantities of mercury in each flux
are indicated in parentheses by (kg total Hg/yr, kg
MeHg/yr).
In 1992, .1.4.1 kg of total Hg entered the lake, 96% of which was carried into the lake by
tributaries, including one that drained a former chlor-alkali plant. The largest sink for Hg
within the lake was through sedimentation, which accounted for 11.1 kg of total Hg. In
that same year, 2.8 kg of total Hg flowed out of the lake downstream and 0.20 kg Hg
(0.1% of total annual Hg load) was incorporated into fish biomass (Henry et al.. 1995).
A budget for Hg methylation was also constructed for the lake, and differs from that of
total Hg in several important points (Figure 12-4). First, only 29% of the annual lake
MeHg load of 0.88 kg entered the lake via tributaries (compared to nearly the entire total
Hg load). The water column, particularly deep, anoxic water, contributed 0.60 kg or 68%
of the annual MeHg load of the lake (Henry etal.. 1995). More than half (about 0.47 kg)
of the annual MeHg load was incorporated into sediments, and while sedimentation rates
were high, about one- to two-thirds of MeHg in sediments was released back into the
water column. Because most of the total Hg in fish is in the form of MeHg, the model
calculated that the MeHg in fish biomass was the same amount as total Hg in fish
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biomass, 0.20 kg, although this value constituted a much higher portion of the annual
MeHg load [23% of annual MeHg; (Henry etal.. 1995)1. This early study showed that the
majority of MeHg in Lake Onondaga was produced within the lake itself, specifically in
the anoxic zone, and that MeHg cycled between the sediments and water column, with a
significant amount stored in fish tissue.
Lake Onondaga was resampled in 2006, 2007, and 2009, following industrial remediation
projects and incorporation of nitrification treatment in the municipal wastewater plant
discharging into the lake (Todorova et al.. 2014). In the 2006-2009 sampling, MeHg
concentrations in the epilimnion were significantly lower than in 1992, and total Hg in
the epilimnion was half of the total Hg concentrations measured in 1992 (Todorova et al..
2014). However, in all years there was a significant peak in MeHg concentrations and in
MeHg:tHG ratios in the epilimnion in the fall following turnover (Todorova et al.. 2014).
indicating that this seasonal event will have significant effects on total MeHg in the food
chain of the lake.
Methylmercury in Sediments and Water Column—Lakes
The sediments under shallow, oxygenated lake waters are an important site of Hg
methylation. Early literature showed that Hg methylation rates are higher in shallow
organic sediments than in deep clay-rich sediments in Southern Indian Lake in Manitoba,
Canada, and that the ratio of methylation:demethylation was also higher in organic
sediments (Ramlal et al.. 1986). In the Quabbin Reservoir, MA, MeHg fractions in
sediments were 3.1-16.3% at 1-m depth, and decreased with increasing lake depth to
0.1-0.2% at a 22 to 23-m depth (Gilmour et al.. 1992). At Lake Clara, WI, Hg
methylation rates were sampled in the water column, flocculant surface sediments, and
deeper sediments at water depths of 1-10.5 m (Korthals and Winfrey. 1987). At all
sampling depths, methylation and demethylation were detectable in the water column and
in the surface sediments. Gross Hg methylation rates were highest in surficial sediments
sampled at depths of 5-7.5 m, but the methylation:demthylation ratio was highest
(5.5-5.8) in surficial sediments at depths of 1-2 m (Korthals and Winfrey. 1987).
Larger, deeper lakes stratify during the summer; waters near the surface (epilimnion) of
the lake are high in oxygen and primary productivity, while deeper waters (hypolimnion)
have low oxygen, lower primary productivity, and high rates of S reduction and MeHg
production. In the experimentally acidified Little Rock Lake, WI, MeHg was low (less
than 0.1 ng/L) in surface waters during the summer months, but increased at anoxic
depths below 4-6 m to 1 ng/L in June, 3 ng/L in July, and 3.5 ng/L in August (Bloom et
al.. 1991). In stratified Lake Onondaga, NY, the water column, particularly deep, anoxic
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water, contributed 0.60 kg or 68% of the annual MeHg load of the lake (Henry et al..
1995). Sampling in Pallette Lake, WI, showed that the oxic-anoxic boundary was at 12 m
depth in this stratified lake, which was also the location of the sulfate-sulfide transition
(Watras et al.. 1995). There were peaks in both sulfate reduction and Hg methylation
from 12-13 m depth, right at the top of the hypolimnion, which was enriched in MeHg
(20-30%) compared to the epilimnion (5-10% MeHg). There was no sulfate reduction
detected in the profundal sediments or deep layers of the hypolimnion. MeHg was
transported from the oxic-anoxic boundary and hypolimnion where it was produced to the
epilimnion, and the flux of MeHg from the hypolimnion (3,400,000-6,800,000 ng/day,
reported as 3.4-6.8 mg/day) was much larger than the flux of MeHg into the epilimnion
from the atmosphere [10,500 ng MeHg/day, or 0.0105 mg MeHg/day; (Watras et al..
1995)1. Ecklev and Hintelmann (2006) showed that in boreal Canadian lakes, potential
methylation is highest approximately 1 m below the oxycline, and this methylation zone
moves up the water column as the oxycline rises due to oxygen depletion at depth in the
summer (Ecklev and Hintelmann. 2006).
Methylmercury in Sediments and Water Column—Wetlands
In wetlands at Bog Lake Fen in the Marcell Experimental Forest, MN, the highest levels
of MeHg production occurred at the upland-peatland interface. When mesocosms
installed in the bog interior received additions of labile C and sulfate, MeHg
concentrations were similar to MeHg concentrations from the upland-peatland interface,
indicating that upland inputs of C may contribute to high methylation rates in the
peatland (Mitchell et al.. 2008a). A larger study that sampled peatlands within the
Marcell Experimental Forest, MN, as well as at the ELA in Ontario found that %MeHg
was highest within 5-10 m of the upland-peatland interface across wetlands, and that
sulfate and pH were higher at the interface, and DOC was lower at the interface, than in
the interior of the peatlands (Mitchell et al.. 2008a). In peatlands in the Arbutus Lake
watershed in the Adirondack Mountains, NY, MeHg and total Hg were higher in pore
water near the top of the peat profile, at 20 to 40-cm depth, than at 80 to 100-cm depth
(Selvendiran et al.. 2008a). and the total Hg and MeHg were highest in both peat and
pore water in the top 15 cm of the peat (Selvendiran et al.. 2008a). The Hg in the top
15 cm of peat was estimated to be about 27% of the total Hg, and MeHg in the top 15 cm
was 30% of the total MeHg in a riparian peatland and 50% of the total MeHg in a beaver
meadow (Selvendiran et al.. 2008a). In freshwater Sphagnum peatlands of the Allequash
Creek watershed in the Northern Highland Lake District of Wisconsin, reducing
conditions exist below 2-cm depth in the peat, and MeHg concentrations peak at 7-cm
depth in the peat (Creswell et al.. 2008).
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In wetlands not dominated by Sphagnum, Hg methylation occurs in sediments and in
periphyton (see Appendix 12.3). In the freshwater marshes of the Florida Everglades,
where periphyton and flocculation often extend the sediment-water interface, methylation
rates were highest in the top 6 cm of sediments (Gilmour et al.. 1998). In coastal marsh
sediments, such as at Kirkpatrick Marsh on the Chesapeake Bay, MeHg was highest in
the top 6 cm of sediment, although total Hg was highest at a depth of 12-15 cm in the
sediments (Mitchell and Gilmour. 2008). and in coastal Georgia marshes, sediment core
incubations showed that MeHg production was highest in the top 4 cm of the marsh
sediments (King et al.. 1999).
Methylmercury in Sediments and Water Column—Estuarine and Marine
Ecosystems
Research shows that Hg methylation occurs in estuaries and in marine sediments.
Incubations of samples collected in the Patuxent River and Estuary, MD, showed that net
MeHg production in sediments was four times the rate of MeHg accumulation in
sediments, suggesting that MeHg in estuary surface waters comes from estuarine
sediment sources as well as sources upstream in the watershed (Benoit et al.. 1998).
MeHg comprised 5% of surface water particulate Hg (unclassified compounds large
enough to be removed by filtration) and 2% of total Hg in surface water, and 0.3% of Hg
in estuarine sediments (Benoit et al.. 1998). More recently, surface sediments were
sampled from the Chesapeake Bay, from four locations in the main channel, two
locations on the continental shelf, and one location on the slope of the continental shelf.
Total Hg in the upper bay ranged from 50 to 171 ng Hg/g sediment (reported as
250-850 pmol/g), in the lower bay and continental shelf ranged from 2.0 to 20 ng Hg/g
sediment (10-100 pmol/g), and at the single continental slope site was 50-70 ng Hg/g
sediment (250-350 pmol Hg/g); (Hollweg et al.. 2009). MeHg was 0.2-1.5% in the top
0-12 cm of sediments across sites, and bottom water salinity was strongly correlated with
%MeHg [r = 0.814; (Hollweg et al.. 2009)1. suggesting that MeHg was produced in
marine sediments.
An incubation study of water samples collected at depths of 20-327 m in the ocean
around the Canadian Arctic Archipelago found measurable rates of Hg methylation at all
depths from all five sites sampled. Methylation of inorganic Hg accounted for 47% of the
MeHg present in the water column at these locations, and demethylation was also widely
observed, indicating that marine MeHg represents marine production of MeHg, not
merely transport of MeHg from other systems (Lehnherr et al.. 2011).
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APPENDIX 13. CLIMATE MODIFICATION OF
ECOSYSTEM RESPONSE TO
NITROGEN AND SULFUR
The scope of this appendix is to identify key papers describing how climate alters
ecosystem response to nitrogen (N) and sulfur (S) addition. Nitrogen and S loading
occurs in many ecosystems concurrently experiencing multiple stressors, including
human-driven climate change. Climate change effects on U.S. ecosystems were recently
summarized in the U.S. National Climate Assessment (Galloway et al.. 2014; Groffman
ct al.. 2014). Each appendix of the ISA evaluating the effects of N enrichment or
acidification includes a section on how climate modifies the ecosystem response to N or
N + S deposition. Additionally, in this appendix, we have excerpted text from Greaver et
al. (2016). to serve as a foundation for the discussion. This paper, which focuses on
empirical observations, provides a current review of how climate (e.g., temperature and
precipitation) modifies ecosystem response to N.
Anthropogenic emissions of greenhouse gases are likely to cause a global average
temperature increase of 1.5 to 4.0°C and a significant shift in the amount and distribution
of precipitation by the end of the 21st century (Collins et al.. 2014). Recent work has
focused on the effects of anthropogenic N on the Earth's radiative forcing (Pinder et al..
2012) and how temperature and precipitation alter ecological responses to N exposure
(Greaver et al.. 2016). Most work is conducted on the effects of climate and N or
acidifying deposition (N + S). There are only a couple of studies on how climate modifies
ecosystem response to S.
Climate effects on ecosystems is a rapidly expanding field. For some processes we are
beginning to understand how temperature and precipitation may interact; however, for
many biogeochemical pool and processes, data are insufficient to quantify either the
direction or magnitude of how climate may alter ecosystem response to N with certainty.
Some global-scale earth systems models now incorporate interactions between N and
carbon (C) in ecosystems. They are summarized in Appendix 6. Greaver et al. (2016)
includes information on terrestrial and surface water ecosystems. A brief summary of
climate modification of estuary response to N is also included in Appendix 13.2.
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13.1. Climate Modification of Soil Acidification and Nitrogen
Enrichment
Climate-driven changes in N cycling may alter the N supply in terms of quantity and
timing of N available for uptake by biota. Alteration of N availability relative to a given
organism's life cycle or physiological thresholds may alter overall ecosystem function.
These effects may be further modified when temperature and precipitation cause direct
stress to biota. In the following sections we describe how climate alters the cycling of N
and how temperature and precipitation interact with two important mechanisms affected
by N availability to taxa: (1) nitrogen-driven eutrophication, which will stimulate the
growth of opportunistic plant and animal species, and (2) acidification, which may
decrease growth and cause mortality among sensitive species. We describe how these
changes in growth will alter C cycling and biodiversity.
13.1.1. Nitrogen Transport and Transformation
13.1.1.1. Excerpt from Greaver et al. (2016)
"Although global N cycling is complex, the movement of N through the biosphere can
largely be explained by describing a few key transformations (Figure 13-1). Atmospheric
N2 is converted into reactive N by lightning or by specialized bacteria capable of
biological N fixation, in addition to the human activities that create reactive N.
Organisms use reactive N to produce proteins and other essential compounds. Dead
organic matter is decomposed by microbial enzymes, producing smaller N containing
organic molecules such as amino acids. This organic N is largely converted to mineral
forms that are readily assimilated by plants and microorganisms. Where reactive N is
present under aerobic conditions, some microorganisms convert ammonium (NH/) to
nitrate (NO, ). a process termed nitrification. Nitrate is mobile in soils and often leaches
into aquatic systems and groundwater. In anaerobic conditions, microorganisms can
convert NO3 to gaseous N via denitrification, emitting N back to the atmosphere.
"In nonagricultural terrestrial ecosystems, atmospheric deposition is the dominant source
of anthropogenic N. Through changes in precipitation, shifts in atmospheric circulation,
and temperature-related effects on the stability of N compounds, climate change is
expected to alter the relative contribution of wet and dry forms of N deposition and shift
the spatial distribution of N deposition (Engardt and Langncr. 2013; Tagaris et al.. 2008).
Local N deposition rates are generally predicted to change by 0-20% (Engardt and
Langner. 2013; Tagaris et al.. 2008).
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Spatial and temporal alteration of snow melt, precipitation and evapotranspiration
I
Spatial and temporal alteration of landscape-level wetness and hydrologic flow
i
Key mechanisms of nitrogen cycling altered due to changes in hydrology
Drier conditions
(-) Nitrogen flushing is likely to
increase N accumulation in
ecosystems, with large N pulses
exported during rainfall events.
(-) Dentrification will generally
decrease with drying, leading to
accumulation of N within the
ecosystem.
(-) Mineralization will generally
decrease under dry conditions.
(-) Nitrogen uptake by vegetation
caused by drought-stress in
vegetation as water becomes the
most limiting factor for growth.
(+) Drought-related plant infestation
and disease wilt generally decrease
under dry conditions.
(+) Fire will release ecosystem N into
the air and increase N available as
throughflow.
N = nitrogen.
Source Greaver et al. (2016).
Figure 13-1 Summary of key interactions between nitrogen,
anthropogenic-driven climate change, and hydrology.
1	"The influence of N addition on reactive N availability within terrestrial and aquatic
2	ecosystems is mediated by microbial transformations and transport within and between
3	ecosystems. Climate change is expected to strongly affect these processes through
4	increasing temperature and through temporal and spatial shifts in temperature and
5	precipitation (Baron et al.. 2012). Warming directly increases the metabolic biokinetics of
6	enzyme activity necessary for microbial N transformation until a temperature optima is
7	reached. Climate change is also expected to cause numerous modifications of the
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Wetter conditions
(+) Nitrogen flushing increases export
from 'upstream' ecosystem and
loading to the 'downstream'
ecosystem.
(+) Denitrification will generally occur
in wetter areas, increased
denitrification will cause more N in
the ecosystem to be lost via the
gas-phase.
(+) Mineralization will generally
increase under wet conditions.
(+) Moisture-related plant infestation
and disease will generally increase
under wet conditions.
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hydrologic cycle, and moisture availability regulates the biokinetic temperature response
(Borken and Matzner. 2009; Rustad et al.. 2001) because water is needed to transport
enzymes and substrates and water influences oxygen availability. Warming is predicted
to intensify the hydrologic cycle, with heavy rainfall events expected to become more
frequent and intense, along with the potential for deepening and lengthening of dry
periods, as well as altered snow accumulation and melt and changes in evapotranspiration
(Collins et al.. 2014; Jimenez Cisneros et al.. 2014). These changes may cause soil
conditions for microbial activity to shift between optimal and inhibitory (Morse et al..
2015b). modifying the link between climate warming and the rate of microbial N
transformations such as decomposition, mineralization, nitrification, denitrification, and
biological N fixation. Of these transformations, the rates of N fixation may be the most
uncertain part of the N cycle (Vitousck et al.. 2013). altering N supply and influencing
the C cycle (Welter et al.. 2015).
"Climate-driven changes to the hydrologic cycle will also alter the quantity of N
transported through a system via waterborne transport and soil water content-mediated N
cycling (Billen et al.. 2013). Alteration of N retention in the soil due to changes in
moisture and flushing may be significant enough to determine whether an ecosystem is
an N source versus N sink (Boulton. 2007; Band et al.. 2001). Greater precipitation
generally increases water flow, which may (1) increase leaching/export of N through
terrestrial landscapes, (2) increase terrestrial N inputs to streams and rivers, and
(3) increase N transport rates through streams and rivers (Whitehead et al.. 2009).
although adaptation by microbes and plants may increase their ability to retain N as
flushing increases, thereby potentially limiting some of the overall impact of increased
precipitation and flow.
"Under dry conditions, landscapes can become more hydrologically disconnected and N
retentive, which can increase N concentrations in subsequent flushing events (Goodridgc
and Melack. 2012; Kaushal et al.. 2008). Drought can inhibit nitrification and cause N to
accumulate in the soil; once precipitation occurs it often results in a pulse of nitrification
that produces nitrate and subsequent nitrate leaching (Lamersdorf et al.. 1998). For
example, the 2012 droughts in the midwestern U.S. were followed during the spring 2013
by extremely high river nitrate concentrations (Beeman. 2013). Likewise, longer periods
between wet cycles lead to accumulation of nitrate and other acidifying solutes in the
soil, causing less frequent, yet more extreme acidification events (Bavlev et al.. 1992).
Beyond simply the total volume, precipitation intensity influences the rate of N flow
through ecosystems. Increased precipitation intensity of cold season frontal storm
systems and warm season convective storms would likely increase the frequency of high
N loading events to aquatic systems. Due to the limited capacity for in-stream removal of
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N during high flow pulse events, most N is transported downstream (Kaushal et al..
2014V
"Such hydrologic cycle changes are also expected to affect the timing of N transport.
Changes in the seasonality of precipitation, and in particular snowmelt, will tend to alter
the timing of N flushing through the ecosystem. There has already been widespread
earlier snowmelt and increased winter thaws associated with warming over the last few
decades (Collins et al.. 2014). but implications for the timing of N export have been
assessed in only a few sites (Casson et al.. 2012). Timing changes ultimately can alter the
magnitude of N export to downstream water bodies, particularly if the timing of flushing
changes relative to the timing of biologically mediated uptake in either terrestrial or
aquatic ecosystems (Baron et al.. 2009). Thus, it is possible to have a modification in
sink/source behavior in regions where annual or seasonal patterns of water-filled pore
space shift with climate change.
"The rate at which denitrification returns reactive N to the atmosphere varies across space
and time, with landscape- to micro-scale denitrification hot spots or moments that depend
on interactions with hydrologic flow paths, the persistence and variability of low oxygen
conditions, and the residence time of water and N, all of which are likely to respond to
climate-driven changes in the hydrologic cycle (Anderson et al.. 2015; Weier et al..
1993). Generally, warmer and wetter conditions under climate change would facilitate
greater rates of denitrification, whereas warmer and drier areas might experience
decreased denitrification or concentrate "hot spots" into smaller areas with higher soil
moisture, substrate concentrations, and fluxes (Duncan et al.. 2013). Alternating wet and
dry states may promote coupled nitrification/denitrification processes or build-up and
flushing of mobile N depending on the ratio of transport to reaction rates. These are
general trends associated with moisture availability and transport. Carbon substrate
availability and other controls on microbial processes, however, will also be influenced
by climate change (discussed below) and can mediate these hydrologic effects."
13.1.1.2. Additional Considerations
There are considerations to note in addition to the excerpt from Greaver et al. (2016).
Additional papers on climate interactions with N and S soil biogeochemistry discussed in
Appendix 4 are summarized in Table 13-1. Featured most notably are the effects of
precipitation on S cycling, snow on soil N cycling, N addition effects on biogenic GHG
flux, and the results of integrating climate parameters into N and S biogeochemistry
modeling.
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Table 13-1 Summary on climate modification sulfur and nitrogen cycling in
Appendix 4 and Appendix 6 in addition to those in Appendix 13.
Indicator/Process T P Snow
Reference
Biogeochemical effects (Appendix 4)
S mineralization X X	Empirical: Watershed mass balances of sulfur are
increasingly regulated by watershed moisture
because high-moisture soil conditions stimulate the
net mineralization of soil sulfur pools.
Mitchell etal. (2011)
S retention
X	Model: Watersheds with higher runoff ratio tend to
convert sooner from net retention to net release of
SCM2".
Rice etal. (2014)
Soil N variables
X Meta-analysis: Meta-analysis data is from
41 publications based on snow depth manipulation
experiments. The general responses of 12 variables
related to terrestrial nitrogen (N) pools and dynamics
to altered snowpack depth are evaluated.
Li etal. (2016c)
N mineralization X
X Empirical: More freeze/thaw cycles anticipated with
climate warming; supported lower rates of N
mineralization.
Duran et al. (2016)
N mineralization X
X Empirical: Snow-manipulation experiment at
Hubbard Brook to evaluate soil freezing later in the
season due to prolonged warm season. Freezing
likely increases soil NO3" levels by physical disruption
(increased fine root mortality) causing reduced N
uptake by plants and reduced competition for
inorganic N, allowing soil NO3" levels to increase
even with no increase in net mineralization or
nitrification.
Groffman et al.
(2001)
N mineralization
X Empirical: N mineralization rates were more strongly Sorensen et al.
related to soil volumetric water content than to root (2016)
biomass, snow or soil frost, or winter soil
temperature.
N effects CO2, CH4,
and N2O
Meta-analysis: 313 observations across 109 studies Liu and Greaver
evaluated the effect of N addition on the flux of three (2009)
major GHGs: CO2, CH4, and N2O.
Soil acidification and X X	Model: 2009-2100—six climate change simulations Pourmokhtarian et
N enrichment	of temperature, precipitation, and photosynthesis al. (2012)
active radiation (PAR). Without CO2 fertilization
effects, net soil mineralization and nitrification
increased, soil and stream acidified, and the
percentage of base saturation in soils declined due to
increased NO3". With CO2 fertilization effects, the N
loss to streams was suppressed due to increased
plant uptake.
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Table 13-1 (Continued): Summary on climate modification of nitrogen (N) effects
on nitrogen cycling in Appendix 4 and Appendix 6 in
addition to those in Appendix 13.
Indicator/Process T P Snow	Reference
Soil acidification, N XX	Model: The ForSAFE-VEG model simulations show Belvazid et al
enrichment, and	that climate and atmospheric deposition have	(2011a)
plant community	comparably important effects on N mobilization in the
soil, as climate triggers the release of organically
bound nitrogen stored in the soil during the elevated
deposition period. Climate has the most important
effect on plant community composition; thus climate
change cannot be ignored in future simulations of
vegetation dynamics.
Soil acidification and X X	Model: PnET-BGC modeling found climate change Wu and Driscoll
N enrichment	played a larger role in ANC than base cation	(2010)
deposition changes. Temperature had a larger effect
than precipitation on decomposition. Net
mineralization and nitrification increased faster with
climate change than plant NO3" uptake.
13.1.2. Nitrogen, Climate, and Carbon Cycling
13.1.2.1. Excerpt from Greaver et al. (2016)
"Changes in N supply alter plant growth and C cycling. The addition of N to terrestrial
and aquatic ecosystems can cause eutrophication, a state of high nutrient availability that
alters ecosystem function. Autotrophs (plant/algae) capture CO2 through photosynthesis,
storing C in biomass until it is oxidized through respiration or combustion and released
back to the atmosphere. In terrestrial systems, N addition usually stimulates autotrophic
growth until biotic N demand has been satisfied, although high rates of N addition may
increase the concentration of acid anions, which often decreases plant growth (see
acidification discussion below). At certain sites, N additions have uneven effects,
stimulating growth of some tree species, while impairing the health and growth of others
(Thomas et al.. 2010). When considering the net effects on multiple tree species, growth
in most forests is stimulated. This additional growth increases the overall amount of C
stored in plant biomass; one unit of N input may cause an additional 24.5 to 177 units of
forest C uptake (dc Vries et al.. 2014a; Liu and Greaver. 2009).
"A number of published meta-analyses evaluate the response of C pools and fluxes to
single stressors of N, precipitation, or temperature. To gain insight on stressor
interactions from single stressor response studies, we have synthesized existing
meta-analyses of terrestrial ecosystem response to additions ofN, precipitation, and
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temperature (e.g., [see Figure 13-2 in this document]). A recent correlation analysis of
growth (in terms of net primary production |NPP|) for 1,247 woody plant communities
across global climate gradients confirms that NPP increases with higher temperature and
precipitation (Chu et al.. 2016). However, ecological changes along broad natural
gradients may differ from the response of ecosystems to comparatively rapid
environmental change caused by human activities (Chu et al.. 2016). This latter process
may be more accurately characterized by addition experiments, such as those summarized
by meta-analysis. Our synthesis of existing meta-analyses indicates that aboveground
NPP is highly responsive to N addition and enhanced precipitation, while temperature
rise does not increase aboveground NPP. This result is consistent with the basic
biokinetic effects of warming on enzyme activity, which would have the counteracting
effects of stimulating both plant C capture (photosynthesis) and plant C release
(autotrophic respiration). Although there are no meta-analyses on warming and
whole-ecosystem autotrophic respiration, Lu et al. (2013 i observed in a meta-analysis
that temperature increased gross ecosystem production, a metric that does not subtract
respiration from gross production (photosynthesis). Therefore, temperature may have
larger effects on plant C fluxes than on NPP. Note that precipitation-induced changes in
NPP may vary depending on whether there is sufficient enhancement of precipitation to
offset increased evapotranspiration in a wanner climate (Bachclct et al.. 2001).
Carbon pools
Ecosystem C
31 ? ?
60
30
0
-30
N T P
GEP

NEE


R
? 34 48

31 34
48

? 34 48

60-

—i—
60

mR
30-
NS NS

30-



i—r	




-30-1


-30-1

Above-ground biomass
37 34 48
60-
30
0
-30
Foliage biomass
34 ? ?
60-
30-
0-
-30
Root biomass
37 34 48
60-
30-
0
-30
a
Fine root biomass
? 99 ?
60-
NS

N T P
60
30-
0
-30
ŁL
N T P
60
30-
0
-30
NS
Above-ground NPP Below-ground NPP Below-ground RautaJrC(lf1s
97 48 48	? 48 48	? 34 100
60-
30-
0
-30
N T P
Organic layer
36 34 ?
Soil C Dissolved organic C Microbial C	Myrcorrhiza
36 34 ?	36 34 ?	36 34 ?	38 ? ?
Decomposition
98 34 ?
60-
30-
0-
-30-
1 NS
EHi—i
60-
30-
0
-30
H2	
N T P
60-

60
30-

30

NS,		

-30-

-30
36 34 100
h«
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Note: Bars indicate response ratios (treatment/control x 100), and error bars represent 95% confidence intervals for the response.
Orange bars indicate the magnitude of response to nitrogen enrichment, blue bars show response to temperature increase, white
bars show response to precipitation increase. Blue zone indicates ecosystem carbon inputs and outputs, green zone indicates plant
responses, and brown zone indicates soil and microbial responses. The upper confidence interval for the precipitation effect on net
ecosystem C exchange is 124.6 and beyond the scale of the chart. The response of aboveground net primary production to
warming was stated to be nonsignificant (Wu et al.. 2011). but no effect size was given.
Source: Greaver et al. (2016) The number above each bar indicates the published source of the effect size as follows: 1. LeBauer
and Treseder (2008): 2. Wu et al. (2011): 3. Lu et al. (2013): 4. Liu and Greaver (2009): 5. Liu and Greaver (2010): 6. Xia and Wan
(2008): 7. Knorr et al. (2005): 8. Dieleman et al. (2012): 9. Treseder (2004): 10. Luetal. (2011b): 11. Liu et al. (2016b).
Figure 13-2 The effects of increased nitrogen, temperature, and precipitation
on terrestrial carbon pools (left panel) and fluxes (right panel)
from published meta-analyses.
"Belowground, initial findings are that N addition tends to increase the C stored in the
soil organic layer and in root biomass (Liu and Greaver. 2010; Xia and Wan. 2008).
while it tends to decrease mycorrhizae/microbial abundance and heterotrophic respiration
(Liu and Greaver. 2010; Treseder. 2004). This offset may result in no net change in soil
respiration (Liu and Greaver. 2010); however, this is an active area of research.
Consistent with the biokinetic effects of warming, long-term data and meta-analyses
show that soil respiration, including decomposition and microbial respiration, is
stimulated by increasing temperature (Lu et al.. 2013; Bond-Lambertv and Thomson.
2010; Rustad et al.. 2001). Most empirical studies show rising temperature stimulates N
release by mineralization (Churkina et al.. 2010). which may be driven more by
temperature effects on moisture (Emmett et al.. 2004). In some dynamic land models, the
additional N from mineralization will stimulate C uptake by plants even more than
current N deposition (Burd et al.. 2016). At the same time, increased N from
mineralization may cause N induced inhibition of decomposition, a feedback mechanism
that might decrease the amount of N released that is currently considered by few models
(Gerber et al.. 2010). The mechanisms causing N driven reduction in decomposition are
not well understood, but are thought to result from changes in microbial community
composition and their production of decomposition enzymes, as well as possible changes
in the character and degradability of soil organic matter (Conant et al.. 2011; Janssens et
al.. 2010). Climate change could also affect decomposition rates by altering both
available soil moisture and microscale connectivity among microorganisms, water, and
nutrients within the soil matrix that in turn may alter microbial processes (Xiang et al..
2008). Although there is no consensus about how dissolved organic carbon (DOC) in
surface water is regulated overall, increasing N and temperature increase DOC
concentrations (Laudon et al.. 2012). While few meta-analyses examine precipitation
effects on the soil C cycle, precipitation tends to increase the root C pool (e.g., [see
Figure 13-2 in this document])
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"Overall ecosystem C balance is assessed by summing measurements of individual pools
to quantify ecosystem carbon (EC) content or by measuring C fluxes to quantify the net
ecosystem exchange (NEE) of C. The meta-analysis we identified indicated that
temperature did not increase NEE lYLu et al.. 2013) positive NEE indicates ecosystem C
gain], and as previously mentioned this is likely because the biokinetic effects of
warming stimulate respiration that offsets the C capture via stimulation of primary
production. Increased precipitation tends to increase NEE (Wuetal.. 2011). likely
because water availability increases photosynthesis, while not increasing plant respiratory
losses. A meta-analysis of N addition studies indicates that adding N to grasslands had no
effect on NEE, but that N addition increased forests EC (Liu and Greaver. 2009). There
may be differences between grasslands and forests in terms of the extent that C gain
simulated by N is offset by heterotrophic and autotrophic respiration. Increasing
temperature may decrease C storage if the warming causes inhibition of photosystems, or
enhances evaporation and reduces water availability. A better understanding of these and
other contributing processes is needed.
"Traditionally, primary production in freshwater systems was thought to be phosphorous
(P) limited, but recent data have shown an increase of limitation by N or colimitation by
N and P (Elser et al.. 2007). In N limited fresh waters, N addition enhances rapid growth
of nitrophilic algae. This is an important food source to consumer species in the trophic
cascade; however, it is unclear if this is an important source of long-term C storage.
Sediments are estimated to be the largest pool of long-term C storage (Cole et al.. 2007).
and N stimulates the production of terrestrial biomass that may be transported to aquatic
sediments. Nitrogen also stimulates primary production of aquatic algae which contribute
to C in sediments (Kastowski et al.. 2011). although few studies have examined this
effect. Gudasz et al. (2010) found a strong positive relationship between increasing
temperature and organic C mineralization. They conclude future organic C burial in
boreal lakes could decrease 4-27% under scenarios of warming due to enhanced
temperature-dependent microbial activities. We were unable to identify studies
examining precipitation effects, or the combined effects of N, temperature, and
precipitation effects, on C storage in freshwater ecosystems. A discussion of biodiversity
associated with eutrophication of fresh waters is included in the biodiversity section."
13.1.2.2. Additional Considerations
There are considerations to note in addition to the excerpt from Greaver et al. (2016).
Additional papers on the interacting effects of climate and N on C cycling discussed in
Appendix 4 and Appendix 6 are summarized in Table 13-2. Notably there is a new
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meta-analysis on N and warming interactions and several new papers on the interacting
effects of N on plant biomass and precipitation.
Table 13-2 Summary on climate modification of nitrogen (N) effects on carbon
(C) cycling in Appendix 4 and Appendix 6 in addition to those in
Appendix 13.
Indicator/Process T P Snow	Reference
Biogeochemical effects (Appendix 4)
19 different C pools X X	Synthesis: A synthesis of meta-analyses for single Greaver et al. (2016)
and processes	factor experiments on N, T, or P manipulation for
within ecosystems	19 different C pools and processes within
ecosystems.
Soil C	X X	Meta-analysis: Results showed that the interaction of Ni et al. (2017)
warming and N deposition greatly increased the soil
C input (+49%) compared with the single factor of
either warming (+5%) or N deposition (+20%). Soil C
loss was not significantly affected by the interaction of
N and warming, likely because increases in
decomposition due to warming are offset by the
decreases by N addition.
N effects CO2, CH4,
and N2O
Meta-analysis: 313 observations across 109 studies
evaluated the effect of N addition on the flux of three
major GHGs: CO2, CH4, and N2O.
Liu and Greaver
(2009)
Decomposition X
Empirical: Increasing rainfall variability and N
addition can stimulate litter decomposition in tall
grass prairies in the U.S.
Schuster (2016)
Bioloqical effects (Appendix 6)
Biomass X
Empirical: Using 1,600 observations, the study
authors found that biomass responses to N increased
linearly with mean annual precipitation (MAP).
Xia and Wan (2008)
Biomass X
Empirical: Using 126 observations, the study authors
found no relationship between biomass and mean
annual precipitation (MAP).
LeBauer and
Treseder (2008)
ANPP	X X	Empirical: Plant growth response to N increased with Tian et al. (2016a)
precipitation until annual precipitation reached
800 mm/yr. The N response efficiency also peaked
with moderate annual temperatures (~8°C) and
declined under cooler or warmer conditions.
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Table 13-2 (Continued): Summary on climate modification of nitrogen (N) effects
on carbon (C) cycling in Appendix 4 and Appendix 6 in
addition to those in Appendix 13.
Indicator/Process T P Snow	Reference
ANPP	X	Empirical: In the Sonoran Desert near Phoenix, AZ, Hall et al. (2011)
60 kg N/ha/yr caused little or no increase in
production among herbaceous annuals in low
precipitation years, moderate N responses with
average rainfall, and strong increases in biomass with
added N during above-normal rainfall seasons.
Aboveground	X	Empirical: Added N was positively correlated with Vourlitis (2012)
biomass and litter	precipitation and was only significant in the high
production	rainfall years.
13.1.3. Climate and Acidification
13.1.3.1. Excerpt from Greaver et al. (2016)
"Atmospheric deposition of acidic N and sulfur (S) compounds (e.g., HNO3, H2SO4) has
directly caused widespread acidification of terrestrial and freshwater ecosystems in many
industrial areas. More broadly, other forms of atmospheric N deposition (e.g., NHx) can
indirectly cause acidification by increasing inorganic N availability enough to induce
nitrification. Freshwater and terrestrial ecosystem acidification is well studied and is
characterized by decreased pH and elevated aluminum (Al) concentrations/mobility in
soils and surface waters that cause plant physiological changes (Schaberg et al.. 2002).
tree mortality (McNultv et al.. 2007). aquatic fauna mortality, and decreased aquatic
biodiversity (Driscoll et al.. 2001b). Acidification driven by N (as opposed to N + S)
occurs at higher levels of N addition than for initial changes to the C cycle. Often N
saturation of the terrestrial ecosystem and subsequent leaching into adjacent aquatic
systems is observed in the process of aquatic acidification (Aberetal.. 1998). The
threshold for the onset of acidification changes across the landscape, depending on
geochemical sensitivity and historical loading of acidifying deposition. Although recent
declines in emissions of SOx and NOx in eastern North America and throughout Europe
(NAPAP. 2011) have led to many improvements in acid-base balances in acid sensitive
ecosystems in recent decades, it is unclear whether sensitive ecosystems will continue to
improve as emissions decline or whether secondary processes will promote, arrest, or
reverse ecosystem recovery as climate changes.
"Potential future changes in the quantity and temporal distribution of precipitation and
temperature (and their interactions) is expected to alter the wet-dry cycles that govern the
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timing and amount of acidic inputs in precipitation, microbial transformation in the soil,
and the flush of acid anions from soils to surface waters. If acid anions build up in soil
during periods of drought, the eventual flushing likely causes a more potent acidification
event (Moslcv. 2015; Whitehead et al.. 2009). If the acidification event occurs during a
time when sensitive biota or lifestages of biota are present, acidification may cause more
adversity to these populations (Kowalik et al.. 2007). Increases in storm frequency
associated with global climate change (Collins et al.. 2014) could increase the frequency
and severity of acidification driven by high levels of sea salt deposition in coastal regions
(Wright and Schindler. 1995). Although the mechanisms of interaction are unclear,
increases in DOC concentrations in aquatic ecosystems across Europe and the U.S. have
been linked to acidification, N cycling, and climate change, with important implications
for water quality and ecosystem function (Evans et al.. 2008).
"As previously mentioned, warmer temperatures increase decomposition and
nitrification. Nitrification will also increase with increased N supply caused by increased
weathering or decomposition (Booth et al.. 2005). The process of nitrification generates
protons that increase the rate of nitrate and base cation leaching to drainage waters
(Murdoch et al.. 1998). The combined increase of NO3 leaching and loss of base cations
has the potential to magnify acidification in forest soils (Fernandez et al.. 2003). Soil
weathering is typically the key buffer to acidic deposition (Li and McNultv. 2007). and
while weathering is increased by both soil temperature and soil moisture (Gwiazda and
Broecker. 1994). it is unclear whether any future change in the magnitude of temperature
and precipitation will be enough to alter base cation supply or influence the acid-base
balance of sensitive ecosystems. Furthermore, it is unclear whether increased supply of N
in soils from either deposition, increased decomposition, or increased nitrogen fixation
may negate the ameliorative effect of enhanced weathering. Some studies show that
climate change will mitigate acidification through increased weathering (Belvazid et al..
2011a). while others show that climate change will aggravate acidification although
increased nitrification outpacing enhanced weathering (Wu and Driscoll. 2010). In
general, increased temperature and precipitation will likely enhance inputs of buffering
agents from weathering and deposition, but also increase inputs of acidifying agents from
deposition and enhanced N cycling. The relative sensitivity of these opposing processes
to a given change in climate remains unresolved.
"Climate change may alter the sensitivity of biota to acidification, creating the need to
adapt to a combination of acidity and climate change stresses. For example, the suitable
habitat for brook trout in the Catskills and Adirondack mountains of the northeastern U.S.
may be constrained as climate change increases downstream water temperatures,
reducing downstream range where the trout can survive, while upstream migration is
limited in part by acid conditions in the headwaters. In another example, Al is toxic to
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many fish and known to be mobilized during acidification events. It is known that the
mortality rate of Atlantic salmon exposed to A1 increases at higher temperatures (ABS
and Muniz. 1993); it is unclear how many other aquatic species would experience
temperature-dependent toxicity, which could make them more vulnerable to acidification
in a warming climate. Overall, there is little knowledge of how the biological thresholds
to acidity will be affected by climate change."
13.1.3.2. Additional Considerations
There are considerations to note in addition to the excerpt from Greaver et al. (2016).
Additional papers on climate interactions with acidification discussed in Appendix 4 and
Appendix 5 are summarized in Table 13-3. Most notably, several new studies model the
interacting effects of climate and deposition on soil biogeochemistry, plant growth, and
plant community composition.
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Table 13-3 Summary on climate modification of acidification in Appendix 4 and
Appendix 5 in addition to those in Appendix 13.
Indicator/Process T P Snow	Reference
Biogeochemical effects (Appendix 4)
Soil acidification X X	Model: 2009-2100—six climate change simulations Pourmokhtarian et
and N enrichment	of temperature, precipitation, and photosynthetically al. (2012)
active radiation (PAR). Without CO2 fertilization
effects, net soil mineralization and nitrification
increased; soil and stream acidified; and percentage
of base saturation in soils declined due to increased
NO3". With CO2 fertilization effects, N loss to streams
is suppressed due to increased plant uptake.
Model: The ForSAFE-VEG model simulations show Belvazid et al.
that climate and atmospheric deposition have	(2011a)
comparably important effects on N mobilization in the
soil, because climate triggers the release of
organically bound nitrogen stored in the soil during
the elevated deposition period. Climate has the most
important effect on plant community composition,
underlining the fact that this cannot be ignored in
future simulations of vegetation dynamics.
Soil acidification X X	Model: PnET-BGC modeling found climate change Wu and Driscoll
and N enrichment	played a larger role in ANC than base cation	(2010)
deposition changes. Temperature had a larger effect
than precipitation on decomposition. Net
mineralization and nitrification increased faster with
climate change than plant NO3" uptake.
Biological effects (Appendix 5)
Growth of red	X X	Empirical: Tree ring analysis was conducted along Wason et al. (2017)
spruce and balsam	an elevation gradient (proxy for climate change). Both
fir	species showed increased growth with increased
precipitation pH. Red spruce appeared to show
increased growth with a warming climate but balsam
fir did not. No changes were noted in the [species?]
distribution in the spruce-fir forest, and the authors
suggest any such change may be part of a longer
term process.
Model: The Annual Radial Model used to investigate Koo et al. (2014)
projected climate change effects and changing air
pollution on red spruce growth. High-elevation
(<1,700 m) red spruce growth would decline if climate
change co-occurred with an increase in air pollution.
However, growth would increase under climate
change, if air pollution decreased. In contrast,
low-elevation red spruce growth would decrease with
future climate change regardless of change in air
pollution.
Soil acidification, N X X
enrichment, and
plant community
Growth of red	X X
spruce
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Table 13-3 (Continued): Summary on climate modification of acidification in
Appendix 4 and Appendix 5 in addition to those in
Appendix 13.
Indicator/Process T P Snow	Reference
Biological effects of freshwater acidification (Appendix 8)
Water DOC	X	The major stressor affecting native coldwater fish Warren et al. (2017)
species in the eastern U.S. is shifting from
acidification to thermal stress and some lakes
recovering from acidification may provide a degree of
protection against climate effects. As DOC in the
water increases with increasing lake pH in recovering
lakes, decreased water clarity may create cooler
refuge habitat for fish.
13.1.4. Nitrogen, Climate, and Biodiversity
13.1.4.1. Excerpt from Greaver et al. (2016)
"Biodiversity, which contributes to the structure and function of ecosystems, is declining
globally (Rockstrom et al.. 2009). Decades of study show that added N reduces
autotrophic diversity in terrestrial and aquatic systems, fungal biodiversity in soils, and,
although less studied, animal diversity in terrestrial and aquatic systems (Howarth et al..
2011; Bobbink et al.. 2010). As previously mentioned, two mechanisms that contribute to
altered biodiversity are eutrophication and acidification. Eutrophication often causes N
stimulated growth for opportunistic species, which may cause competitive exclusion of
poorer competitors and soil acidification driving cation imbalances and physiological
stresses, suppressing seed germination and seedling regeneration (Sullivan et al.. 2013;
Roem et al.. 2002). In addition, N may alter physiology and/or community properties,
increasing the risk to secondary factors such as pests, fire, frost, and drought (Bobbink et
al.. 2010). Lastly, direct damage to vegetation from ammonia (NH3), nitrogen dioxide
(NO2), nitrogen monoxide (NO), peroxyacetyl nitrate (PAN), and nitric acid (HNO3)
exposure is known to occur, but most likely occurs in highly polluted areas and in close
proximity to high emission sources (Greaver et al.. 2012). In terrestrial ecosystems, all
processes generally reduce local autotroph diversity and homogenize habitats into
communities with small numbers of generally fast-growing or acid-tolerant autotrophic
species. These changes propagate through the food web, leading to increases of generalist
pests, herbivores, and parasitic soil bacteria, and decreases in specialist herbivores and
beneficial microbial communities that occur belowground (de Sassi et al.. 2012).
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"Biodiversity responses on land may be moderated by the type of climate change
occurring and the mechanism of N response. In terrestrial systems that get warmer and
wetter, eutrophication may be amplified if endogenous N sources are low or dampened if
endogenous sources are high/more liberated to meet community demand. The response of
the acidification pathway may depend on whether the change in net fluxes of cations
from climate change exceeds the net fluxes of N [i.e., from enhanced deposition of
cations and N, decomposition or weathering, and leaching (Howarth et al.. 2011; Roemet
al.. 2002)1. Many of these processes may be dampened in terrestrial systems that are
anticipated to get warmer and drier (e.g., southwestern U.S.) due to the drier conditions
reducing biological activity (Bobbink et al.. 2010). Colder regions such as montane,
alpine, and tundra systems are often strongly N limited and poorly buffered; thus, an
extended growing season under climate change will lead to greater opportunities for all
operating processes. Climate change may also magnify the effects of secondary stressors
in several ways, including increased pest populations under warmer wetter conditions, as
well as increasing fire potential and drought vulnerability as more aboveground plant
tissue is produced under elevated N (Pise et al.. 2011). Nonetheless, field evidence for
interactions between N and climate change under controlled conditions is scarce. The few
existing studies find additive effects in Mediterranean California (Zavalcta et al.. 2003).
and no interactive effects of precipitation and N addition in Minnesota (Reich et al..
2014; Reich. 2009). Not all terrestrial ecosystems are anticipated to be equally sensitive
to these pressures. In grasslands, many forbs and slow-growing species such as native C4
grasses appear especially vulnerable to added N (Clark and Tilman. 2008; Zavaleta et al..
2003). In a study of northeastern forests, all three tree species with negative growth
responses to N deposition were evergreen conifers (e.g., Pinus resinosa, Picea rubens,
Thuja occidentalis), while all five tree species with positive growth responses were
broadleaf species with arbuscular mycorrhizal associations [e.g., Acer rubrum, A.
saccharum, Fraxinus americana, Liriodendron tulipifera, and Prunus serotina (Thomas
et al.. 2010)1. However, contingent factors underly these general patterns, as there were
tree species from each group that did follow these generalities.
"In aquatic systems, elevated temperatures and N inputs from increased rain and glacial
retreat will likely magnify changes in algal assemblages that can propagate through the
food web (Hobbs et al.. 2010; Elseretal.. 2009a). In freshwater aquatic biodiversity
research, there is a substantial amount of work on lakes investigating the effects of
warming via gradient studies (latitude or altitude), warming experiments, time series, and
palaeoecology (Jeppesen et al.. 2014). Fish community assemblages, size, structure, and
dynamics are likely to change with continued global warming, and in some cases the
elevated temperatures that have already occurred in the past decades. Fish cannot
thermoregulate, but only physically move to areas with appropriate temperatures, if those
are accessible. In general, changes in fish composition, particularly in shallow lakes, are
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characterized by a decline in abundance of several cold-stenothermal species (Kangur et
al.. 2013; Winfield et al.. 2010) and in increase in eurythermal species, which exhibit a
wide range of thermal tolerance (Jcppcscn et al.. 2012). Many fish species are also
adapted to specific oxygen concentrations; when temperature increases, oxygen may drop
to critical levels as warm water holds less oxygen and the respiration rates increase.
Warming effects on the biodiversity of grazing macroinvertebrates and zooplankton is
mixed across studies (Qzcn et al.. 2013; Burgmer et al.. 2007). Warming is shown to
increase cyanobacteria biomass (Kostcn et al.. 2012) and biofilm biomass (Williamson et
al.. 2016). while warming effects on phytoplankton show mixed results (Mccrhoff et al..
2012; Feuchtmavr et al.. 2009).
"Warming may cause increases in evaporation that will lower water level and increase
salinity ("Williams. 2001). Increasing salinity of freshwater systems tends to have a
negative effect on phytoplankton, zooplankton, macroinvertebrates, and fish (Jeppesen et
al.. 2007). Climate change will alter the transport, availability, and timing of N in
ecosystems, and recent data indicates an increase of primary production limitation by N
or colimitation by N and P (Elser et al.. 2009a). In N-limited fresh waters, N enrichment
enhances rapid growth of nitrophilic algae that outcompete other populations for light and
resources such as P and silicon (Si), leading to dominance by a few algal species and a
reduction in the nutritional quality of invertebrates as food for fish (Baron et al.. 2012;
Hobbs et al.. 2010; Elser et al.. 2009a). The combination of increasing both N and
temperature may be synergistic and sometimes difficult to uncouple (Kangur et al..
2013). as both may stimulate hypoxic conditions, consequently altering community
structure (Ozen et al.. 2013; Feuchtmavr et al.. 2009) and the frequency, intensity, extent,
and duration of harmful algal blooms (Paerl etal.. 2011).
"Presently, there are at least 78 listed or candidate species for threatened or endangered
status in North America that have N impacts identified as a primary contributor, and an
estimated 15-37% of species may be at risk from climate change (Hernandez et al..
2016). In total, there are numerous pathways whereby these dominant global change
factors can interact to impact biodiversity, and it is likely, although not definitive, that N
and climate often have additive and potentially amplifying effects on decreasing
biodiversity in many systems."
13.1.4.2. Additional Considerations
There are considerations to note in addition to the excerpt from Greaver et al. (2016).
Additional papers on climate interactions with N and biodiversity discussed in
Appendix 6 are summarized in Table 13-4. Most notably there are several new studies on
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the interacting effects of climate and N deposition on plant species richness, and
community composition of plants, soil microbes and soil animals.
Table 13-4 Summary on climate modification of nitrogen (N) effects on
biodiversity in Appendix 6 in addition to those in Appendix 13.
Indicator/Process
T
p
Snow
Reference
Plant species
richness
X

Empirical: No significant interaction between N and
temperature on plant species richness for
closed-canopy systems (deciduous, evergreen, and
mixed forests), but significant interaction in
open-canopy systems (grasslands, shrublands, and
woodlands). In these open-canopy ecosystems, N
had a more negative effect on species richness at
lower temperatures.
Simkin et al. (2016)
Plant community
composition
X
X
Model: In the ForSAFE-VEG model projections of
plant community composition in three French forests
two N reduction scenarios showed that recovery of
these plant communities occurred only if climatic
factors are held constant at current levels.
Rizzetto et al. (2016)
Plant community
composition
X
X
Model: Understory plant community composition in
northern hardwood forests at Bear Brook Watershed
in Maine and Hubbard Brook in New Hampshire: the
simulated plant community composition returned
toward preindustrial conditions over the next century
only in a scenario in which N deposition rates
returned to background and climate was kept stable.
Phelan et al. (2016)
Soil microbial
community

X
Empirical: N additions and lower precipitation
interacted to shift microbial community composition in
forests soils in northeastern China.
Wanq et al. (2014c)
Soil microbial
community

X
Empirical: N additions and lower precipitation
interacted to shift microbial community composition in
grassland soils in Inner Mongolia.
Li et al. (2016a)
Ectomycorrhizal
species
composition

X
Empirical: Both changes in precipitation and nitrogen
were associated with shifts in ectomycorrhizal
species composition in European Scots pine.
Jarvis et al. (2013)
Nematode
predators,
microarthropod
herbivores, and
taxa richness of
nematodes and
microarthropods

X
Empirical: In Minnesota grassland, a decline was
observed in several categories of biota within a soil
food web (numbers of nematode predators,
microarthropod herbivores, and taxa richness of
nematodes and microarthropods) under increased N,
and a decrease in ciliate protists under high N and
drought.
Eisenhauer et al.
(2012)
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Table 13-4 (Continued): Summary of climate modification of nitrogen (N) effects
on biodiversity in Appendix 6 in addition to those in
Appendix 13.
Indicator/Process T P Snow	Reference
Algal species	X	In regions with no evidence of increased atmospheric Ruhland et al
composition	nutrient inputs, warming trends are observed to	(2015).
enhance competitiveness of planktonic diatoms like
Asterionella formosa, which are typically associated
with elevated N, indicating that climate change has
significant direct and indirect effects on algal species
composition. Climate change may thus enhance the
effects observed in areas with nutrient increases
alone.
13.2. Estuaries
In addition to terrestrial and freshwater systems described above, climate change
modifies key processes in estuarine and near-coastal systems linked to nutrient inputs.
Atmospheric deposition may represent more than 40% of N in coastal water bodies
(Appendix 7). N sources to near-coastal ecosystems include direct deposition to the water
surface and all other N sources upstream. Coastal eutrophication is a process of
increasing nutrient over-enrichment indicated by water quality deterioration, resulting in
numerous adverse effects including hypoxic zones (water with dissolved oxygen that is
too low to support marine life), species mortality, proliferation of phytoplankton and
macroalgae (including harmful algal blooms), and decreased coverage of submerged
aquatic vegetation lYBricker et al.. 2007); Appendix 101.
Indicators of nutrient enrichment may be affected by climate-related changes in estuaries
including temperature, precipitation, wind patterns, stronger estuary stratification,
increased metabolism and organic production, and sea-level rise [see discussion in
Appendix 10; (Altieri and Gedan. 2015; Statham. 2012; Rabalais et al.. 2009)1. For
example, eutrophic conditions and the extent and duration of hypoxia are predicted to
increase with changes in temperature and precipitation (Altieri and Gedan. 2015;
Rabalais et al.. 2009; Boesch et al.. 2007). Donev et al. (2012) reviewed effects of
climate change, including coastal hypoxia, on marine biodiversity and reported
population-level shifts. In estuaries affected by nutrient enrichment and
climate-associated alterations described above, conditions favor a shift in phytoplankton
composition toward greater abundance and distribution of toxic cyanobacteria associated
with an increased prevalence of harmful algal blooms (Paerl et al.. 2016a). In a
meta-analysis of temperature effects on benthic macrofauna, survival times significantly
decreased in hypoxic conditions under warmer temperatures (Vaquer-Sunver and Duarte.
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2011). Decreases in estuarine biodiversity associated with N loading can be magnified by
hydrologic factors. Glibert et al. (2014) observed a regional change in phytoplankton
community composition and decreased coverage of submerged aquatic vegetation
following a shift from long-term dry to long-term wet conditions in the early 2000s in
shallow coastal lagoons along the coast of Maryland and Virginia. Shifts in
phytoplankton community structure are known to occur in estuaries due to N enrichment,
but climate changes may at times outweigh the impacts of eutrophication (Pacrl et al..
2010).	Thus, physical changes caused by climate change should be considered when
modeling phytoplankton community response to N enrichment (Pacrl et al.. 2014).
13.3. Wetlands
Changes in mean annual temperature and frequency and magnitude of precipitation will
affect the responses of all wetlands to N loading (Appendix 11).
Increasing temperatures may strengthen N effects upon wetland ecosystems. Temperature
effects on wetlands have been demonstrated in European bog ecosystems, where
increased temperatures increased the cover of woody species and decreased Sphagnum. N
addition and temperature are known to synergistically depress Sphagnum production,
with a 1°C increase in summer temperature having an impact on the Sphagnum
equivalent to an additional 40 kg N/ha/yr (Limpens et al.. 2011).
Hydrologic regimes are important controls on wetland cycling and productivity, so
changes in the magnitude and frequency of precipitation can have strong effects on
ecosystem N retention and C storage. Increased precipitation is shown to increase
Sphagnum sensitivity to N addition-induced decreases in production (Limpens et al..
2011).	Experimental mesocosms modeling changes in precipitation to salt marshes found
that precipitation delivered in infrequent, heavy storm events decreased N retention and
plant productivity, even though storm events delivered a higher N deposition load to the
marsh (Hanson et al.. 2016; Oczkowski et al.. 2016). Shifts in precipitation towards less
frequent, more intense rain events may strengthen N deposition-induced decreases in salt
marsh N retention while weakening N deposition-induced increases in plant productivity.
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APPENDIX 14. ECOSYSTEM SERVICES
This appendix is a review of the state of the science on ecosystem services altered by N
and S deposition. The state of the science in 2008, as presented in the 2008 Integrated
Science Assessment for Oxides of Nitrogen and Sulfur-Ecological Criteria (hereafter
referred to as the 2008 ISA), is summarized and sets the foundation for the discussion on
the new literature (published from 2008-present).
14.1. Introduction
The term "ecosystem services" refers to a concept that ecosystems provide benefits to
people, directly or indirectly (Costanza et al.. 2017). and these benefits are socially and
economically valuable goods and services deserving of protection, restoration, and
enhancement (Bovd and Banzhaf. 2007). The concept of ecosystem services recognizes
that human well-being and survival are not independent of the rest of nature, humans are
an integral and interdependent part of the biosphere (Costanza et al.. 2017). In some
cases, and in line with more conventional economic thinking, ecosystem services analysis
can result in attaching monetary values to ecosystem outcomes. However, monetary
estimation may not always be possible and this does not diminish the importance of the
ecosystem services to human well-being. At a minimum, ecosystem services analysis
involves discussion and, ideally, quantification of ecological outcomes understood by
households, communities, and businesses. Explicitly linking ecosystem services with
affected people has proven difficult because of the broad definition of ecosystem services
and the numerous types of services that could be affected. An analysis of ecosystem
services specifically altered by NOx, SOx, and PM would translate the effects of ambient
concentrations and deposition into biological, physical, or monetary metrics that give
insight to public welfare effects.
The 2008 ISA documented the ecosystem services frameworks that were published at that
time and provided several examples of how N and S deposition would alter services
based on those frameworks. Since the 2008 ISA, new ecosystem services frameworks
have been developed and several new approaches have been applied to conduct
ecosystem services analysis specifically evaluating outcomes altered by N and S
deposition. The conclusions considering the full body of literature published are that
(1) there is evidence that N and S emissions/deposition have a range of effects on
ecosystem services and their social value; (2) there are some economic studies that
demonstrate such effects in broad terms; however, it remains methodologically difficult
to derive economic costs and benefits associated with specific regulatory
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decisions/standards; and (3) thousands of causal relationships are now documented
between N and S air pollution and changes in Final Ecosystem Goods and Services
(FEGS), defined as the "components of nature, directly enjoyed, consumed or used to
yield human wellbeing" (Bovd and Banzhaf. 2007). The following sections discuss the
relevant ecosystem services frameworks (Appendix 14.2). the studies in the U.S. on
ecosystem services altered by Acidification and Eutrophication (Appendix 14.3). the
studies of ecosystem services analyses conducted in countries outside of the U.S.
(Appendix 14.4 and Appendix 14.5). and a summary (Appendix 14.6).
14.2. Ecosystem Services Frameworks
Ecosystem services frameworks can help identify causal pathways between specific
ecosystem changes and potential human beneficiaries. A single classification system may
be desirable to promote standardization and communication. However, classification
systems do not by themselves solve the core issue of how to empirically develop causal
relationships between N and S and final ecosystem goods and services. Here an overview
of key frameworks is presented to give context on conceptual advances in ecosystem
services research.
Several ecosystem services frameworks were published in 2008 and described in the
2008 ISA. These frameworks provided qualitative and/or quantitative metrics by which to
identify the services that ecosystems provide to benefit human welfare and society fWRI.
2000; Costanza et al.. 1997; Daily. 1997; Pimentel etal.. 1997). It was well documented
at that time that some goods and services have explicit market value; however, the value
of other services is more difficult to assess (U.S. EPA. 2008a; Goulder and Kennedy.
1997). The most accepted framework for ecosystem services identified by the 2008 ISA
was the 2005 Millennium Ecosystem Assessment (MEA) by Hassan et al. (2005). Since
the 2008 ISA, several new ecosystem services frameworks and classification systems
have been published. These include The Economics of Ecosystems and Biodiversity
[TEEB; (Sukhdev et al.. 2014)1. the Common International Classification of Ecosystem
Services [CICES; (Haines-Young and Potschin. 2011)1. U.S. EPA Final Ecosystem
Goods and Services Classification System [FEGS-CS; (Landers and Nahlik. 2013)1. and
The National Ecosystem Services Classification System [NESCS; (U.S. EPA. 2015d)l.
The goal of the TEEB initiative is to demonstrate the economic value of services
provided by species and ecosystems and then capture the value of these services through
market- or policy-based approaches (Sukhdev et al.. 2014). The TEEB initiative uses four
primary categories of ecosystems services. Three of the four categories are from the
MEA (provisioning, regulating, and cultural services). However, TEEB replaces the
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MEA's "supporting services" with "habitat services," which emphasizes the provisioning
of species and genetic diversity.
The CICES framework was developed in part as a reaction to the increasingly diverse
ways that ecosystem services were discussed, categorized, and assessed by different
organizations around the world, the CICES program has as a primary goal to create a
more uniform approach to assessing ecosystem services (Haines-Young and Potschin.
2011). The classification approach CICES developed recognizes only final outputs from
ecosystems, in other words, "products" that are directly consumed or used by people
(Haines-Young and Potschin. 2011). Thus, the concept of "supporting services"
identified by the MEA was dropped under the assumption that these functions would be
recognized for their role in facilitating the final outputs. CICES developed a hierarchical
structure, beginning with the three broad categories (provisioning, regulating, and
cultural), which were broken down into 23 "service groups" and then 59 "service types"
(Haines-Young and Potschin. 2011).
The FEGS-CS framework uses the concept of Final Ecosystem Goods and Services
(FEGS), which are defined as a subset of ecological outcomes, specifically the
"components of nature, directly enjoyed, consumed or used to yield human wellbeing"
(Bovd and Banzhaf. 2007). The U.S. EPA FEGS-CS approach defines and classifies
338 unique FEGS, each uniquely numbered by a combination of environmental class or
subclass and a beneficiary category or subcategory (Landers and Nahlik. 2013). In
addition to FEGS-CS, the U.S. EPA also developed NESCS to specifically analyze the
human welfare impacts of policy-induced changes to ecosystems (U.S. EPA. 2015d). The
goal of NESCS is to support analysis of changes from baseline conditions, such as
cost-benefit analysis, and support systematic accounting of changes in ecosystem
services. The NESCS structure defines categories and numeric codes that are designed to
help identify flows of services from ecosystems to human beings, while distinguishing
between the producers (i.e., "supply-side") and users (i.e., "demand-side") of the service.
Lastly, Bell et al. (2017) developed the Stressor—Ecological Production function—final
ecosystem Services (STEPS) framework. STEPS produces "chains" comprised of the
biological indicator, the ecological production function (EPF, which uses ecological
components to link the biological indicator to a final ecosystem service), and the user
group who directly uses, appreciates, or values the component. To this end, STEPS is
used in conjunction with other frameworks that identify user groups, like FEGS-CS. The
framework uses a qualitative score (high, medium, low) to describe the strength of
science (SOS) for the relationship between each component in the EPF.
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14.3. United States Applications
Among the ecological effects of N and S deposition, the ecosystem services altered by
acidification are the most well documented, with several new papers published since
2008 (Beier et al.. 2017; Caputo et al.. 2017; Irvine et al.. 2017; O'Dea et al.. 2017;
Banzhaf et al.. 2016; Office of Air and Radiation. 2011; Chestnut and Mills. 2005;
Banzhaf et al.. 2004; U.S. EPA. 1999; NAPAP. 1991). In the 2008 ISA, no publications
were identified describing how ecosystem services were altered as a result of N driven
eutrophication; however, several new comprehensive studies have been published since
2008 (Clark et al.. 2017; Rhodes et al.. 2017; Compton et al.. 2013; Birch et al.. 2011;
Compton et al.. 2011). Also new to the literature since 2008 is a group of publications
(Bell et al.. 2017; Clark et al.. 2017; Irvine et al.. 2017; O'Dea et al.. 2017; Rhodes et al..
2017) evaluating the FEGS altered by four modes of ecological response to N and S
deposition (aquatic acidification, terrestrial acidification, aquatic eutrophication, and
terrestrial eutrophication). In these analyses critical load exceedances for N related air
pollution were used as a model stressor from which a total of 1,104 unique chains linking
stressor to beneficiary were identified. These analyses, built using a wholistic STEPS
approach, represent a great advance for identifying the wide range of ecosystem
processes and types of final ecosystem services affected by N and S deposition. However,
it also underscores the many information gaps that exist for quantifying the dose-response
relationships between N and S air pollution exposures and final ecosystem services and
the values associated with changes in these services.
Two recently published studies (Jaramillo and Muller. 2016; Holland. 2015) have used
the Air Pollution Emissions Experiments and Policy (APEEP or AP2) to calculate the
value of damages from emissions of SO2, VOCs, NOx, PM2 5, PM10, and NH3. The model
connects changes in emissions to concentrations of several criteria air pollutants. A large
majority of the damage estimates are associated with human health impacts and
ozone-related damages to crops and forests. A very small portion of damages are linked
to forest and recreation impacts from a mix of ambient pollutants including NOx and
SOx, but this portion of the model was developed before 2006, has not been updated, and
is not well documented (i.e., the model structure is described but not the estimated
parameters).
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How do N deposition impacts oil lace lichen affect forest ecosystem services?
Native to the Pacific coast of North America, lace lichen (Ramalina memiesii [Ramalinaceae])
grows from Alaska to Baja California. It lias a unique morphology that adds to the natural
beauty of coastal woodlands along the Pacific Coast (Hastings Reserve. 2015). It is a food
source for browsing mammals, including deer, cows, and squirrels, and is also used as a nesting
material for hummingbirds and orioles (Hastings Reserve).
Photos source: 2016 AOE workshop report.
Unfortunately, lace lichen abundance is threatened by nitrogen and sulfur pollutants" deposition
(Hernandez et al.. 2016). It is very sensitive to air pollution, including exposure to nitric acid
(Riddell et al„ 2008). and it has been found to contribute to the global uptake of sulfur dioxide,
an airborne pollutant (Gries et al.. 1997). Lace lichen was formerly found in the San Jacinto
Mountains near Los Angeles and on the surrounding coastal plain but now it only grows at
elevations above the smog layer (Hastings Reserve. 2015).
Declines in lace lichen abundance can reduce ecosystem services to humans in several ways. In
addition to reducing the aesthetic benefits provided by lichens, declines in lace lichen may be
indirectly contributing to declining populations of the spotted owl, which is ail endangered,
widely recognized, and charismatic species native to Pacific Northwest forests. This indirect
impact may be occurring because the northern flying squirrel, which is an almost exclusive food
source for the spotted owl, relies exclusively on forage lichens as a winter time food source.
14.3.1. Acidification
Prior to 2008, ecosystem acidification in the U.S. was evaluated with regard to how
ecosystems services respond to lowering NOx and SOx emissions. Initial monetary
valuation of the total annual value for improvements in recreational fishing (NAPAP.
1991) were improved with updated economic studies and new models for estimating the
changes in fish stocks and catch rates in the Adirondacks U.S. EPA (1999). A few years
later, Chestnut and Mills (2005) compared the actual benefits of reducing emissions of
NOx and SOx in Title IV of the Clean Air Act Amendments (CAAA) by the estimate of
benefits made in 1990. They conclude that quantitative assessment was problematic at
that time due to a lack of measurement units to gauge changes in the quality and quantity
of ecosystem services and a lack of dose-response relationships to indicate how quality
and quantity may change as a function of changes in pollution exposures. A different
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approach, that did not rely on dose-response, documented that New York households are
willing to pay an average of $45 to $100 annually to reduce the number of acidified lakes
by 40% and improve forest health in the Adirondacks (Banzhaf et al.. 2004).
Several ecosystem services valuations for acidification published (since 2008) are given
in Table 14-1. Rea et al. (2012) paired biogeochemical modeling using the Model for
Acidification of Groundwater in Catchment (MAGIC) with two different approaches, one
benefit transfer and one random utility, to estimate ecosystem services values for
remediation scenarios of acidic deposition on lakes in New York's Adirondack Park.
Beier etal. (2017) and Caputo et al. (2017) have also calculated new monetary valuations
of acidification effects in the Adirondacks (Table 14-1). New work in the southern
Appalachian region used an approach that weds biogeochemical (e.g., MAGIC) and
economic modeling to evaluate the cost of aquatic and terrestrial acidification (Banzhaf et
al.. 2016).
Two new studies used FEGS-CS to quantify how many FEGS are altered by ecosystem
acidification and associated CL or TL exceedance (Irvine et al.. 2017; O'Dea et al.. 2017).
O'Dea et al. (2017) identified human beneficiary groups for each ecological endpoint
affected by aquatic acidification, including effects on otter, mink, loon, aquatic
vegetation, shellfish, brook trout, and bass. Irvine et al. (2017) documented the effects of
terrestrial acidification on two acid-sensitive tree species, balsam fir (Abies balsamea)
and white ash (Fraxinus americana) (Table 14-1).
Overall, the new literature since the 2008 ISA includes studies that better characterize
ecosystem services valuation by pairing biogeochemical modeling and benefit transfer
equations informed by WTP surveys, especially for the Adirondacks and Shenandoah
regions. Aside from valuation, the estimate of the total number of ecosystem services
affected by N and S deposition is more well quantified by the new studies that use
FEG-CS. However, for many regions and specific services, poorly characterized
dose-response between deposition, ecological effect, and services are the greatest
challenge in developing specific data on the economic benefits of emission reductions
(NAPAP. 2011).
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Table 14-1 Ecosystem services research related to ecosystem acidification.
Mode
Region
Effect
Reference
Aquatic
acidification
Adirondacks
In 1990, the estimates of total annual value for improvements
due to lowering NOx and SOx emissions in recreational
fishing were $12 to $24 million.
NAPAP (1991)
Aquatic
acidification
Adirondacks
Reported estimates of annual value of improvements in
conditions for recreational fishing in the Adirondacks of about
$65 million for the reductions in acid deposition as a result of
the 1990 CAAA, primarily Title IV.
U.S. EPA (1999)
Aquatic
acidification

Compare the actual benefits of reducing emissions of NOx
and SOx in Title IV of the Clean Air Act Amendments (CAAA)
by the estimate of benefits made in 1990.
Chestnut and Mills
(2005)
Aquatic
acidification
Adirondacks
New York households are willing to pay an average of $45 to
$100 annually to reduce the number of acidified lakes by 40%
and improve forest health in the Adirondacks.
Banzhaf et al.
(2004)
Aquatic	Adirondacks Linked NOx and SOx emissions estimates to CMAQ	Office of Air and
acidification	modeling, MAGIC modeling, then a Random Effects Model Radiation (2011)
that was developed to account for fishing site choices made
by recreational fishers based on attributes of sites specifically
in the Adirondack region. The difference in economic welfare
values between the value of fishable (i.e., not impaired) lakes
in the with-CAAA scenario and the without-CAAA scenario
represented the benefits to recreational fishing in the
Adirondack region associated with the CAAA.
The changes in percent base saturation levels in timber	Office of Air and
harvest areas were mapped in relation to potential changes in Radiation (2011)
the growth and health of tree species present in these areas
and the likely effects of altered tree growth and health on
timber harvest rates and volumes. Of the forest types of
interest, the paper birch forest type experiences the greatest
increase in percent base saturation due to the CAAA,
followed by the eastern hemlock and the sugar
maple/beech/yellow birch forest types.
Biogeochemical modeling (e.g., MAGIC) is combined with a Rea et al. (2012)
benefit transfer approach based on a stated preference study,
which captures ecosystem services values as a whole. Total
values range from $547 million to $1 billion per year. In
contrast, an approach based on a preference study, which
only captures recreation fishing values, estimated benefits
ranging from $7 million to $9 million per year.
Terrestrial Adirondacks
Acidification
Aquatic	Adirondacks
acidification (44 lakes)
Aquatic and Southern
terrestrial Appalachia
acidification (169 sites)
Hypothetical scenarios of lower emissions are coupled with
biogeochemical modeling (e.g., MAGIC) and a novel stated
preference survey to determine individuals' WTP is used to
generate aggregate benefits of about $3.7 billion, or about
$16 per year per household in the region.
Banzhaf et al.
(2016)
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Table 14-1 (Continued): Ecosystem services research related to ecosystem
acidification.
Mode	Region	Effect	Reference
Terrestrial Adirondacks Blending regression analysis of tree growth and soil condition Beier et al. (2017)
acidification	with a benefit transfer approach, it was estimated that
acid-impaired hardwood forests provide roughly half of the
potential benefits of forests on moderate to well-buffered
soils—an estimated loss of-$10,000 ha-1 in net present
value of wood products, maple syrup, carbon sequestration,
and visual quality.
Aquatic	Adirondacks Acidic deposition has had little effect on lakes water suitable Beier et al. (2017)
acidification (52 study for drinking in the region. However, as pH declines in lakes,
lakes)	the estimated value of recreational fishing decreases
significantly due to loss of desirable fish. The expected value
of recreational fishing increased with increasing pH, from a
minimum of $4.41 angler/day in unstocked, acid-impaired
lakes to a maximum of $38.40 angler/day in well-buffered
lakes that were stocked with trout.
Aquatic	Adirondacks Analysis of benefit transfer using willingness-to-pay (WTP) for Caputo et al. (2017)
acidification	recreational fishing and logistic regression offish populations
demonstrated that under low emissions scenarios combined
with fish stocking. These are the same results reported in
Beier et al. (2017).
Aquatic	U.S. national- Using FEGS-CS to quantify the total number of causal	O'Dea et al. (2017)
acidification scale	relationships between aquatic acidification driven by N and S
deposition and beneficiaries. Effects were documented on
otter, mink, loon, aquatic vegetation, shellfish, brook trout,
and bass.
Terrestrial Northeastern Using FEGS-CS to identify 160 chains with 10 classes of Irvine et al. (2017)
acidification U.S.	human beneficiaries for balsam fir and white ash combined,
the authors concluded there are resources at risk that the
public may value. Estimated ~1,200 chains if all vulnerable
tree species are evaluated.
14.3.2. Eutrophication
1	Since the 2008 ISA, several comprehensive studies have been published on the
2	ecosystems services related to N pollution in the U.S. (Table 14-4). These include an
3	evaluation of services effected by multiple N inputs (including N deposition) to the
4	Chesapeake Bay (Birch etal.. 2011) and syntheses by Compton et al. (2011) and
5	Compton et al. (2013) of the cost-benefits on N loading across the nation for services that
6	included coastal fish harvests, recreational uses of inland and coastal waters, lakefront
7	property values, and water treatment costs. Further expanding on this work, Sobotaetal.
8	(2015) calculated the amount of N loading from human activity that is leaked into
9	adjacent ecosystems and the associated effects on ecosystem services. In their work,
10	Sobotaetal. (2015) specifically identified the costs of the atmospheric portion of total N
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loading. (Baron et al.. 2012) evaluated the interactions between climate and N,
identifying some key services in which a changing climate will modify the response to N.
Birch etal. (2011) examined the spatial and temporal movement of reactive N through
multiple ecosystems and media in the Chesapeake Bay watershed and estimated damage
costs in several categories (Figure 14-1). Their analysis estimates the contribution of
atmospheric N sources to the total annual N loads delivered to the bay estuary. In other
words, roughly 20% of N related ecosystem services impacts in the estuary are
attributable to atmospheric N deposition in the watershed.
Building on the work by Birch etal. (2011). a review and synthesis of the published
literature by Compton et al. (2011) compared estimates of the average damage costs/unit
of N across a range of affected services. Furthing building on this work, Sobotaetal.
(2015) estimated how N leaked to the environment (air/deposition, surface freshwater,
groundwater, and coastal zones) alters services by multiplying watershed-level N inputs
(8-digit U.S. Geologic Survey hydrologic unit codes; [HUC8s]) with published
coefficients describing nutrient uptake efficiency, leaching losses, and gaseous emissions.
Sobotaetal. (2015) then applied a per-unit damage cost values ($/N) for a wide variety
of ecological effects (based on estimates from various sources, Table 14-2) to the leakage
estimates.
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Why are aquatic grasses on the path between N deposition and ecosystem services in the
Chesapeake Bay?
The Chesapeake Bay is the largest estuary (4,480 sq miles) in the U.S. draining from eight major river
basins covering 64,000 sq miles of watershed in six states and a population of 18 million watershed
residents. The public's use of the bay is significant with more than 700 public access points on the bay and
its tributaries. More than 3,600 species of plants and animals are supported by the bay, including over
80 resident and migratory waterfowl. About 500 million pounds of seafood is produced annually by the
bay (https://www.chesapeakebay.net).
Unfortunately, this vital natural resource and estuarine ecosystem are under stress, and nitrogen deposition
in the Chesapeake Bay watershed is one of the main contributing sources. The Chesapeake Bay Program
reports that one-third of the nitrogen loading the bay comes from atmospheric deposition, often as nitrogen
oxides or ammonia (https://www.chesapeakebav.net/issues/air pollution). Nitrogen air emission sources
include automobiles, fossil fuel-fired electric generating units, and intensive animal operations both within
and upwind of the bay watershed.
One of the most important indicators of environmental stress to the Chesapeake Bay is the loss of
submerged aquatic vegetation (SAV). Nutrient over-enrichment and sediment overloading the bay has led
to SAV loss. Although SAV appears to be recovering as NOx emissions have declined in recent decades,
the bay still contains only an estimated 53% of the SAV it once supported. The Chesapeake Bay
Program's target restoration goal is 185,000 acres
fhttps://www.chesapeakebay.net/state/uriderwater grasses).
Photo courtesy of Chesapeake Bay Program
SAV is a critical indicator of the bay's health in part because of the essential role it plays in supporting a
wide range of final ecosystem services to humans. The 80,000 acres of underwater grasses protect and
host young and molting blue crabs and juvenile, for example
(https://wwyv.chesapeakebav.net/discover/facts). SAV supports natural ecosystems in additional ways,
including serving as a food source for wildlife, waterfowl, and humans and as habitat (and protection) for
wildlife and waterfowl (Barbier et al.. 2011; Kemp et al.. 2005) (Johnston et al- 2002). In turn, SAV
facilitates important recreation, tourism, and commerce, including wildlife and waterfowl watching,
fishing, and crabbing. In addition to these valuable indicators of overall WQ and ecosystem health, SAV
deters and dissipates wave energy and promotes sedimentation for landowners along the bay, thus,
mitigating erosion and protecting and sustaining land value.
The SAV connection is, therefore, one important pathway through which N deposition affects the well-
being of human populations in and around the bay.
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Chemical Nitrogen Cascade: Chesapeake Bay Watershed
I! mi Villi in

Utilmes
*2,0M
In&uniy
«,WM1
Mobile S.UL6
170,0X1
<3iie7 iixircev
14,DM
r
SH,_ Umbiinm

Mm-Afjiwltirf
KjOflQ
FIGURE 1. Chemical Nriroqen Cascade in tfce Chesapeake Bny Walenkts" IteucvyeBrt- S« S3 for s-oinces and colculaticMK.
N = nitrogen; N.E. = no estimate available; N20 = nitrous oxide; NH3 = ammonia; NOx = nitrogen oxide.
Source: Birch et al. (2011).
Figure 14-1 Economic nitrogen cascade in the Chesapeake Bay Watershed.
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Table 14-2 Potential damage costs of nitrogen (N) ($/kg N; 2008 or as reported)
to air, land, and water resources in the conterminous U.S. in the early
2000s as synthesized by Sobota et al. (2015). Low, median, and high
costs derive from the specific damage cost reference. Negative
values indicate an economic benefit.
Cost ($/kg N)
N damage type
System
Low
Median
High
Reference
From atmospheric NOx
Increased incidence of
respiratory disease
Air/Climate
12.88
23.10
38.63
Van Grinsven et al. (2013)
Birch et al. (2011)
Declining visibility—loss of
aesthetics
Air/Climate
0.31
0.31
0.31
Birch et al. (2011)
Increased effects of airborne
particulates/increased carbon
sequestration in forests
(includes benefits)
Air/Climate
-11.59
-4.51
2.58
Van Grinsven et al. (2013)
Increased damages to
buildings from acid
Land
0.09
0.09
0.09
Birch et al. (2011)
Increased ozone exposure to
crops
Land
1.29
1.51
2.58
Van Grinsven et al. (2013) Van
Grinsven etal. (2013): Birch et
al. (2011)
Increased ozone exposure to
forests
Land
0.89
0.89
0.89
Birch et al. (2011)
Increased loss of plant
biodiversity from N
enrichment
Land
2.58
7.73
12.88
Van Grinsven etal. (2013)
From atmosphere NH3
Increased incidence of
respiratory disease
Air/Climate
2.58
4.93
25.75
Van Grinsven et al. (2013) Van
Grinsven etal. (2013): Birch et
al. (2011)
Declining visibility—loss of
aesthetics
Air/Climate
0.31
0.31
0.31
Birch etal. (2011)
Increased effects of airborne
particulates/increased carbon
sequestration in forests
(includes benefits)
Air/Climate
-3.86
-1.93
-1.93
Van Grinsven et al. (2013)
Increased damages to
buildings from particulates
Land
0.09
0.09
0.09
Birch et al. (2011)
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Table 14-2 (Continued): Potential damage costs of nitrogen (N) ($/kg N; 2008 or as
reported) to air, land, and water resources in the
conterminous U.S. in the early 2000s as synthesized by
{Sobota, 2015, 2944570@@author-year}. Low, median,
and high costs derive from the specific damage cost
reference. Negative values indicate an economic benefit.



Cost ($/kg N)


N damage type
System
Low
Median
High
Reference
Increased loss of plant
biodiversity
Land
2.58
7.73
12.88
Van Grinsven etal. (2013)
From N2O
Increased ultra-violet light
exposure from
ozone—humans
Air/Climate
1.29
1.33
3.86
Van Grinsven et al. (2013)
Compton et al. (2011)
Increased emission of
greenhouse gas
Air/Climate
5.15
13.52
21.89
Van Grinsven et al. (2013)
Increased ultra-violet
exposure from ozone—crops
Air/Climate
1.33
1.33
1.33
Birch etal. (2011)
From surface freshwater N loading
Declining waterfront property
value
Freshwater
0.21
0.21
0.21
Dodds et al. (2009)
Loss of recreational use
Freshwater
0.17
0.17
0.17
Dodds et al. (2009)
Loss of endangered species
Freshwater
0.01
0.01
0.01
Dodds et al. (2009)
Increased eutrophication
Freshwater
6.44
16.10
25.75
Van Grinsven etal. (2013)
Compton et al. (2011)
Undesirable odor and taste
Drinking
water
0.14
0.14
0.14
Kusiima and Powers (2010)
Nitrate contamination
Drinking
water
0.54
0.54
0.54
Compton et al. (2011)
Increased colon cancer risk
Drinking
water
1.76
1.76
5.15
Van Grinsven etal. (2013)
From groundwater N loading
Undesirable odor and taste
Drinking
water
0.14
0.14
0.14
Kusiima and Powers (2010)
Nitrate contamination
Drinking
water
0.54
0.54
0.54
Compton et al. (2011)
Increased colon cancer risk
Drinking
water
1.76
1.76
5.15
Van Grinsven etal. (2013)
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Table 14-2 (Continued): Potential damage costs of nitrogen (N) ($/kg N; 2008 or as
reported) to air, land, and water resources in the
conterminous U.S. in the early 2000s as synthesized by
{Sobota, 2015, 2944570@@author-year}. Low, median,
and high costs derive from the specific damage cost
reference. Negative values indicate an economic benefit.



Cost ($/kg N)


N damage type
System
Low
Median
High
Reference
From coastal N loading
Loss of recreational use
Coastal
zone
6.38
6.38
6.38
Birch et al. (2011)
Declines in fisheries and
estuarine/marine habitat
Coastal
zone
6.00
15.84a
26.00
Van Grinsven et al. (2013) Van
Grinsven et al. (2013):
ComDton et al. (2011)
aExcludes $56/kg N from submerged aquatic vegetation loss in the Gulf of Mexico from Compton et al. (2011).
Still another advance is offered by Clark et al. (2017)and Rhodes et al. (2017). who
synthesized information on N induced terrestrial and aquatic eutrophication, respectively,
from the published literature to link the ecological effects of critical load exceedances
with human beneficiaries by using STEPS and FEGS-CS. Rhodes et al. (2017) identified
that N loading to aquatic ecosystems affects 154 chains that link changes in biological
indicators of aquatic eutrophication (a shift in phytoplankton community) to FEGS,
13 ecological production functions (EPF) within three different ecosystems (alpine lakes,
lakes, and estuaries) and 18 classes of human beneficiaries that potentially will be
effected by a change in one of these endpoints. For terrestrial eutrophication, Clark et al.
(2017) identified that N critical load exceedences affected beneficiary types through
582 individual chains in the five ecoregions examined (Eastern Temperate Forests,
Marine West Coast Forests, Northwestern Forested Mountains, North American Deserts,
Mediterranean California) and 66 FEGS across a range of final ecosystem services
categories (21 categories; e.g., changes in timber production, fire regimes, and native
plant and animal communities) (Figure 14-3 and Table 14-3).
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Level I CEC ecoregions
Eastern Temperate Forests	North American Deserts
Marine West Coast Foreste M Northwestern Foisted Mountains
Mediterranean California	01 tier Ecoregions
Map of (a) Level 1 ecoregions examined and (b) the count of unique chains, biological indicators, Final Ecosystem Goods and
Services, and beneficiaries for each ecoregion.
Source: Clark et al. (2017).
Figure 14-2 Map of ecosystem services altered by nitrogen critical load
exceedance.
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Table 14-3 Numbers of chains, Final Ecosystem Goods and Services (FEGS),
and beneficiaries (bens) associated with each initial biological
indicator Clark et al. (2017).
Biological Indicator
No. Chains
No. FEGS
No.
Benefits
Increased grass-to-forb ratio and/or increase in total biomass
192
11
12
Decreased lichen biodiversity
62
7
11
Decreased native mycorrhizal diversity
61
5
10
Decreased abundance of ectomycorrhizal fungi
43
3
12
Increased bark beetle abundance
34
4
12
Decreased survival of bigtooth aspen
32
4
11
Decreased survival of scarlet oak
32
4
11
Decreased survival of trembling aspen
32
4
11
Decreased abundance of arbuscular mycorrhizal fungi
25
3
10
Decreased growth of red pine
25
3
10
Increased cover of understory nitrophilic species
18
2
10
Increased bacteria-to-fungi ratio
17
2
9
Change in herbaceous community composition
7
1
7
Increase in N leaching
2
1
2
FEGS = Final Ecosystem Goods and Services.
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Table 14-4 Ecosystem services research related to Nitrogen (N)-driven
eutrophication.
Reference
Mode	Region	Effect	HERO ID
Estuarine N Chesapeake Recreation and commercial fishing are the main ecosystem Birch et al. (2011)
driven	Bay	services impacts; however, it is difficult to quantify these
eutrophication	impacts due to data limitations. The analysis finds that,
although atmospheric releases of N are lower than direct
releases to land and water, their total damages are larger yet
have associated abatement costs that are relatively low.
Includes	Damage to services associated with productivity, biodiversity, Compton et al.
acidification	recreation, and clean water are less certain and although (2011)
and	generally lower, these costs are quite variable (<$2.2-$56
eutrophication	per kg N). In the current Chesapeake Bay restoration effort,
for example, the collection of available damage costs clearly
exceeds the projected abatement costs to reduce N loads to
the Bay.
Includes	National	Annual damage costs associated with anthropogenic N
acidification scale,	leakage range from $1.94 to $2,255/hectare per year,
and	calculated for Nationally, the total quantifiable damages were estimated to
eutrophication 8-digit HUCs be between $81-$144 billion per year. Between 14-24% of
the potential damage costs were associated with fossil fuel
combustion. Areas with the largest damage costs
corresponded to areas with the largest N inputs and
leakages, such as the upper Midwest and central California.
Sobota et al. (2015)
Terrestrial N National scale N critical load exceedences affected beneficiary types	Clark et al. (2017)
driven	through 582 individual and 66 FEGS across a range of final
eutrophication	ecosystem services categories.
Aquatic N National scale Using STEPS and FEGS-CS, the authors identify 154 chains Rhodes et al. (2017)
driven	that link changes in biological indicators of aquatic
eutrophication	eutrophication (a shift in phytoplankton community) to FEGS,
and 18 classes of human beneficiaries.
14.3.3. Nitrogen and Climate Modification
1	Some work has discussed the implications of climate change and N loading on ecosystem
2	services such as fish harvests, property values, water treatment, human and wildlife
3	health, ocean acidification, and freshwater diversity (Baron et al.. 2012). The combined
4	effect of N loading and climate change on the economic value of water resources and
5	related products has yet to be evaluated, and even the separate economic effects of N
6	loading or climate change are difficult to determine. However, climate change will
7	modify the effects of N loading on some key ecological services. Porter etal. (2013) and
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Compton et al. (2013) also identified some effects on the interactions between nitrogen
and climate, which include:
•	As eutrophication increases with warmer water temperatures, there will be costs
associated with upgrades of municipal drinking water treatment facilities, the
purchase of bottled water, and the health costs of NO, in drinking water leading
to toxicity and disease (Compton et al.. 2011).
•	The U.S. will increasingly rely on groundwater for drinking water under future
climate change scenarios (U.S. Global Change Research Program. 2009). creating
a strong potential for increased costs for treating exposure to NOs -stimulated
disease.
•	Increasingly, N enrichment is correlated in waters with pathogen abundance and
human and wildlife diseases (Johnson et al.. 2010). Climate warming opens the
possibility for more vector-transmitted diseases to migrate to higher latitudes,
where N loading may enhance their success (Johnson et al.. 2010).
•	Acidification causes direct harm to calcifying shellfish and crustaceans (Howarth
et al.. 2012). Changes in climate and the N cycle will intensify ocean acidification,
and there are feedbacks from acidification to N cycling (Donev et al.. 2009).
Table 14-5 provides a summary of recent studies and findings regarding the
economic effects of ocean acidification on U.S. fisheries.
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Table 14-5 Summary of recent literature examining economic impacts of ocean
acidification on U.S. fisheries.
Region
Type of Fish
Key Findings
Reference
Northeast
Crustaceans
Ocean acidification alone results in very minor changes in
Ainsworth et al.
Pacific from
(especially
landings relative to the 2060 baseline projections; however,
(2011)
northern
shrimp),
ocean acidification in combination with three other climate

California to
echinoderms,
change effects (primary productivity, zooplankton

southeast
molluscs, and
community structure, and ocean deoxygenation) reduces

Alaska
euphaeids
landings by about 20%, which suggests that synergies exist.

U.S.
Sea scallops
Increased scallop growth rates from warming predicted to
Coolev et al.


outweigh decreasing growth rates from ocean acidification
(2015)


until 2030, at which point the negative influence of ocean



acidification will become the dominant effect.



After 2030, fewer scallops will attain largest ("U10")size



before they are harvested in current fishing levels are



maintained.

U.S.
Various
Considerable revenue declines, job losses, and indirect
Coolev and

commercial
economic costs are possible due to ocean acidification
Donev (2009)

species with a
broadly damaging marine habitats, altering marine resource


focus on molluscs
availability, and disrupting other ecosystem services.



Under a moderate 2% discount rate, U.S. ex-vessel revenue



losses are predicted to be $0.6-$2.6 billion and broader



economic losses are predicted to be $1.5-$6.4 billion.

U.S.
Shellfish
Analysis of sensitivity and adaptive capacity to OA across
Ekstrom et al.


23 regions indicates that the most socially vulnerable
(2015)


communities can be found on the U.S. East Coast and Gulf



of Mexico.



The East Coast is vulnerable mostly due to economic



dependence whereas the Gulf of Mexico is vulnerable



because of low adaptive capacity.

Washington,
Shellfish
Survey results from a sample of shellfish industry
Mabardv et al.
Oregon,

participants (70% owners or managers of hatcheries)
(2015)
California

indicated that about half had experienced negative impacts



from OA and that more than half the industry felt they would



be somewhat or definitely able to adapt to OA.

Global,
Shellfish
The economic cost of mollusc loss estimates for the U.S.
Narita et al.
disaggregated

are estimated to be roughly $400 million USD, assuming no
(2012)
by region

change in income but taking into account welfare losses



from shellfish price changes.

14.4. European and Canadian Applications
Since the 2008 ISA, several qualitative and quantitative assessments of either N
deposition or the combination ofN and S deposition have been conducted in the U.K.
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(Jones et al.. 2014). across Europe (dc Vries et al.. 2014b; de Vries et al.. 2014c; Van
Grinsven et al.. 2013). and Canada (Ahcrnc and Posch. 2013). Although these
assessments have varied considerably in their approaches, all have used simplified
methods that intentionally omitted much of the mechanistic and spatial complexity in
how deposition affects ecosystem services. Some of the qualitative and conceptual
assessments provided descriptions of impacts of N deposition on ecosystems (de Vries et
al.. 2014b; de Vries et al.. 2014c; Jones et al.. 2014).
In Canada, Aherne and Posch (2013) examined critical load exceedances to assess the
impacts of N and S deposition on water quality (eutrophication), soil quality
(acidification), and plant species diversity. The approach was quantitative, but did not
assign values for ecosystem services. Instead, the analysis focused on the sources of these
emissions and understanding what regions within Canada faced a loss of these services
because critical loads had been exceeded.
Studies in Europe found that N deposition has both positive and negative effects on
ecosystem services, de Vries et al. (2014b) modeled ecosystem response to N deposition,
using the Very Simple Dynamic soil model. The model quantified changes in these
services for the 2000-2050 period using three emissions scenarios: 1980 emissions,
current legislation, and a scenario where deposition does not exceed critical loads for any
ecosystem. The biggest differences among the emission scenarios were that regulations
imposed since the 1980s created significant decreases in areas receiving critical N loads
for water NOs and Al, but also likely decreased CO2 uptake by -20%. A second study by
Van Grinsven et al. (2013) conducted economic analysis of the influence of N cycling in
Europe and found that most economic impacts were related to atmospheric N emissions.
Although there were positive impacts on agricultural production from fertilization, there
were large negative impacts of atmospheric N emissions on human health (direct effects
on human health are outside the scope of the secondary NAAQS review) and ecosystem
function. Overall, the costs of added reactive N (atmospheric and agricultural) in the
European Union were between Ł75 and Ł485 billion/year. Within the U.K., Jones et al.
(2014) conducted a quantitative analysis that focused on six main ecosystem services
impact pathways (Figure 14-3): timber production, livestock production, carbon
sequestration, greenhouse gas (GHG) emissions, recreational fishing, and biodiversity
appreciation. For the first three pathways, declining N deposition is estimated to result in
ecosystem services losses of Ł27.2 million/year, due to decreased fertilization of
woodlands, grasslands, and heathland. For the other three pathways, declining deposition
is estimated to result in ecosystem services gains totaling Ł93 million/year, due to
decreased GHG emissions, improved water quality for recreational fishing, and enhanced
nonuse values for biodiversity through habitat protection. The estimated net gains to U.K.
society are, therefore, Ł65.8 million/year, with a 95% confidence interval of
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Ł15.1—Ł123.2 million pounds/year. Because of the large number of assumptions and
simplifications, these studies present a broad conceptual analyses demonstrating the
potential of ecosystem services to quantify the impacts of reactive N in ways that could
be valuable for policymakers, rather than as definitive assessments.
Benefltsfcosts from declining nitrogen deposition, 1987-2007
200
-100
Lower bound
estimate
Central estimate
Upper bound
estimate
C02 = carbon dioxide; N20 = nitrous oxide; recr = recreational.
Source: Jones et al. (2014).
Biodiversity
¦	Recr. fishing
N20 emissions
¦	C02 sequestration
¦	Livestock
¦	Timber
Figure 14-3 Benefits and costs associated with the 25% decline in nitrogen
deposition in the U.K. since 1990.
14.5. Global Perspective
There is one study on ecosystem services with a global perspective published since 2008
and one study published in 2005 not included in the 2008 ISA. Alcamo et al. (2005)
modeled global and regional impacts of stressors that included critical load exceedance
for acidification and eutrophication on provisions, regulations, support, and cultural
ecosystem services. However, these impacts represent the combined effects of eight
drivers (including climate change, land use change, mineral extractions, and N fertilizer
use). Consequently, the analysis does not define ecosystem services changes that are
specifically attributable to changes inNorS deposition.
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Erisman et al. (2013) examined the impact of humans on the global N cycle and how
these impacts relate to changes in ecosystem services. They group services into the
following categories:
•	Drinking water (nitrates),
•	Air pollution effects on human health,
•	Air pollution effects on crops (due to ozone),
•	Freshwater eutrophication (defined as areas where the concentration of nitrate
exceeds 1 mg NO3 -N/L) (UNEP. 2007; Camargo and Alonso. 2006; Vorosmartv
et al.. 2005).
•	Biodiversity loss (they identify the global critical load for biodiversity loss as
5-10 kg/ha/yr) (Bobbink et al.. 2010),
•	Stratospheric ozone depletion,
•	Changes in climate, and
•	Coastal dead zones.
It is especially noteworthy that the authors selected thresholds of level of N addition to an
ecosystem which begins to cause adverse effects. The paper concludes that better
quantitative relationships need to be established between N and the effects on ecosystems
at smaller scales, including a better understanding of how N shortages can affect certain
populations.
14.6. Summary
Since 2008, several studies have identified a range of ways in which N-S deposition
could affect socially valuable ecosystem services. Several new ecosystem services
frameworks and classification systems have been published, including The Economics of
Ecosystems and Biodiversity [TEEB; (Sukhdev et al.. 2014)1. the Common International
Classification of Ecosystem Services [CICES; (Haines-Young and Potschin. 2011)1. and
U.S. EPA Final Ecosystem Goods and Services Classification System [FEGS-CS;
(Landers and Nahlik. 2013)1.
For acidification, the literature since the 2008 ISA includes studies that better
characterize ecosystem services valuation by pairing biogeochemical modeling and
benefit transfer equations informed by WTP surveys, especially for the Adirondacks and
Shenandoah regions (Table 14-1). Aside from valuation, the estimate of the total number
of ecosystem services affected by N and S deposition is more well quantified by the new
studies that use FEG-CS. However, for many regions and specific services, poorly
characterized dose-response between deposition, ecological effect, and services are the
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greatest challenge in developing specific data on the economic benefits of emission
reductions (NAPAP. 2011).
In the 2008 ISA there were no publications specifically evaluating the effects of N
deposition on ecosystem services associated with N driven eutrophication. Since the 2008
ISA, several comprehensive studies have been published on the ecosystems services
related to N pollution in the U.S. (Table 14-4). These include an evaluation of services
effected by multiple N inputs (including N deposition) to the Chesapeake, a syntheses of
the cost-benefits on N loading across the nation, and analysis of the amount of N leaked
from its intended use and the effects on ecosystem services. In their work (this work
specifically identified the costs of the atmospheric portion of total N loading). Aside from
valuation, the estimate of the total number of ecosystem services affected by N is more
well quantified by the new studies that use FEG-CS (Bell et al.. 2017; Clark et al.. 2017;
Irvine et al.. 2017; O'Dea et al.. 2017; Rhodes et al.. 2017). In these analyses critical load
exceedances for N related air pollution were used as a model stressor from which a total
of 1,104 unique chains linking stressor to beneficiary were identified.
In the time since the 2008 ISA, several qualitative and quantitative assessments of either
N deposition or the combination of N and S deposition have been conducted in the U.K.,
across Europe, Canada, and at the global scale. Although these assessments have varied
considerably in their approaches, all were intentionally simplified to neglect much of the
mechanistic and spatial complexity on how N deposition affects ecosystem services. At
the global scale, an analysis selected general quantitative thresholds ofN loading for the
onset of ecologically adverse effects. However, the paper concludes that better
quantitative relationships need to be established between N and the effects on ecosystems
at smaller scales.
The conclusions considering the full body of literature are that (1) there is evidence that
N and S emissions/deposition have a range of effects on U.S. ecosystem services and
their social value; (2) there are some economic studies that demonstrate such effects in
broad terms; however, it remains methodologically difficult to derive economic costs and
benefits associated with specific regulatory decisions/standards; and (3) thousands of
causal relationships are now documented between N and S air pollution and changes in
Final Ecosystem Goods and Services. There are resources in the literature that would be
useful in quatification and valuation, but would require additional analyses to connect to
N and S deposition.
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14.7. Supplemental Materials: Ecosystem Services Profiles of
Select Species
1	Several species profiles have been created to better characterize the ecosystem services
2	provided by species affected by NOy and SOx. The species included here are identified as
3	threatened and endangered with N identified as a contributing stressor (Hernandez et al..
4	2016). The species are: balsam fir (Abies balsamea), eelgrass (Zostera marina), green
5	turtle (Chelonia mydas), white ash (Fraxinus americana), and lace lichen (Ramalina
6	menziesii). These profiles identify the geographic distribution within the U.S., ecological
7	function, Class I areas, FEGS, and cultural importance; information on economic
8	valuation is presented when available.
14.7.1. Balsam Fir
9	Scientific Name: Abies balsamea.
10	• Family: Pinaceae (pine).
11	Symbolic Role: NA.
12	Federal or State Threatened or Endangered Species Listing Status: Endangered
13	species in Connecticut.
14	Geographic Range/Distribution:
15	• Native and Current Range: The range of balsam fir is North America, including
16	most of eastern Canada. In the U.S., it extends from Minnesota to Maine and
17	south into central Pennsylvania and the mountains of Virginia and West Virginia.
18	Of the two main varieties—balsamea and phanerolepis—it is the latter that is
19	found in Virginia and West Virginia.
20	Overlap with Class I Areas:
21	• http://www.epa.gov/visibilitv/maps.html.
22	• http://www.cpa.gov/ttn/oarpg/t 1 /fir notices/classimp. gif.
23	Habitat: Grows best in areas characterized by cool temperatures and moist soils. In the
24	U.S., it is most commonly found in mixed stands, where associates include red spruce,
25	black spruce, paper birch, aspen, and red maple (Uchvtil. 1991; USDA. 1990b).
26	Primary Threats:
27	• Related to N or S Deposition: none
28	• Other: Balsam fir are particularly vulnerable to insect damage from spruce
29	budworms (Uchvtil. 1991) and from decay caused by the red heart fungus
30	[Haematostereum sanguinolentum (USDA. 1990b)I.
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Ecosystem Role and Function:
•	Nutrition: Provides some nutrition for mice, voles, red squirrels, and grouse
OJSDA. 1990b).
•	Cover: Provides winter cover to ungulates, including white-tailed deer and
moose, and to grouse and songbirds (Uchvtil. 1991). Also provides cover for
small mammals including martens and snowshoe hares (Darvcau et al.. 2001; de
Bellefeuille et al.. 2001; Darveau et al.. 1998).
•	Other: As with other tree species in general, balsam fir trees provide a number of
other intermediate ecosystem services that are indirectly valued by humans
including nutrient cycling, carbon sequestration, soil stabilization and erosion
control, water regulation and filtration, and air pollutant filtration (Kricgcr. 2001).
Final Ecosystem Services:
•	Direct Human Uses:
•	Raw material in wood products and other industries: Used as pulpwood
and lumber for a variety of products. Oleoresin from balsam fir bark blisters
has various uses including as a medium for mounting microscopic specimens
and cement for optical systems (USDA. 1990b). It is also used in developing
fragrances (Zcrbc et al.. 2012).
•	Energy/fuel source: Wood wastes for producers using balsam fir wood or
pulp are sometimes used for energy (USDA. 1990b).
•	Cultural and human health. Balsam fir are commonly grown and used as
Christmas trees and as material for wreath-making. In 2009 they accounted
for 7% of Christmas tree sales in the U.S. (USDA. 2010). A large number of
ethnobotanical uses of balsam fir by Native Americans has also been
documented. The Native American Ethnobotany database maintained by
University of Michigan (2015) references documents citing 87 specific uses
by 13 different tribes, including a wide variety of medicinal uses
(e.g., dermatological and venereal aid; cold and cough remedy) and uses as a
material for rugs and bedding, sewing, and canoe waterproofing.
14.7.2. Eel Grass
Scientific Name: Zostera marina (Zosteraceae).
Symbolic, Federal or State Threatened or Endangered Species Listing Status: Listed
as a conservation species of least concern.
Geographic Range/Distribution:
• Native Range: Eelgrass is native to coastal regions of both the Pacific and
Atlantic Oceans along the coast of North America.
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•	Current Range: Eelgrass is currently found in coastal regions in the Pacific
Ocean along the western U.S. and Canada. It is also found along the eastern coast
of North America from Canada to the Mid-Atlantic region in the U.S. (LJSDA.
2015a). Seagrasses have been disappearing from their native range at an average
rate of 110 km2/year since 1980 (Wavcott et al.. 2009) and are especially sensitive
to changes in temperature along the southern limits of their current range in North
America (Kaldv. 2014).
Overlap with Class I Areas:
•	http://www.epa.gov/visibilitv/maps.html.
•	http://www.cpa.gov/ttn/oarpg/t 1 /fir notices/classimp. gif.
Habitat: Coastlines in both the Atlantic and Pacific oceans. It can be found in bays,
estuaries, lagoons, and beaches.
Primary Threats:
•	Related to N or S Deposition: (Hernandez et al.. 2016).
•	Other: Threats to eelgrass include dredging for coastal development, overfishing
of predator species that lead to losses of herbivores that clean the leaves, and
sediment and nutrient loading from coastal development (Wavcott et al.. 2009;
Orth et al.. 2006). More broad changes in climate, such as temperature and
sea-level rise also contribute to the loss of seagrasses (Kaldv. 2014; Douglass et
al.. 2010).
Ecosystem Role and Function: Eelgrass are an important marine species for primary
productivity (Duartc. 2013). and provides a suitable reproductive habitat and nursery
grounds for many species of fish and shellfish (Barbier et al.. 2011). Birds, invertebrates,
green turtles, and manatees use eelgrass as a food source. Some commercially viable fish
species like pacific salmon, pacific herring, and Dungeness crabs rely on eelgrass to
provide cover and habitat (Plummer et al.. 2013; Orth et al.. 2006). Another function of
eelgrass is its ability to improve water quality through nutrient uptake and particle
deposition. Eelgrass is an important species for coastal erosion protection because it helps
control wave attenuation (Barbier et al.. 2011). Eelgrass also provides indirect human
benefits through carbon sequestration because it uses dissolved carbon in seawater to
grow, and at the end of its lifecycle, the carbon is buried and stored as detritus (Barbier et
al.. 2011).
Final Ecosystem Services:
•	Direct Human Uses:
•	Cultural: Native Americans used dried eelgrass to bake into cakes and also
for smoking deer meat (Felger and Moser. 1985). Native Americans consume
the roots of the plant or use it to flavor other foods (University of Michigan.
2015).
•	Raw Material: Eelgrass is used to pack crabs to keep them moist during
transport in the Mid-Atlantic region of the U.S. (Barbier et al.. 2011).
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14.7.3. Green Turtle
Scientific Name: Chelonia mydas (Cheloniidae).
Symbolic Role: Green sea turtles are symbols of guidance and protection in Hawaiian
culture.
Federal or State Threatened or Endangered Species Listing Status: Listed as an
endangered species (Scminoff. 2004).
Geographic Range/Distribution:
•	Native Range: Green sea turtles have distinct native populations in both the
Pacific and Atlantic oceans.
•	Current Range: In the North American Atlantic, nesting sites are found in
coastal regions in Florida, as well as Georgia, South Carolina, and North Carolina
(IJSFWS. 2015). In the Pacific, green turtles nest along the French Frigate Shoals
in the northwestern Hawaiian Islands (NQAA Fisheries Pacific Islands Regional
Office. 2015a).
Overlap with Class I Areas:
•	http://www.epa.gov/visibilitv/maps.html.
•	http://www.epa.gov/ttn/oarpg/t 1 /fir notices/classimp. gif.
Habitat: Mature green sea turtles are found in shallow coastal regions with large
seagrass beds such as bays, reefs, and inlets. Green turtles nest on open beaches with
minimal disturbance (IJSFWS. 2015).
Primary Threats:
•	Related to N or S Deposition: (Hernandez et al.. 2016).
•	Other: Other major threats to green sea turtles include entanglement in fishing
gear, incidental by catch, and collisions with fishing boats or jet skis. Green turtles
also face threats from illegal human harvesting and habitat loss and degradation
due to human development along coastal areas and reefs (Malama na Honu.
2015).
Ecosystem Role and Function: Green sea turtles are herbivores that mostly feed on
seagrasses and algae in near-shore marine ecosystems. Green sea turtle eggs are a food
source for many coastal species like ghost crabs or marine birds, as well as other
scavenging animals. Tiger sharks are the only nonhuman predator for large juvenile and
mature turtles (NQAA Fisheries Pacific Islands Regional Office. 2015b). Green sea
turtles are valued for their natural beauty, and several conservation groups are dedicated
to increasing sea turtle populations (Komoroskc et al.. 2011). Turtles can also be a source
of ecotourism (Campbell. 2002).
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1	Final Ecosystem Services:
2	• Direct Human Uses:
3	• Cultural: Some native Hawaiians consider the green turtle to be a personal
4	or family deity, which should not be harmed. The green turtle is featured in
5	Hawaiian culture through depictions in petroglyphs (NQAA Fisheries Pacific
6	Islands Regional Office. 2015a).
14.7.4. White Ash
7	Scientific Name: Fraxinus americana.
8	• Family: Oleaceae (olive).
9	Symbolic Role: NA.
10	Federal or State Threatened or Endangered Species Listing Status: No special status.
11	Geographic Range/Distribution:
12	• Native Range: Most of eastern North America from southern Canada to northern
13	Florida and as far west as Minnesota (Griffith. 1991).
14	• Current Range: In addition to its native range, white ash is cultivated in other
15	areas, including Hawaii (Griffith. 1991).
16	Overlap with Class I Areas:
17	• http://www.epa.gov/visibilitv/maps.html.
18	• http://www.epa.gov/ttn/oarpg/t 1 /fir notices/classimp. gif.
19	Habitat: Usually found mixed with other hardwoods (not the dominant species) in
20	various topographic conditions. It can grow in a wide variety of soil types but does best
21	in soils with high nitrogen and calcium content, as well as moist and well-drained soils
22	(USD A. 1990a).
23	Primary Threats:
24	• Related to N or S Deposition: (Hernandez et al.. 2016).
25	• Other: White ash has experienced extensive dieback and mortality in its native
26	range since the 1920s, due to a variety of factors, including ash yellows (AshY
27	phytoplasma), canker fungi, viruses, drought (Hibbcn and Silverborg. 1978). and
28	more recently, invasions of the emerald ash borer [Agrilus planipennis; (Poland
29	and McCullough. 2006)1. White ash has also been found to be sensitive to ground
30	level ozone exposures (Chappelka et al.. 1988).
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Ecosystem Role and Function:
•	Nutrition: Samaras (seeds) are good forage for wood ducks, northern bobwhites,
purple finches, pine grosbeaks, fox squirrels, mice, and many other birds and
small mammals. It is browsed by white-tailed deer and cattle, and its bark is
occasionally used as food by beavers, porcupines, and rabbits (Griffith. 1991).
•	Cover: It provides cover to primary cavity nesters, such as red-headed,
red-bellied, and pileated woodpeckers, and to secondary nesters such as wood
ducks, owls, nuthatches, and gray squirrels (Griffith. 1991).
•	Other: As with other hardwood tree species in general, white ash trees provide a
number of other intermediate ecosystem services that are indirectly valued by
humans, including nutrient cycling, carbon sequestration, soil stabilization and
erosion control, water regulation and filtration, and air pollutant filtration
(Krieger. 2001).
Final Ecosystem Services:
•	Direct Human Uses (Final Ecosystem Services)
•	Raw material in wood products and related industries: White ash trees
provide a dense, shock resistant, and durable wood that is used in a variety of
products, including baseball bats, tool handles, furniture, cabinets, canoe
paddles, snowshoes, and railroad ties (USDA NRCS).
•	Energy/fuel source: Used for fuel wood (USDA NRCS).
•	Human health: Leaves from the white ash have reported beneficial uses,
including the relief of swelling and itching from mosquito bites and
prevention of snake bites (Griffith. 1991).
•	Cultural: A large number of ethnobotanical uses of white ash by Native
Americans has been documented. The Native American Ethnobotany
database maintained by University of Michigan (2015) references documents
citing 38 specific uses by 9 different tribes, including a wide variety of
medicinal uses (e.g., gastrointestinal, dermatological, and gynecological aids)
and uses as a material for snow gear, basketry, furniture, canoes, hunting, and
fishing.
•	Aesthetic: Provides attractive and vivid fall foliage, including combinations
of orange, yellow, maroon, and purple. For landscaping uses, it is also
considered to be a good and relatively quick-growing shade tree (University
of Kentucky. 2015).
•	Nonuse value: Survey-based economic studies have found evidence of
nonuse values among the general population for protecting white ash from air
pollution in the Adirondacks region (Banzhaf et al.. 2006).
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14.7.5.
Lace Lichen
Scientific Name: Ramalina menziesii (Ramalinaceae).
Federal or State Threatened or Endangered Species Listing Status: Listed as a
conservation species of least concern.
Geographic Range/Distribution:
•	Native Range: Lace lichen is native to the western coast of North America; in the
past it was found in the San Jacinto Mountains near Los Angeles and on the
surrounding coastal plain but now only grows at elevations above the smog layer
(Hastings Reserve. 2015).
•	Current Range: Lace lichen grows along the Pacific Coast of North America
from Alaska to Baja California. It grows at elevations up to 3,500 feet ("Werth and
Sork. 2008).
Overlap with Class I Areas:
•	http://www.epa.gov/visibilitv/maps.html.
•	http://www.epa.gov/ttn/oarpg/t 1 /fir notices/classimp. gif.
Habitat: Grows on tree species in coastal areas with enough moisture from fog and
strong winds (Egan and Holzman. 2003).
Primary Threats:
•	Related to N or S Deposition: (Hernandez et al.. 2016).
•	Other: Other threats to lace lichen include changes in climate. This species
requires a good amount of moisture and a temperate climate to survive; it is absent
from drier areas within its range (Egan and Holzman. 2003).
Ecosystem Role and Function: Lace lichen has a mutualistic relationship with its host
tree species. Three common host species in California are the valley oak, blue oak, and
coastal live oak (Werth and Sork. 2008). The lichen benefits from the microclimate and
structure provides by the host tree and it benefits the tree by capturing wind-borne
nutrients and moisture and depositing them back to the surrounding soil (Knops et al..
1996). This species is particularly useful in the Santa Lucia Mountains where nitrogen
limits plant growth (Hastings Reserve). Lace lichen is a food source for browsing animals
like deer and cows. It is also used as a nesting material for hummingbirds and orioles
(Hastings Reserve). Lace lichen has also been found to contribute to the global uptake of
sulfur dioxide, an airborne pollutant (Gries et al.. 1997).
Final Ecosystem Services:
•	Direct Human Uses:
• Aesthetic: Lace lichen has a unique morphology that adds to the natural
beauty of coastal woodlands along the Pacific Coast (Hastings Reserve.
2015).
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1	• Education/research: Lace lichen is very sensitive to air pollution, including
2	exposure to nitric acid (Riddcll et al.. 2008) and can be used as an indicator
3	species to measure air quality (Gciscr et al.. 2010). New growth of lichen can
4	also indicate improved air quality conditions.
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APPENDIX 15. OTHER ECOLOGICAL EFFECTS OF
PARTICULATE MATTER
Evidence for effects of particulate matter (PM) on ecological receptors include direct
effects of airborne PM on radiative flux and both direct and indirect effects of deposited
particles. In addition to nitrogen (N) and sulfur (S) and their transformation products
(discussed in Appendix 3-Appendix 12). other PM components such as trace metals and
organics are deposited to ecosystems and may subsequently impact biota. Direct effects
include alteration of leaf processes from deposition of PM ("dust") to vegetative surfaces
(Appendix 15.4.1). Indirect effects encompass physiological responses associated with
uptake of PM components and alterations to ecosystem structure and function. Exposures
can occur directly to surfaces of vegetation (leaves, bark, twigs), by ingestion, or via soil
or water through uptake by roots or other biological tissues. Bioaccumulation and
biomagnification of PM components can lead to effects at higher trophic levels. Direct
and indirect effects of PM on vegetation, fauna, and the soil environment, and evidence
for effects at higher levels of biological organization (communities, ecosystems) from
PM deposition are evaluated in the context of what was known from previous PM
assessments (U.S. EPA. 2009a. 2004) and newly available studies.
15.1. Introduction
PM-associated components include N and S and their transformation products, trace
metals, organics, base cations, and salts. Base cations (especially Ca2+, Mg2+, and K+)
from atmospheric deposition can help ameliorate the effects of acidic deposition
associated with oxides of N and S, although under very high base cation deposition, plant
health can be adversely affected (U.S. EPA. 2009a). Particulate salt may be added to an
ecosystem from deicing salt (U.S. EPA. 2009a'). The focus of this appendix is on the PM
effects not associated with N and S components that are covered elsewhere in this ISA.
The primary non-N and S components covered in this appendix include trace metals
(Appendix 15.3.1) and organics (Appendix 15.3.2).
Effects of PM on ecological receptors can be both chemical and physical (U.S. EPA.
2009a. 2004). As described in the 2009 Integrated Science Assessment for Particulate
Matter (2009 PM ISA), particulates that elicit direct and indirect effects on ecological
receptors vary in terms of size, origin, and chemical composition. Particle composition is
attributed to ecological outcomes to a greater extent than particle size (Grantz et al..
2003). PM-associated metals and organics are linked to responses in biota; however, the
heterogeneous nature of PM composition and distribution coupled with variability
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inherent in natural environments confound assessment of the ecological effects of
particulates.
Ecological effects evaluated in the 2009 PM ISA included direct effects on metabolic
processes of plants, the contribution of PM to total metal loading which alters soil
biogeochemistry and microbiology, the effects on plant and animal growth and
reproduction, and the contribution to total organics loading resulting in trophic transfer.
Newly available information on the ecological effects of PM published since the 2009
PM ISA is summarized in the following sections along with key studies from previous
PM assessments (U.S. EPA. 2009a. 2004). Studies conducted in the U.S. are the focus of
this review, but research from other countries is included if it has advanced the study of
PM effects on biota. Such studies may include those providing documentation of
additional physiological effects of PM toxicity, new techniques for PM assessment, and
further characterization of effects on communities and ecosystems. Together with the
information available in the 2009 PM ISA, the body of evidence is sufficient to infer a
likely causal relationship between deposition of PM and a variety of effects on
individual organisms and ecosystems.
Direct Effects of Particulate Matter on Radiative Flux
The direct effects of particles suspended in the atmosphere on radiative flux and
subsequent effects on vegetation were described in previous PM assessments (U.S. EPA.
2009a. 2004). Briefly, increased atmospheric particulates and aerosols can alter radiative
flux by both radiation attenuation and by changing the efficiency of radiation interception
in the vegetation canopy by converting direct radiation to diffuse radiation (Hovt. 1978).
Diffuse radiation is more uniformly distributed throughout the vegetation canopy and can
reach leaves more evenly than direct radiation, increasing canopy photosynthetic
productivity. In a study reviewed in the 2009 PM ISA, decreased crop yield in China was
associated with reduction in solar radiation due to regional haze (Chamcidcs et al.. 1999).
More recent studies provide additional evidence of the role of PM in altering radiative
flux and plant response. In an assessment of aerosol effects on plant productivity in the
eastern U.S., increased canopy photosynthesis due to increases in diffuse radiation,
despite a simultaneous decrease in direct radiation due to light scattering by aerosols, was
observed on a regional scale (Matsui et al.. 2008). Surface downwelling solar radiation
was reduced by 14.9 W/m2 in 2000 and 16.0 W/m2 in 2001, while photosynthesis and
stomatal conductance increased simultaneously. In a recent study from Beijing, China,
the expression of plant proteins important to photosynthesis were shown to be directly
impacted by presence of aerosols and PM in the air column (Yan et al.. 2014). Expression
of plant ribulose 1,5-biphosphate carboxylase/oxygenase (RuBisCO), a key enzyme
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involved in plant photosynthesis, increased in areas with strong diffuse solar radiation,
(identified by measurement of aerosol optical depth and particles with diameters of 0.1 to
1.0 |im). In the same areas, light-harvesting complex II protein and oxygen-evolving
enhancer protein decreased. These enzymes play a role in plant responses to direct solar
radiation when photosynthesis capacity is exceeded.
15.2. Particulate Matter Deposition to Ecosystems
Once airborne PM is deposited to aquatic or terrestrial systems, it can elicit additional
effects on ecosystem receptors. Large particles tend to be deposited near their sources,
such as along roadsides or next to mining, smelting, and other industrial operations, while
smaller particles can be transported long distances (Grantz et al.. 2003). Deposition to
ecosystems can be in the form of wet, dry, or occult deposition (U.S. EPA. 2009a; Grantz
et al.. 2003). Once deposited, PM-associated components may remain on biological
surfaces or be transferred between environmental compartments (e.g., water, soil,
sediment, biota).
More often, the chemical constituents drive the ecosystem response to PM, rather than
size class (U.S. EPA. 2009a; Grantz et al.. 2003). Exposure to a given mass concentration
of PM may lead to widely differing phytotoxic and other environmental outcomes
depending upon the particular mix of PM constituents involved. Deposition of the metal
and organic constituents of PM are discussed below; however, a rigorous assessment of
each chemical constituent (e.g., polyaromatic hydrocarbons [PAHs], mercury [Hg],
cadmium [Cd]) is not given. An overview of trace metals and organics and their
depositional effects in ecosystems was provided in the previous PM assessments (U.S.
EPA. 2009a. 2004) and only summarized below.
15.2.1. Metals
All but 10 of the 90 elements that comprise the inorganic fraction of the soil occur at
concentrations of <0.1% (1,000 jxg/g) and are termed "trace" elements or trace metals.
Trace metals with a density greater than 6 g/cm3, referred to as "heavy metals" (e.g., Cd,
copper [Cu], lead [Pb], chromium [Cr], Hg, nickel [Ni], zinc [Zn]), are of particular
interest because of their potential toxicity to plants and animals. For example, plant
toxicity to trace metals is most frequently associated with Cu, Ni, and Zn (U.S. EPA.
2004). Some metals such as Zn, manganese (Mn), Cu, and iron (Fe) play important
physiological roles in trace quantities; however, other metals such as Hg, Cd, Cr, and Pb
have no known biological function. Although some trace metals are essential for
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vegetative growth or animal health, they are all toxic in large quantities. Trace metals are
naturally found in small amounts in soils, groundwater, and vegetation. They may enter
the ecosystems as both fine and coarse particles. Heavy metals tend to be associated with
fine particles. Trace elements exist in the atmosphere in particulate form as metal oxides
(Ormrod. 1984). Aerosols containing trace elements derive predominantly from industrial
activities. Once deposited to biological surfaces, metals can be taken up by biota,
accumulate in tissues, and elicit toxic effects (Gall et al.. 2015). These responses are
highly variable across organisms. Only the bioavailable fraction is available for uptake by
biota. In plants, metal uptake is generally via soil to root transfer (McBridc et al.. 2013).
although recent studies reviewed in Appendix 15.4 provide evidence for foliar transfer of
metals following atmospheric deposition. The ecological effects of atmospherically
deposited Pb, a metal associated with PM, are included in the ISA for Pb (U.S. EPA.
2013b).
15.2.2. Organics
Organic compounds can be in the gas phase or in association with particles (Grantz et al..
2003). Organic compounds that may be associated with deposited PM include persistent
organic pollutants (POPs), pesticides, semivolatile organic compounds (sVOCs), PAHs,
and flame retardants, among others (U.S. EPA. 2009a). Dry deposition of organic
materials is often dominated by the course fraction, but fine PM may act as a carrier for
materials such as pesticides. PAHs are widespread in the environment. Of the
anthropogenically derived PAHs, 8 are considered carcinogenic and 16 have been
classified by the U.S. EPA as priority pollutants. In general, high molecular weight PAHs
are more toxic and have a greater tendency to be associated with PM (Mcsquita et al..
2014a). Derived from vehicular traffic and other sources, they are common air pollutants
in metropolitan areas. PAHs and other organics can be transferred to higher trophic
levels. Some atmospheric contaminants like POPs, polybrominated diphenyl ethers
(PBDEs), and other brominated flame retardants have been shown to accumulate in biota
in polar regions and at other remote locations where they are carried by long-range
atmospheric transport from lower latitudes (U.S. EPA. 2009a). The Western Airborne
Contaminants Assessment Project (WACAP) summarized in the 2009 PM ISA provided
evidence for airborne transfer and bioaccumulation of organics to remote national parks
and Class I areas (Landers et al.. 2008). and newer studies from this project have added to
this information (Pritz et al.. 2014; Landers et al.. 2010). Seven ecosystem compartments
(air, snow, water, sediments, lichens, conifer needles, and fish) were analyzed for a suite
of contaminants to determine where the pollutants were accumulating, identify ecological
indicators, and assess the source of the air masses most likely to have transported the
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contaminants to the parks. Overall, semivolatile organic compounds and metals were
detected throughout park ecosystems, and biomagnification of organics through the food
web indicated contaminants were present at multiple trophic levels.
As reviewed in previous PM assessments, vegetation itself is an important source of
hydrocarbon aerosols (U.S. EPA. 2009a. 2004). Terpenes, particularly a pinene, (3 pinene,
and limonene, released from tree foliage may react in the atmosphere to form submicron
particles. These naturally generated organic particles contribute significantly to the blue
haze aerosols formed naturally over forested areas (U.S. EPA. 2004; Geron et al.. 2000).
15.3. Effects of Particulate Matter on Vegetation
Particles deposited to foliar surfaces may lead to both structural and functional alterations
in plants. PM may physically obstruct processes associated with the leaf surface or be
taken up across the cuticle and into the plant tissues where it affects plant metabolic
activities. This section first considers direct impacts of PM deposition to vegetative
surfaces. Next, effects following uptake of PM components into plant tissue across leaf
surfaces or root uptake via soils are described. Uptake by plants can occur at the
soil/plant interface and at the air/plant interface (Krupa et al.. 2008). Once uptake occurs,
plant physiological responses may include decreased gas exchange, altered metabolism
and photosynthesis, altered pigment and mineral content, and altered enzyme activity
(Naidoo and Chirkoot. 2004). There is some evidence that metals reduce frost hardiness
and impair nutrition (Taulavuori et al.. 2005; Kim et al.. 2003).
15.3.1. Vegetative Surfaces
As described in the 2009 PM ISA, foliar surfaces are covered with a waxy cuticle layer
that helps reduce moisture loss and ultraviolet (UV) radiation stress (U.S. EPA. 2009a).
This epicuticular wax consists largely of long chain esters, polyesters, and paraffins,
which accumulate lipophilic compounds. Organic air contaminants in the particulate or
vapor phase can be adsorbed to, and accumulate in, the epicuticular wax of leaf surfaces.
Low solubility limits foliar uptake and direct heavy metal toxicity because trace metals
must be brought into solution before they can enter leaves or bark of vascular plants (U.S.
EPA. 2009a). Particles embedded in waxes may remain for extended periods of time and
remain associated with the leaf while other particles may be removed via precipitation,
wind, or leaf-fall. Particle capture efficiency varies by plant species, and the role of
vegetative surface characteristics in trapping PM, especially in polluted urban
environments, is extensively reviewed in the literature. PM trapping by vegetation
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depends on distance from source, leaf shape, canopy structure, leaf surface wettability,
presence of hairs, properties of the epidermal layer, and phyllotaxy (Popek et al.. 2015;
Popek et al.. 2013; Saebo et al.. 2012).
The 2009 PM ISA also reviewed studies that show effects of direct deposition of
particulates to aboveground plant organs. These effects include altered plant metabolism
and photosynthesis by the blocking of sunlight, obstruction of stomatal apertures,
increasing leaf temperature, and leaf surface injury (U.S. EPA. 2009a; Naidoo and
Chirkoot. 2004; Grantz et al.. 2003). Leaf surface pH can be altered by deposition of
alkaline particles, which can hydrolyze epicuticular surfaces (Grantz et al.. 2003). The
diverse composition of ambient PM and the effects of other air pollutants confound
characterization of the direct effects of PM on foliar surfaces (Grantz et al.. 2003).
A number of recent studies published since the 2009 PM ISA have examined the effects
of PM deposition to vegetative surfaces. For example, some have further characterized
the effects of particle size and composition on leaf surface processes. Following 60-day
experimental application of urban dust collected from roadsides in India to common
annual plant species (Abelmoschus esculentus, Celosia cristata, Coleus blumei,
Cyamopsis tetragonolobus, Gomphrena globosa, Impatiens balsamina, Ocimum sanctum,
Phaseolus vulgaris, Solanum melongena, Zinnia elegans), scanning electron micrographs
of leaf surfaces indicated that larger particles piled up around the stomata, while finer
particles clogged the stomatal openings (Rai et al.. 2010). In the applied dust, about 75%
was in the 2.5 to 10 |im size fraction, with course particles composing approximately
10% of PM and ultrafines comprising 15%. Experimental application of PM
(monometallic PM oxides: cadmium oxide [CdO], antimony trioxide | Sb:Ch |. and zinc
oxide [ZnO] and process PM enriched with Pb) to cabbage (Brassica oleracea) and
spinach (Spinacia oleracea) leaves resulted in a coverage rate of particles on the leaf
surface estimated at approximately 2% following washing (Xiong et al.. 2014). Most of
the particles were concentrated in stomatal apertures, with up to 12% of the area
occupied. The presence of PM in the stomata was correlated to particle size and solubility
of the associated metal oxides. Newer studies (Qguntimehin et al.. 2010; Qguntimehin et
al.. 2008) support observations of visible foliar injury reported in previous PM
assessments, with some suggesting that drought may affect the particle capture efficiency
of trees, possibly leading to decreased drought tolerance (Burkhardt and Parivar. 2015.
2014; Rasanen et al.. 2014).
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15.3.2. Foliar Uptake of Particulate Matter
As reported in the 2009 PM ISA, fine PM has been shown to enter the leaf through the
stomata and penetrate into the mesophyll layers where it alters leaf chemistry and
physiology (Da Silva et al.. 2006; Naidoo and Chirkoot. 2004). Organic compounds can
be deposited as particles on the leaves or be taken up through the cuticle or stomata in the
gas phase, although these pathways are not well characterized for most trace organics
(Oguntimchin et al.. 2010). As reviewed in the 2009 PM ISA, for lipophilic POPs, such
as polychlorinated dibenzodioxins (PCDDs) and polychlorinated biphenyls (PCBs), the
air/plant response route generally dominates (Lee et al.. 2003; Thomas et al.. 1998). but
uptake through aboveground plant tissue also occurs. The pathways depend on the
chemical and its physical properties, such as lipophilicity, water solubility, vapor
pressure, and Henry's law constant. Environmental conditions can also be important,
including temperature and organic content of soil, plant species, and the foliar surface
area and lipid content.
(Eichcrt et al.. 2008) demonstrated hydrophilic particle penetration into the substomatal
cavity of faba bean (Vicia faba) leaves. Since the 2009 PM ISA, development and
refinement of techniques such as micro x-ray fluorescence, scanning electron microscopy
coupled with energy dispersive x-ray microanalysis, and secondary ion mass
spectrometry have been applied toward characterizing particulate transfer from foliar
surfaces to the interior of the plant (Schrcck et al.. 2014; Burkhardt et al.. 2012; Schreck
et al.. 2012). Repeated deliquescence and efflorescence of soluble PM may have the
effect of slowly advancing solutes into the stomatal pore where they may become
bioavailable (Burkhardt et al.. 2012). Recently, stomatal uptake of aqueous solutions was
confirmed by environmental scanning electron microscopy (Burkhardt et al.. 2012).
Deposited aerosols on leaf surfaces may alter surface tension and hydrophobicity,
enabling stomatal liquid water transport and facilitating foliar uptake. Several studies
provide additional evidence of particles entering plant tissues through stomata openings
(Schreck et al.. 2012; Uzu et al.. 2010). Schreck et al. (2014) observed particles in
stomata and damage to guard cells from Pb-rich particles from a Pb recycling factory.
Other pathways of deposited metals to vegetative surfaces may include diffusion across
the cuticle (Schreck et al.. 2012; Uzu et al.. 2010). Nonstomatal uptake of atmospheric
Hg into plant leaves has recently been demonstrated in addition to stomatal pathways
(Stamenkovic and Gustin. 2009). Biogeochemical transformations were observed in
metal-rich PM at the leaf surface (Schreck et al.. 2012). The role of the phyllosphere
(i.e., the microbial communities on foliar surfaces) on the uptake and chemical
transformation of deposited particles and in promoting plant growth is an emerging area
of study; however, interactions associated with PM are not well characterized at this time
(Wevens et al.. 2015; Xiong et al.. 2014).
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15.3.3.
Particulate Matter Impacts on Gas Exchange Processes
In earlier assessments of ecological effects of PM, deposition of particles to vegetation
was known to interfere with gas exchange processes such as photosynthesis, respiration,
and transpiration (U.S. EPA. 2009a. 2004). Several new studies support observations
from previous PM reviews on the exchange of oxygen and carbon dioxide across the leaf
surface. Most studies involve the application of dust to leaves. A greenhouse study with
lettuce (Lactuca serriola) leaves showed decreased gas-exchange parameters (net
photosynthetic rate and stomatal conductance) and a change in transpiration rate
following PM (dust from vehicular traffic with a dominance of fine particles) application
to leaves (Pavli'k et al.. 2012). Fly ash dusting of 0.5 to 1.5 g/m2/day to rice reduced
photosynthesis, stomatal conductance, and transpiration (Raja et al.. 2014).
Photosynthesis, transpiration rate, and water use efficiency was decreased in peach
(Primus persica) leaves dusted with cement kiln dust, and to a lesser extent with soil dust
(Malctsika et al.. 2015). Exposure of the leaves of Prunus padus to PM in Poland resulted
in decreased photosynthesis and chlorophyll fluorescence, but similar effects on the
photosynthetic apparatus were not found in Prunus serotina (Popck et al.. 2017). In a
study of native and non-native plant species in Argentina, Gonzalez et al. (2014) reported
that gas-exchange parameters, including maximum assimilation rate, stomatal
conductance, and transpiration rate, were significantly decreased in most plant species by
ambient dust accumulation on leaf surfaces compared with those parameters in plants
from which deposited dust had been removed. Similarly, decreased photosynthesis rate,
increased stomatal resistance, and lower chlorophyll content was observed with increased
PM accumulation in plant species growing in the city center of Warsaw, Poland
compared to a moderately clean environment (Przvbvsz et al.. 2014). The photosynthetic
apparatus of the lichen species Evernia prunastri was significantly affected by dust
enriched by calcium (Ca), Fe, and titanium (Ti) near a quarry and cement factory in
Slovakia (Paoli et al.. 2015). In open-top chamber experiments, leaf net photosynthesis
rate decreased in maize (Zea mays) plants exposed to 50 ng Hg/m3 compared to plants
exposed to 2 ng Hg/m3 (Niu et al.. 2014). Photosynthetic rate and chlorophyll content
were significantly reduced in cherry tomatoes (Lycopersicon esculentum) misted with
fluoranthene for 30 days to represent atmospheric wet deposition to the leaf surface
(Oguntimehin et al.. 2010).
15.3.4. Plant Physiology
In the 2009 PM ISA, particulates were associated with phytotoxic responses, including
induction of phytochelatins, alteration of pigments, and changes in mineral content and
enzyme activity (U.S. EPA. 2009a). Newly available studies support these findings.
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Chlorophyll content as well as specific leaf area and relative water content of selected
tree species was significantly reduced by the greater dust load associated with a polluted
site in India compared to a low pollution location (Chaturvcdi ct al.. 2013). Decreased
leaf chlorophyll and increased phenol content was also observed in peach leaves with
accumulated cement or soil dust (Malctsika et al.. 2015). In lettuce leaves, PM
application caused a depletion in amino acids that are metabolized during photosynthesis
as well as an increase in proline (Pavli'k et al.. 2012). Proline and malondialdehyde
concentrations were elevated in maize leaves treated with Hg (Niu et al.. 2014). Recently,
the use of foliar fatty acid composition as an indicator of foliar (and root) metal uptake
was demonstrated with lettuce leaves exposed to PM from smelter emissions (Schrecket
al.. 2013).
15.3.5. Uptake of Particulate Matter by Plants from Soils
Plants can also uptake PM that has been deposited to soils. As reviewed in the 2009 PM,
ISA uptake from soils varies by the PM component and plant species (U.S. EPA. 2009a).
There was some evidence that shallow-rooted plant species are most likely to take up
metals from the soil (Martin and Coughtrev. 1981). The ability of plants to take up
contaminants from soil and water has been applied toward environmental cleanup efforts,
a process known as phytoremediation (Hooda. 2007; Padmavathiamma and Li. 2007;
Clemens. 2006). Plants that are hyperaccumulators of metals are especially useful in
phytoremediation efforts (Prasad and DeOliveira. 2003).
Plants respond to high concentrations of metals in soil through a variety of mechanisms,
and plant species differ substantially in their response to heavy metal exposure. As
reviewed in the 2009 PM ISA, mechanisms of metal tolerance included exclusion,
excretion, genetics (Yang et al.. 2005; Patra et al.. 2004). mycorrhizal interactions (Gohre
and Paszkowski. 2006). storage capability and accumulation (Clemens. 2006). various
cellular detoxification mechanisms (Gratao et al.. 2005; Hall. 2002). and chelation with
phytochelatins (Memon and Schroder. 2009).
Since the 2009 PM ISA, additional studies have focused on the use of vegetation to
remediate sites with both organic and inorganic contaminants and the use of mycorrhizal
fungi and bacteria to promote phytoremediation (Appendix 15.5.4). In addition,
soil-bound PAHs associated with soil organic matter, historically thought to be generally
not easily available for root uptake, have been shown in recent studies to be readily
adsorbed to root surfaces, with some evidence showing translocation to the shoots
(Desalme et al.. 2013).
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15.3.6.
Effects on Plant Growth and Reproduction
PM effects on plant growth reviewed in the 2009 PM ISA include reduced vigor and
impaired root development (U.S. EPA. 2009a). In general, plant biomass is negatively
correlated with organics (Dcsalmc et al.. 2013) and metal concentrations (Audet and
Charcst. 2007). Inhibition of aboveground biomass was observed in lettuce leaves with
traffic-associated PM applied directly to vegetation (Pavlik et al.. 2012). Increased trace
metal concentration along with the decrease in biomass was measured in the treated
leaves. Decreased primary root elongation and shoot and root rates were observed in
tomatoes (Solarium lycopersicum) grown for 18 days in particulate matter with a nominal
mean aerodynamic diameter less than or equal to 10 ^m (PMio) collected from an urban
background site in Italy (Darcsta et al.. 2014). Decreased chlorophyll a and increased
carotenoids in the exposed plants were noted along with increased reactive oxygen
species production in roots from the PMio substrate. Phenanthrene, a common PAH in
air, applied to soils via simulated atmospheric deposition in an exposure chamber was
shown to significantly decrease root length and shoot biomass of leek (Allium porrum)
seedlings (Dcsalmc et al.. 2012). In a similar study, shoot biomass of red clover
(Trifolium pratense) exposed to atmospheric phenanthrene decreased by around 30%
while ryegrass (Loliumperenne) was unaffected (Dcsalmc et al.. 201 la). No effects on
root biomass were observed in either plant. With phenanthrene exposure to red clover,
more carbon (C) was retained in leaves with decreased C allocation to stems and roots
(Desalme et al.. 201 lb).
Daily application of urban dust to common annual plant species for 60 days showed
reduction in plant growth, number of leaves, and leaf area (Rai et al.. 2010). Growth and
yield of rice was significantly influenced at higher rates of fly ash deposition due to
increased heat load and reduced intracellular carbon dioxide (CO2) concentration (Raja et
al.. 2014). Following dusting of the plants by fly ash at 0.5, 1.0, and 1.5 g/m2/day, a
significant decrease in grain yield of 12.3, 15.7, and 20.2%, respectively, was observed.
Farahat et al. (2016) studied the effects of dust deposition from a rock quarry on
old-growth eastern hemlock (Tsuga canadensis) in Quebec, Canada. Mean radial growth
declined 43% after the construction of the quarry compared with growth rates before
construction. Major changes in the hemlock occurred 3-16 years after the quarry was
established, indicating that the effect on growth may be indirect through the alteration of
soil pH in this forest In a study from Beijing, China, application of local road dust to
Sophora japonica seedlings significantly affected growth characteristics such as leaf N,
shoot biomass, root-shoot ratio, photosynthesis, and total chlorophyll (Bao et al.. 2016;
Bao et al.. 2015).
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Recently, the effect of air pollutants on the timing and budburst of vegetation has been
assessed in several studies (Jochncr et al.. 2015; Kozlov et al.. 2007). The timing of full
flowering of hazel (Corylus avellana) was found to be significantly related to PM in
ambient air in urban areas of Munich, Germany (Jochncr et al.. 2015).
These effects on phenological onset were also observed with ozone (O3), nitrogen dioxide
(NO2), and nitrate (NO3) in hazel and additional tree species in the same study. In a field
and greenhouse study in Russia with mountain birch (Betula pubescens), no impacts of
air pollution on phenology were observed between trees surveyed in proximity to a Ni-Cu
smelter and trees at further distance from the emission source (Kozlov et al.. 2007).
Some studies have examined the effects of PM on plant reproduction. Some potential
effects demonstrated were those in flowering phenology and timing (Jochner et al.. 2015;
Rai et al.. 2010; Honour et al.. 2009). Jaconis et al. (2017) examined the effect of PM on
stigmatic clogging in Cichorium intybus on road sides in Cincinnati, OH. The authors
reported no relationship between PM levels and pollen germination across all of the roads
studied.
15.4. Effects of Particulate Matter on the Soil Environment
As described in the 2009 PM ISA, the soil is one of the most dynamic environments of
biological interaction in nature (U.S. EPA. 2009a). The upper soil layers where deposited
particles accumulate are typically active sites of litter decomposition and plant root
uptake. Processes within the rhizosphere, the soil around plant roots that mineral nutrients
must pass through, may be affected by PM components. Soil-associated microbial
communities of bacteria and fiingae such as actinomycetes and basidiomycetes,
respectively, break down organic matter for nutrient cycling and make elements available
for plant uptake. Soil mycorrhiza, (fungi that colonize plant roots) form a symbiosis to
provide nutrients in exchange for carbon from the plant. The effects of PM-associated
organics and metals on the soil environment is largely dictated by bioavailability to soil
microflora and plant roots.
15.4.1. Bioavailability in Soils
Soils are heterogeneous and the effects of PM deposition and subsequent bioavailability
of PM components depend on soil characteristics (e.g., pH, organic content).
PM-associated organics partition between the soil and the atmosphere based on soil
properties, such as organic matter content, moisture, porosity, texture, and structure, as
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well as the physiochemical properties of the pollutant, including vapor pressure and water
solubility (U.S. EPA. 2009a'). These compounds may be taken up by roots or be
associated with organic matter (Fismcs et al.. 2002). For heavy metals, accumulation in
soil is influenced by a variety of soil characteristics, including pH, Fe and aluminum
oxide content, amount of clay and organic material, and cation exchange capacity [CEC;
(Hernandez et al.. 2003)1. Bioavailability depends on metal speciation, soil pH, and
degree of binding to dissolved organic matter [DOM; (Sauve. 2001)1.
In the 2009 PM ISA, several models, isotopic studies, and sequential extraction methods
for determining metal bioavailability in soils were reported (Feng et al.. 2005; Collins et
al.. 2003; Shan et al.. 2003). However, actual measurement of biological effects is
generally regarded as the preferred criterion to assess bioavailability (Almas et al.. 2004).
Since the 2009 PM ISA, Biomet biosensors have been used to assess bioavailability of
metals in contaminated topsoils and subsoils (Almendras et al.. 2009). Biomet biosensors
use genetically engineered bacteria that give off light when exposed to heavy metals that
are bioavailable. Almendras et al. (2009) also used sequential extraction to determine
solubility. At three sites near a smelter in Chile, they found that Cu and Zn had the
greatest solubility, arsenic (As) had less solubility, and Pb and Fe were the least soluble
of those heavy metals. Further, as was determined to be the most bioavailable, while Cu
and Zn were less bioavailable. The authors were unable to determine bioavailability of Pb
using the same technique.
The mobility of heavy metals, an indicator of potential bioavailability, was determined in
surface soil using sequential extraction (Svendsen et al.. 2011). In five sites at decreasing
distances from a smelter in Norway, Cd was determined to be most mobile and Cu the
least mobile (Svendsen et al.. 2011). Zn and Pb were moderately mobile. These results
were in contrast to the results of (Almendras et al.. 2009).
15.4.2. Soil Nutrient Cycling
In previous PM reviews, toxicity of metals, especially Zn, Cd, and Cu, to soil microflora
was found to reduce decomposition processes in soils and interfere with nutrient cycling
(U.S. EPA. 2009a). In previous PM reviews, changes to microbial enzymatic activity, soil
basal respiration rate, and soil microbial biomass were all associated with increased metal
content in soils. Studies reviewed in the 2009 PM ISA report increased leaf litter
accumulation near point sources.
Since the 2009 PM ISA, many studies showing the deleterious effects of PM on nutrient
cycling have been published. The effect of PM on the microbial coefficient of soils
contaminated by atmospheric emissions from an active mining and smelting complex was
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determined in one study. The microbial coefficient, which is the ratio of microbial
biomass carbon to total organic carbon, was negatively correlated with metal
concentrations (Shukurov et al.. 2014). The microbial coefficient gives a measure of
substrate bioavailability.
A recent study demonstrated the role of atmospheric deposition of particulate heavy
metals in inhibiting C cycling. In this study, a cropping system was exposed to
atmospheric deposition while a similar cropping system was designed to exclude it by
using a greenhouse. Total organic carbon and water soluble organic carbon increased in
the exposed plots and decreased in the greenhouse plots; slow decomposition in soils
polluted with heavy metals could have caused the increased levels of total organic carbon
and water soluble organic carbon observed in the open plots (Pandev and Pandev. 2009).
New studies show decreases in enzyme activity and varying sensitivity of enzymes to
metal pollution. Soil dehydrogenase activity was diminished in polluted soils compared
to control soils (Boiarczuk and Kieliszewska-Rokicka. 2010). High soil dehydrogenase
activity indicates a living microbial community; additionally, dehydrogenase activity
increases with microbial biomass C, organic matter levels, and basal respiration
(Boiarczuk and Kieliszewska-Rokicka. 2010). In another study, enzyme activities of
alkaline phosphatase and fluorescein diacetate were greater in greenhouse treatments than
in open treatments exposed to deposition over the study period, showing that the potential
for hydrolysis was elevated in the greenhouse treatments (Pandev and Pandev. 2009). As
a result, organic C was mineralized to carbon dioxide and released from the soil more
quickly than it was in the treatments exposed to deposition. Qu et al. (2011) found that
enzyme activity was higher at sites farther from an active mining operation than those
close to it. However, urease activity was more affected by heavy metals than
dehydrogenase and phosphatase activity. Dehydrogenase contributes to many oxidative
activities, which degrade soil organic matter (Qu etal.. 2011). Phosphatase conducts
hydrolysis to convert organic phosphorus compounds into inorganic phosphorus
compounds, and urease converts urea into carbon dioxide and ammonia through
hydrolysis (Ou etal.. 2011).
A new study compared soil functional activity measured by the bait-lamina assay and by
the Biolog assay. Feeding activity was increased at the less polluted site compared to the
more polluted site (Boshoff et al.. 2014). Substrate utilization rate and richness of used
substrates were decreased in the more polluted site compared with the less polluted site
after 2 and 6 days. On the other hand, diversity of used substrates at all plots resembled
one another by the end of the experiment. Feeding activity, substrate utilization rate,
richness of used substrates, and diversity of used substrates all decreased as As, Cu, and
Pb concentrations increased. The Biolog assay was a better indicator of soil functional
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activity than the bait-lamina assay (Boshoff ct al.. 2014). The bait-lamina assay indicates
the feeding activity of soil fauna; the Biolog assay measures the metabolic use pattern of
the culturable microbial community.
The effects of PM on microbial activity and biomass have been widely studied in the
literature since the release of the 2009 PM ISA. Additions of Cr and Zn to soils have been
found to reduce microbial respiration (Akcrblom et al.. 2007). The highest concentrations
of Pb and molybdenum (Mo) decreased respiration, as did higher levels of Ni and Cd. On
the other hand, low levels of Ni and Cd increased microbial respiration. In this study,
metals were added in low, medium, and high doses to soil from the humus layer of a
Swedish forest (Akcrblom et al.. 2007). Similarly, in another study, microbial activity
(for all metals except Cu) decreased with increasing metal concentrations, and a negative
correlation was also determined between metal concentrations and microbial biomass
(Anderson et al.. 2009a). Notably, microbial biomass levels were all similarly reduced
among contaminated sites compared to the control site, even though metal concentrations
among the contaminated sites varied. At the same time, microbial activity and biomass
increased with increasing concentrations of physicochemical variables (such as nitrate,
magnesium [Mg], and potassium [K]). Microbial activity, indicated by the amount of
carbon substrates used, was measured using the Biolog assay (Anderson et al.. 2009a).
Using similar methods, (Anderson et al.. 2009b) found that microbial activity and
biomass in polluted sites were reduced compared to activity and biomass in the control
site, prior to the experiment, which consisted of adding metals to soils from the sites.
Microbial activity and biomass were lower at all sites after metal addition (Anderson et
al.. 2009b). In Azarbad et al. (2013). basal respiration and substrate-induced respiration
increased with organic matter and pH and decreased as the toxicity index (an indicator of
heavy metal contamination) increased. The toxicity index was more important in
explaining effects of heavy metal contamination on substrate-induced respiration than
basal respiration. In the same study, microbial biomass, represented by total phospholipid
fatty acids (PLFAs), increased as toxicity index decreased (Azarbad et al.. 2013). Chodak
et al. (2013) found that microbial biomass and basal respiration were affected mainly by
environmental variables, such as total nitrogen and organic carbon, and heavy metals
impacted them much less. In another study, microbial biomass C and substrate-induced
respiration were lower in open treatments exposed to atmospheric deposition than in
greenhouse treatments as time progressed (Pandev and Pandev. 2009).
15.4.3. Soil Community Effects
At the time of the 2009 PM ISA, toxicity of heavy metals to soil-associated microbes
were well characterized in laboratory studies, but less was known about the relative
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sensitivity of fungi, bacteria, and actinomycetes and community-level changes. New
techniques described in the 2009 PM ISA, such as the use of phospholipid and
nucleic-acid biomarkers (Jovnt et al.. 2006). have enabled researchers to better
characterize microbial community composition and biodiversity. In a recent study,
tolerance of microbial communities to heavy metal PM has been shown to vary with
ecosystem type. Culturable bacterial communities in contaminated meadow soils
experienced greater tolerance to pollution, but the culturable communities in
contaminated forest soils did not (Stefanow icz et al.. 2009). The authors suggested that
similarities between the degree of heavy metal contamination in control and polluted
forest soils, caused by increased pH in polluted soils, may have contributed to the lack of
tolerance of bacterial communities in forest soils because pH influences bioavailability of
metals. Additionally, the absence of increased tolerance in forest soils might have been
caused by a lack of variation in metal sensitivity of bacterial species.
The impact of metal-associated PM on the abundance of microbes has been studied in
recently published papers. Total numbers of silver birch (Betulapendula) seedling root
tips colonized by mycorrhizae decreased in seedlings grown in polluted soils relative to
those in control soil (Boiarczuk and Kieliszewska-Rokicka. 2010). In another study,
culturable bacteria were more abundant at the sites farther from an active mining
operation than at the sites close to it (Qu etal.. 2011). Lenart-Boron and Wolnv-Koladka
(2015) found no correlation between metal concentrations and abundance of microbial
groups (mesophilic bacteria, fungi, actinomycetes, and Azotobacter spp.). The effects of
PM from heavy metals on the metabolic quotient of soils that were contaminated by an
active mining and smelting complex in Uzbekistan have been studied recently in the
literature. Metabolic quotient, an indicator of the influence of environmental variables on
microbial communities, increased with increasing heavy metal concentrations (Shukurov
etal.. 2014).
Recently published studies show that the response of soil-associated microbial
community structure and function to heavy metal PM varies. (Akcrblom et al.. 2007)
found that most metals altered community structure in a similar way. Phospholipid fatty
acids (PLFAs) can indicate microbial community structure, and different PLFAs are
associated with different groups of microbes [e.g., fungi, bacteria; (Akcrblom et al..
2007)1. Relative abundance of PLFAs associated with Gram-positive bacteria had a
positive relationship with metal concentrations, but relative abundance of PLFAs
associated with Gram-negative bacteria had a negative relationship with metal
concentrations. The relative abundance of actinomycetes PLFAs had a positive
relationship with metal concentrations. The reaction of the relative abundance of certain
PLFAs found in Gram-negative bacteria varied depending on exposure to particular
metals, and the reaction differed most between Cr and Cd. Relative abundance of a fungal
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PLFA increased with increasing Cr and Zn concentrations (Akcrblom et al.. 2007). Likar
and Regvar (2009) determined that ascomycetes fungi were more common in polluted
plots, and ectomycorrhizal ascomycetes mainly comprised one taxonomic group. On the
other hand, basidiomycetes fungi were more common in control plots and in plots that
were less contaminated, and ectomycorrhizal basidiomycetes were more diverse than
ectomycorrhizal ascomycetes, comprising several taxonomic groups. Dark septate
endophytes in the ascomycete group were common in polluted plots and comprised
several taxonomic groups (Likar and Regvar. 2009). The increased number of DNA
sequences from a particular dark septate endophyte genus (Phialophora) in plots that
contained greater concentrations of Cd and Pb suggests that this genus could help goat
willow (Salix caprea) take up more nutrients under metal stress (Likar and Regvar.
2009); however, the authors did not actually measure nutrient uptake of plants in this
study. Anderson et al. (2009b) found that before the experiment, which consisted of
adding metals to soils with varying levels of initial contamination by an abandoned
smelter in Anaconda, MT, the structure and function of the microbial community from
each site differed from one another, and after the experiment, the communities remained
structurally and functionally different.
In another study, diversity and evenness of soil fungi communities in forest litter and
those of arbuscular mycorrhizal fungi communities in forest litter were lower in a site
polluted by emissions from an active copper smelter than they were in the reference site;
community structure of soil fungi in the polluted site was not similar to that in the control
site (Mikrvukov et al.. 2015). Spatial autocorrelations existed in both the control and
polluted sites for soil fungi communities, but the spatial structure of the communities in
the two sites did not resemble one another. Arbuscular mycorrhizal fungi (AMF)
community structures at control and polluted sites did resemble one another. Spatial
autocorrelation only existed in the polluted sites for arbuscular mycorrhizal fungi
communities (Mikrvukov et al.. 2015). Spatial heterogeneity tended to be greater in
contaminated areas than in uncontaminated areas (Mikrvukov et al.. 2015). Ou et al.
(2011) found that microbial community diversity increased as distance from an active
mining operation increased. In a different study, ecophysiological index values at two
long-term contaminated sites were substantially high for oligotrophs and copiotrophs
(Margesin et al.. 2011). Both sites contained bacterial communities with high Shannon
diversity index values (a common index used to characterize species diversity), and the
metabolically active groups at both sites were Proteobacteria and Actinobacteria
(Margesin et al.. 2011).
Recent studies have shown the effects of environmental variables in comparison to the
effects of heavy metal associated PM. In a study by Anderson et al. (2009a). microbial
species richness was influenced by physicochemical characteristics (Mg, CEC, Ca,
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ammonium [NH/]), but was not affected at all by metal concentrations; however,
community structure in polluted sites was altered by metal concentrations compared to
control sites. In a different study, toxicity index, which represents the degree of heavy
metal pollution, had a positive impact on functional diversity of fungi, and nutrient status
had a negative impact on functional diversity of fungi (Azarbad et al.. 2013). On the other
hand, toxicity index did not influence functional diversity of bacteria, but functional
diversity increased as acidity decreased. PLFAs that represent Gram-positive bacteria
increased with toxicity index, while the PLFA associated with fungi decreased as toxicity
index increased. The level of organic matter and pH were factors that influenced PLFAs
as well as degree of metal contamination (Azarbad et al.. 2013). Chodak et al. (2013)
determined that pH had the largest impact on bacterial community diversity, but heavy
metals also modulated diversity. Bacterial community structure was mainly affected by
pH. Although pH was the main factor controlling microbial community structure, heavy
metals also influenced microbial community structure (Chodak et al.. 2013).
In addition to soil microbial communities, the effects of metal-associated PM on other
soil fauna have been recently published. Shukurov et al. (2014) found that abundance of
nematodes was greater at the plots further from the emission source; similarly, the
abundance of nematodes, number of bacterivorous nematodes, and number of
omnivorous predator nematodes were negatively correlated with heavy metal
concentrations.
The harmful effects of organic PM on soil microbes have been published in recent
studies. AMF infectivity in the top soils was reduced compared to control soils when
exposed to phenanthrene for 2 weeks (Desalme et al.. 2012). In this study, phenanthrene
was pumped into an exposure chamber containing soils to simulate atmospheric
exposure. However, in another study using a similar experimental setup, mycorrhizal
symbiosis in red clover was similar in control and polluted chambers, and mycorrhizal
symbiosis in both treatments remained efficient (Desalme et al.. 201 la). In the same
study, Rhizobium nodule (bacteria) symbiosis in clover was lower in the upper layer of
polluted soil than in the upper layer of control soil (Desalme et al.. 201 la).
15.4.4. Soil Microbe Interactions with Plant Uptake of Particulate Matter
In the 2009 PM ISA, the role mycorrhiza in modulating toxicity of deposited metals was
reported. This role includes improving nutrient uptake and decreasing metal uptake
(Vogelmikus et al.. 2006; Nogueira et al.. 2004; Berthelsen et al.. 1995) in the plant by
acting as a sink for metals (Carvalho et al.. 2006; Berthelsen et al.. 1995). often
preventing the metals in the roots from allocating to shoots (Spares and Siqueira. 2008;
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Zhang et al.. 2005; Kaldorf et al.. 1999). In contrast, other studies reviewed in the
2009 PM ISA showed that mycorrhizae may facilitate accumulation of metals in plants
and enhance the translocation of metals from the root to the shoot (Zimmcr et al.. 2009;
Citterio et al.. 2005; Vogel-Mikus et al.. 2005). There was limited evidence for bacteria
and mycorrhiza working together to improve plant tolerance to metals (Vivas et al.. 2006;
Vivas et al.. 2003).
Since the 2009 PM ISA, additional studies show mycorrhizae modulates metal uptake in
plants. Seedlings grown in polluted soil with mycorrhizae preferentially sequestered
heavy metals (Cu and Pb) to the roots instead of the shoots (Boiarczuk and Kieliszewska-
Rokicka. 2010). The ratio of mycorrhizal fine roots to nonmycorrhizal fine roots was
lower for silver birch seedlings grown in polluted soils (mixture and fully polluted soil
near a copper foundry in Poland) than for tree seedlings grown in control soil; however,
the mycorrhizae were still able to affect metal uptake by the plants.
Nonmycorrhizal microbes (e.g., bacteria, fungal hyphae, spores from saprobes) may also
decrease metal uptake by plants. In a greenhouse experiment using soils collected at
varying distances from abandoned smelters in the U.S., wavy hairgrass (Deschampsia
flexuosa) plants grown in soils with high amounts of contamination and treated with
nonmycorrhizal microbial wash accumulated less Zn in the shoots compared to plants
without the nonmycorrhizal microbial wash (Glassman and Casper. 2012). In contrast,
although there was an AMF treatment in this experiment, AMF did not impact the
amount of Zn in the shoots of plants.
AMF and nonmycorrhizal microbes such as plant growth promoting bacteria can
contribute to phytoremediation of polluted soils; much new research on these topics has
been published since the release of the previous ISA (Cabral et al.. 2015; Meier etal..
2012; Gamalero et al.. 2009).
15.4.5. Effects of Particulate Matter on Physical Properties of Soils
PM can affect physical properties of soils, such as bulk density, porosity, and water
holding capacity. Changes in these properties can decrease plant growth and yield. This
topic was not discussed in the previous PM ISA. However, new studies on the effects of
PM on physical properties of soils have been published since the release of the previous
ISA. Pandev and Pandev (2009) found that bulk density was elevated during the study
period in an experimental cropping system open to deposition. Porosity and water
holding capacity in the open system were diminished over the same time period. The
opposite was the case for the experimental cropping system closed to deposition. The
authors suggested that the influx of deposition of particulate matter during the study
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period caused a rise in bulk density, and the elevated bulk density caused the decrease in
porosity and water holding capacity (Pandev and Pandev. 2009).
15.5. Effects of Particulate Matter on Fauna
Toxicity of PM to fauna varies depending upon the PM composition, concentration,
bioaccessibility and bioavailability, biological species sensitivity, and organism lifestage.
The bioavailability or bioaccessibility of PM-associated trace metals and organics is
dependent upon the physical, chemical, and biological conditions under which an
organism is exposed at a particular geographic location. Organisms may have
detoxification mechanisms that increase tolerance to stressors. Pathways of PM exposure
to aquatic and terrestrial organisms include ingestion, absorption, and trophic transfer.
Several types of studies have been used to assess the effects of PM on terrestrial and
aquatic organisms, including laboratory bioassays and field studies.
15.5.1. Laboratory Bioassays
The use of ecotoxicity assays to screen and assess PM effects on biota has expanded
considerably since the 2009 PM ISA. All but one of the new studies used PM from
outside of the U.S., so the pollutant mix may not be representative of exposures occurring
in the U.S. These studies are reported below. However, several caveats must be noted in
correlating effects in these studies to natural environments. Extracts of PM from air filters
may not be representative of exposure routes and conditions in aquatic and terrestrial
habitats (Kovats et al.. 2012). Bioavailability of PM in the natural environment is affected
by many factors not represented in bioassay techniques. Furthermore, because these
studies include a mix of PM constituents, it may be difficult to identify the active
component(s). Results from toxicity assays using PM extracts (such as dichloromethane,
methanol, or water extracts) should be interpreted with caution because additional
compounds might remain unextracted, underestimating or overestimating the whole PM
toxicity to aquatic organisms (Mesquita et al.. 2014a; Kovats et al.. 2012).
Recently the Vibrio fischeri bioluminescence inhibition test (Microtox®) has been applied
as an ecotoxicological screening tool for PM (Roig etal.. 2013; Kovats et al.. 2012). In
this bioassay, the rate of inhibition of light emission from the bacteria is measured after
exposure to a pollutant. In aqueous PM extracts from air filters from urban, rural, and
industrial areas of Catalonia, Spain, no correlation was observed between inhibition of
light emission in the in vitro test and PMio concentration (Roig et al.. 2013). Metals
(except Cr) and PCDD/F congeners in the samples were significantly correlated to the
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bioassay results, indicating that PM composition, rather than concentration, explains the
experimental observations.
Toxicity of PM extracts to aquatic biota have been assessed in several studies published
since the 2009 PM ISA. Acute (24-hour) toxicity of ambient PM with a nominal mean
aerodynamic diameter less than or equal to 2.5 |im (PM2 5) extracted from air filters in
Atlanta, GA to the freshwater rotifer Brachionus calyciflorus was assessed by (Vermaet
al.. 2013). The toxicity of methanol extracts was much higher than that of water extracts,
and there were no conclusive associations of rotifer toxicity with PM chemical
composition. In zebrafish (Danio rerio) embryos exposed to organic extracts of PM
samples from coal-burning particles in a semirural area in Italy, cytochrome P4501A
gene expression was significantly induced (Olivarcs et al.. 2013). In another zebrafish
study, urban PM had adverse developmental effects (spinal deformations, malformation
of swim bladder, etc.) on embryos (Mcsquita et al.. 2014b). These observations correlated
with PAH content of the PM samples. In a comparison of PM in rural and urban air
particles collected in Spain, induction of AhR signaling pathway correlated with PAH
concentrations in all locations (Mesquita et al.. 2015). The greatest dioxin-like activity
and embryotoxicity was observed in the finest PM fraction (<0.5 |im). During the winter,
maximal assay activity occurred in the rural samples due to biomass-burning emissions.
A series of in vivo studies on the effects of PM2 5 exposure on the nematode
Caenorhabditis elegans indicated a suite of responses in this organism. Zhao et al.
(2014b) observed altered development, lifespan, reproduction, locomotion, and
defecation behavior associated with long-term exposure to PM2 5 collected from a
traffic-dense area in Beijing, China. PM components included S, Cd, Pb, Zn, Cu, and
PAHs. The insulin signaling pathway was identified as a possible molecular target of the
traffic-related PM2 5 in nematodes (Yang etal.. 2016; Yang et al.. 2015). In nematodes
exposed to PM2 5 in coal ash, altered functioning of neurons that control defecation
behavior and gene expression patterns required for control of oxidative stress were
reported, as were effects on development, reproduction, locomotion behavior, lifespan,
and susceptibility to metal toxicity (Wu et al.. 2017; Sun et al.. 2016; Sun et al.. 2015b).
Elemental analysis of the PM from coal combustion indicated Fe, Zn, and Pb were the
most common metals followed by As, Cd, Cr, Cu, and Ni. Fluoranthene, pyrene, and
12 additional PAHs were detected in the PM2 5 in the coal ash.
Earthworms often constitute a large percentage of soil animal biomass and are considered
to be relatively sensitive bioindicators of soil contamination by metals and organics.
Several studies in the 2009 PM ISA reported uptake of trace metals and organics by
earthworms (Hobbelen et al.. 2006; Parrish et al.. 2006). while fewer studies reported
responses due to atmospheric deposition of PM (Massicotte et al.. 2003). Earthworms can
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also modify PM bioavailability and bioaccessibility in soils through bioturbation. For
example, the presence of earthworms has been shown to significantly increase
soil-to-plant transfer of metals via modifications to soil and increased bioaccessibility of
metals to roots (Lcvcquc ct al.. 2014).
In a recent study, earthworms (Eisenia andrei), directly exposed to powder from PMio
quartz filters (24.9 mg/g PMio and 14 jxg/g PAHs) in artificial soil for a range of final
concentrations from 15 to 30 jxg/g PMio, showed genotoxic effects (Vcrnile et al.. 2013).
DNA damage measured by comet assay was observed starting at 22.5 jxg/g PMio
(0.012 jxg/g PAHs) in samples collected from an urban site in Italy.
15.5.2. Wildlife as Biomonitors of Particulate Matter
The 2009 PM ISA reviewed several studies in which resident biota was used to
biomonitor urban air pollution including PM. For example, snails (Helix spp.) accumulate
trace metals and agrochemicals and can be used as effective biomonitors for urban air
pollution (Regoli et al.. 2006; Viard et al.. 2004; Beebv and Richmond. 2002). Biological
effects that have been demonstrated include oxidative stress, growth inhibition,
impairment of reproduction, and induction of metallothioneins, which are involved in
metal detoxification (Regoli et al.. 2006; Gomot-De Vaufleurv and Kerhoas. 2000).
Biomonitoring of aquatic species was also reported. For example, Coelho et al. (2006)
investigated Hg concentrations in Scrobiculariaplana, a long-lived, deposit-feeding
bivalve in southern Europe. The use of sentinel species to detect the effects of complex
mixtures of air pollutants is of particular value because the chemical constituents are
difficult to characterize, exhibit varying bioavailability, and are subject to various
synergistic effects.
New biomonitoring studies of PM effects on biota include vertebrate and invertebrate
organisms. Use of the land snail Helix aspersa as a biomonitor for PAHs associated with
atmospheric particles in Porto Alegre, Brazil indicated, in general, that higher
genotoxicity in this indicator organism was associated with PM size fractions 
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in excreta were always higher in the wild urban pigeon population, while methemoglobin
levels in blood showed significant differences from season to season, but values in the
indoor birds were significantly lower only in summer and autumn (Sicolo et al.. 2009).
The authors chose this model because pigeons have a small habitat and high metabolic
rate, and the main route of exposure to urban air is through inhalation and intake of
contaminated food particles. Assay results were compared with measurements of
contaminants in the urban air (carbon monoxide [CO], PMio, NO2, O3, SO2, benzene
[CeHe], PAHs in PM2 5), and a positive correlation with PAH and protoporphyrin was
observed along with a positive correlation between O3 and DNA damage index (Sicolo et
al.. 2010; Sicolo et al.. 2009). Feathers, lung, liver, and kidney tissues were sampled from
pigeons in two urban areas in the U.S. (Glendora, CA and Midland, TX) to evaluate Hg
exposure (Cizdziel et al.. 2013); however, no significant differences were observed
between the two locations. In another study in two atmospherically polluted sites in North
Texas, Hg was found in the blood and feathers of eastern bluebird (Sialis sialis), Carolina
wren (Thryothorus ludovicianus), wood duck (Aix sponsa), great egret (Ardea alba), and
great blue heron (Ardea herodias) (Schulw itz et al.. 2015). Levels reported in these birds
may be high enough to affect fitness (Schulwitz et al.. 2015).
Several species have been evaluated as indicators of metal pollution from a long-term
monitoring site near a former Cu-Ni smelter in Haijavalta, Finland. Metal concentrations
in bird feces were quantified at the site to measure exposure via atmospheric deposition;
however, measured decreases in the excrement did not directly reflect emission patterns
(Bcrglund et al.. 2015). The authors suggest this is due to uptake of metals from food
items in contact with contaminated soils. Plasma carotenoid levels measured in the birds
on the site were not directly correlated to metal pollution or nesting survival (Ecva et al..
2012). Leg deformities of oribatid mites were also evaluated as a biomonitor but
determined to be an unreliable indicator of heavy metals in soil at the site (Ecva and
Penttinen. 2009).
15.5.3. Biomagnification
Biomagnification of PM-associated metals and organics was reviewed in the
2009 PM ISA and are summarized here. The basic understanding of the process of
biomagnification has not appreciably changed since the previous assessment. As reported
in the 2009 PM ISA, biomagnification is the progressive accumulation of chemicals with
increasing trophic level (Leblanc. 1995). At the time of the 2009 PM ISA,
biomagnification had been demonstrated for a few trace metals in terrestrial and aquatic
systems (U.S. EPA. 2009a). Plant uptake is often the first step for a metals to enter higher
levels of the food web. Consumers of vegetation may often receive heavy loading of
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metals from their diets. Metals may also bioaccumulate in some species, and tissue
concentrations are magnified at the higher trophic levels. Organic Hg is the most likely
metal to biomagnify, in part because organisms can efficiently assimilate methylmercury
and are slow to eliminate it lYCroteau et al.. 2005; Reinfelder et al.. 1998); Appendix 121.
There is also evidence that the trace metals Cd, Pb, Zn, Cu, and selenium (Se)
biomagnify.
Biomagnification of organics has been extensively documented in aquatic and terrestrial
ecosystems, and these compounds are detected in biota at remote locations due to
long-range atmospheric transport processes (U.S. EPA. 2009a'). Organics, such as PAHs,
can be transferred to higher trophic levels, including fish, and this transfer can be
mediated by aquatic invertebrates in the fish diet (U.S. EPA. 2009a'). In high mountain
lakes, atmospheric inputs dominate as sources of contamination. In addition, such lakes
tend to have relatively simple food webs. In a study reviewed in the 2009 PM ISA,
(Vives et al.. 2005) investigated PAH content of brown trout (Salmo trutta) and their food
items. Total PAH concentrations tended to be highest in organisms that occupy littoral
habitats, and lowest in pelagic organisms.
The study of trophic transfer and biomagnification is limited by the difficulty in
discriminating food webs and the uncertainty associated with assigning trophic position
to individual species (Croteau et al.. 2005). Use of stable isotopes can help establish
linkages. However, it is difficult to determine the extent to which biomagnification
occurs in a given ecosystem without thoroughly investigating physiological biodynamics,
habitat, food web structure, and trophic position of relevant species. Thus, developing an
understanding of ecosystem complexity is necessary to determine which species are at
greatest risk from toxic metal exposure (Croteau et al.. 2005).
15.6. Effects of Particulate Matter on Ecological Communities and
Ecosystems
Evidence for PM effects on higher levels of biological organization is primarily from
field studies conducted in proximity to point sources, such as smelters, mining, and other
industrial sources. These areas where PM deposition decreases from the source have been
assessed through ecological gradient studies. Several long-term monitoring studies have
been conducted on ecosystem responses to atmospheric deposition of metals, including
the Haijavalta smelter in Finland where metal-rich dust emissions have decreased by 99%
during the 23-year study period (Eeva and Lehikoinen. 2015). Additional evidence for
PM effects on ecological communities include studies from urban areas and model
microecosystems.
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15.6.1.
Gradient Effects near Smelters
The 2009 PM ISA reviewed multiple studies conducted near a Cu-Ni smelter in
Haijavalta, Finland. These studies documented effects on multiple species operating at
different trophic levels (Kiikkila. 2003; Helmisaari et al.. 1999; Malkonen et al.. 1999;
Pennanen et al.. 1996). The species composition of vegetation, insects, birds, and soil
microbiota changed, and tree growth was reduced 4 km from the smelter. Within 1 km,
very few organisms were documented (Kiikkila. 2003).
Since the 2009 PM ISA, additional studies have been published from the long-term
monitoring of the ecological communities near the Cu-Ni smelter site in Haijavalta,
Finland. Several studies in passerine birds (pied flycatcher [Ficedula hypoleuca\ and
great tit [Parus major]) show some biological recovery and a reduction in metal
concentration with decreased atmospheric emissions over time, even though metal levels
in tissues remain high in birds near the smelter, likely due to contamination of soils and
food items (Bcrglund et al.. 2012; Berglund et al.. 2011). Similarly, tissue levels in pied
flycatchers remained high near a smelter in Northern Sweden even after a 93 to 99%
reduction in metal emissions (Berglund and Nvholm. 2011). With decreased metal
pollution at Haijavalta, breeding parameters in pied flycatchers and great tits have
improved since the 1990s. However, clutch size and fledgling number remain below
those of birds from a reference area (Eeva and Lehikoinen. 2015; Eeva et al.. 2009).
Interruption of egg laying was assessed in passerine birds at the Haijavalta Smelter site
and was attributed to both cold weather and pollution (Eeva and Lehikoinen. 2010).
These laying gaps occurred with greater frequency, and most commonly at the beginning
of the laying sequence, close to the smelter. The authors suggested that heavy metals at
the site interfere with Ca availability and metabolism during the breeding season when
extra Ca is needed for egg shell formation and growth.
Eeva et al. (2010) sampled land snail shells from pied flycatcher nests along the
Haijavalta smelter pollution gradient to assess effects of air pollution on shell mass,
abundance, and diversity of land snail communities. The snails represent an important
source of Ca for the birds, especially for egg shell formation and development of
nestlings. Shell size decreased near the smelter, indicating snail growth was also
impacted at the most highly polluted areas. The highest diversity, largest size, and
greatest abundance of snail populations were observed in the moderately polluted areas
compared to snails in the vicinity of the smelter and in remote unpolluted areas.
Recent literature continues to show a gradient of response in vegetation to trace metal
deposition from smelting operations including evidence from the U.S. The forest
surrounding two historical Zn smelters in Palmerton, PA has been used as a study site to
assess ecological response and recovery following historic emissions. A portion of this
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forest includes the Appalachian National Scenic Trail, and a barren area of approximately
600 to 800 ha remains where Zn, Mn, and Cd were once emitted (Bever et al.. 2013).
Operation at the smelter ceased in 1980, but effects from the contaminated soils persist.
In a greenhouse study using seedlings of native tree species, as little as 10% of soil
collected in proximity to the smelter site and mixed with reference soil caused reduced
growth or mortality of the tree seedings (Bever et al.. 2013). Red maple (Acer rubrum)
was most sensitive, followed by gray birch (Betulapopulifolia), northern red oak
(Quercus rubra), chestnut oak (Quercus prinus), and eastern white pine (Pinus strobus).
In a field survey relating Zn soil concentrations to forest response, canopy closure and
shrub cover was decreased to half at approximately 2,060 mg/kg Zn, while tree-seedling
density was reduced by 80% at 1,080 mg/kg Zn (Bever etal.. 2011). Vegetative
communities near the Nikel smelter in Russia near the border with Finland and Norway
showed varied response to deposition of heavy metals (Mvking et al.. 2009). Epiphytic
lichens were most affected followed by decreased abundance and species richness of
lichens and bryophytes closer to the smelter. No effects on tree crown condition or
growth were observed.
15.6.2. Urban Environments
Resident biota in urban areas are affected by PM along with other stressors associated
with the built environment such as increased temperature due to heat island effects
(Pickett et al.. 2011). McDonnell et al. (1997) observed differences in urban forest
structure and function compared to rural and suburban areas of New York. Oak forests
near the city were characterized by reduced soil fungal and invertebrate populations,
elevated soil metals (presumably from air pollution), and lower quality of leaf litter. The
removal of PM by urban vegetation to mitigate air pollution has been extensively
reviewed, but few of these studies consider ecological impacts of PM (Escobedo et al..
2011; Pataki etal.. 2011).
15.6.3. Aquatic Ecosystems
In the 2009 PM ISA, PM components in aquatic ecosystems were primarily considered in
the context of aquatic food web transfer of organic contaminants (U.S. EPA. 2009a). In
another study reviewed in the 2009 PM ISA, bioassay procedures with green algae were
used to provide an initial screening of ambient PM toxicity from two urban/industrial and
one rural site near Lake Michigan (Sheeslev et al.. 2004). Results suggested that toxicity
was not related to the total mass of PM in the extract but to the chemical components of
the PM. In a study of the effects of contaminated tunnel wash water runoff in Norway,
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growth of sea trout (Salmo trutta) was significantly reduced in the stream receiving the
wastewater (Meland et al.. 2010). The contaminated wash water had elevated
concentrations of traffic-related contaminants including metals, PAHs, and road salt from
combustion and highway runoff. Fractionation of water samples indicated some of the
metals and PAHs were highly associated with particles. Size ratios of detected PAHs
suggested that sources included tire wear and incomplete combustion.
15.6.4. Experimental Microecosystems
Since the 2009 PM ISA, several studies have employed a microecosystem to assess
effects of air pollutants at the community level of biological organization to see how
microbial communities living on terrestrial mosses (bryophytes) respond to atmospheric
deposition of PM components (metals, PAHs, NO2). In a series of bioindicator studies
conducted in France, species-specific responses to PM components and effects on total
biomass were observed in the associated microbial communities (Meyer et al.. 2010a;
Mever et al.. 2010b). Primary producers, decomposers, and predators responded
differently to the pollutants (Mever et al.. 2010a; Mever et al.. 2010b). Testate amoebae
are especially sensitive to changes in atmospheric pollution in these microsystems and
were proposed as a biomonitoring tool (Mever et al.. 2012). Total testate amoebae
abundance and the abundance of five species of testate amoebae decreased with increased
deposition of the PAH phenanthrene in the experimental microecosystem (Mever et al..
2013).
15.7. Summary of Ecological Effects of Particulate Matter
Since publication of the 2009 PM ISA, new literature builds upon the existing knowledge
of ecological effects associated with PM components, especially metals and organics. In
some instances, new techniques have enabled further characterization of the mechanisms
of PM on soil processes, vegetation, and effects on fauna. New studies provide additional
evidence for community-level responses to PM deposition, especially in soil microbial
communities. However, uncertainties remain due to the difficulty in quantifying
relationships between ambient concentrations of PM and ecosystem response. Overall,
the body of evidence is sufficient to infer a likely causal relationship between
deposition of PM and a variety of effects on individual organisms and ecosystems,
based on information from the previous review and new findings in this review.
In regard to direct effects of PM on radiative flux, a newly available research method
links changes in expression of proteins involved in photosynthesis to changes in radiation
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associated with aerosols and PM. Although this method has not been widely applied, it
may represent an important way to study mechanistic changes to photosynthesis in
response to more diffuse radiation resulting from PM in the air column.
In general, new studies on PM deposition to vegetation support findings in previous PM
reviews on altered photosynthesis, transpiration, and reduced growth. Additional
characterization of PM effects at the leaf surface since the 2009 PM ISA has led to a
greater understanding of PM foliar uptake. Alterations in leaf fatty acid composition are
associated with metals transferred to plant tissues from PM deposition on foliar surfaces
(Schreck et al.. 2013).
Several studies published since the 2009 PM ISA show PM effects on soil physical
properties and nutrient cycling. Previous findings in the PM ISA of changes to microbial
respiration and biomass are further supported by new studies. Microbial community
responses to PM vary in tolerance to heavy metals and organics.
In fauna, results from ecotoxicity assays with PM extracts using bacteria, rotifers,
nematodes, zebrafish, and earthworms support findings in the 2009 PM ISA that toxicity
is not related to the total mass of PM in the extract, but to the chemical components of the
PM. In nematodes exposed to PM from air filters, the insulin-signaling pathway was
identified as a possible molecular target. Use of wildlife as PM biomonitors has been
expanded to new taxa since the last PM review. Several studies in invertebrates and birds
report physiological responses to air pollutants, including PM.
For ecosystem-level effects, a gradient of response with increasing distance from PM
source was reported in the 2009 PM ISA. Newly available studies from long-term
ecological monitoring sites provide limited evidence for recovery in areas such as those
around former smelters due to the continued presence of metals in soils after operations
ceased. A novel experimental microecosystem using microbial communities living in
terrestrial mosses indicates that PM deposition alters responses of primary producers,
decomposers, and predators.
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APPENDIX 16. CASE STUDIES
This appendix includes six case studies which are meant to identify the ecological effects
of nitrogen (N) and sulfur (S) with a specific focus on national parks, other protected
areas, and areas with long-term research. The locations of these case studies were chosen
because they are areas for which a substantial amount of published work on ecological
response to N and/or S deposition is available. The locations include the northeastern
U.S., Adirondack State Park, southeastern Appalachia, Tampa Bay, Rocky Mountain
National Park, and southern California. These case studies identify current acidification
and nutrient status, as well as empirical and modeled critical loads. A recent model of
reactive N deposition to Federal Class I areas suggests that Nr deposition exceeds the
most sensitive critical loads in many of these areas [Figure 16-1. from Lee et al. (2016)1.
The scientific characterizations of ecological responses to N and S in these case studies
set the scientific foundation for further analysis of risk and exposure.
0.00	2.25	4.50	6.75	9.00 [kg N/ha/yr]
Notes: color indicates magnitude of the reactive N exceedance. The size of Class I areas is not represented. Air quality model grid
cells containing Class I areas are shown as colored regardless of the fraction of grid cell area the Class I area covers. Bold line
divides Western and Eastern U.S.
Source: Figure 9 from Lee et al. (2016).
Figure 16-1 CL exceedance in Class I areas.
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16.1.
Northeastern U.S. Case Study: Acadia National Park,
Hubbard Brook Experimental Forest, and Bear Brook
Watershed
16.1.1. Background
This case study is meant to identify effects of nitrogen (N) and sulfur (S) in the
northeastern U.S., with a specific focus on national parks and areas with long-term
research data. This case study identifies current acidification and nutrient status and
empirical and modeled critical loads (CLs). The 2008 NOx-SOx ISA included a case
study of acidification in the Adirondack region of New York [Section 3.2.2.4 of 2008
ISA (U.S. EPA. 2008aYI. This case study is considered as a supplement to that earlier
case study. Further information about the Adirondack region can be found in Appendix 4
and Appendix 5.
16.1.1.1. Description of Case Study Region
Acadia National Park (ACAD) in coastal Maine, the Hubbard Brook Experimental Forest
(HBEF) in the White Mountains of New Hampshire, and the Bear Brook Watershed
(BBW) in southeastern Maine were chosen for this case study to represent northeastern
U.S. ecosystems known historically to be sensitive to acidification and nutrient
enrichment from atmospheric S and N deposition (Figure 16-2. Table 16-1). In this case
study, we highlight multiple decades of research and monitoring at these locations to
provide insights into the ecosystems' responses to S and N deposition and look for any
indications of recovery.
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Northeast Case
Study Region
I Caw Study locations
hm Native American
Reservations
USA City Populations
•	1,000,000 plus
•	500.000 999,999
250,000 - 499,999
•	100,000 - 249,999
•	50,000 99,999
10,001 • 49,999
New*,.-
Hampshire
Acadia National
Park
Oc*sn
Figure 18-2 Locations of northeastern U.S. case study areas and nearby
human population centers.
Table 16-1 Selected characteristics of northeastern case study areas.
Case Study Area
Elevation
Geology
Dominant Vegetation
Focus
ACAD
0 to 466 m
Granitic
Red spruce, mixed forest
National Park
HBEF
183 to 1,002 m
Metamorphic
Northern hardwood
forest, mixed forest,
spruce-fir forest
Experimental forest, site of
diverse biogeochemical research
and monitoring
BBW
172 to 450 m
Metamorphic
Northern hardwood forest
Site of watershed manipulation
experiment
ACAD = Acadia National Park; BBW = Bear Brook Watershed; m = meters; HBEF = Hubbard Brook Experimental Forest.
16.1.1.1.1. Acadia National Park
1	Acadia National Park is located at the interface between northern coniferous forests and
2	temperate deciduous woods and hosts plant species from two distinct regions
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(http://www.acadiacentennial2016.org/visit-acadia/acadias-treasures/'). ACAD
encompasses 45,000 acres on two islands and a mainland peninsula on Maine's coast. It
receives about 140 cm/year of rainfall (Nielsen and Kahl. 2007). Geologically, the park
consists of granite mountains with surrounding sedimentary and metamorphic rock. The
soils have low buffering capacity, and steep slopes at high elevations make soil and
runoff susceptible to acidification. The park's ecosystems include alpine heath, stunted
growth woodlands, jack pine forests that include pitch pine, rocky woodlands of black
spruce and heaths, mature spruce, and fir. Red spruce and sugar maple are the
predominant tree species.
Eighty freshwater plant species have been documented in ACAD with another
12 semiaquatic shorelines species. Seven of the aquatic and semiaquatic plants are listed
or proposed for inclusion on Maine's List of Endangered and Threatened Species.
Another 30 of these species are considered "locally rare"
(https://www.nps.gov/acad/learn/nature/plants.htm').
Hubbard Brook Experimental Forest
Hubbard Brook Experimental Forest is located in the southern portion of the White
Mountain National Forest in central New Hampshire (Figure 16-3). The Atlantic Ocean is
about 116 km to the southeast. Most of the surrounding land is also within the National
Forest. The area is hilly and steep in places with coarse-textured acidic soils. The bedrock
is dominated by metamorphosed igneous and sedimentary rocks. Bedrock is covered by
about 2 m of glacial till.
Northern hardwood forests occupy lower elevation slopes with spruce-fir occurring on
upper elevation slopes. Annual average precipitation is about 140 cm (one-third to
one-quarter of which falls as snow; http://hubbardbrookfoundation.org/wp-
content/uploads/2010/12/long term effects.pdf).
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f - ^, Ml. Cuahman
ak^.977 m
-—b A\ \ - J
Bird Trsneect Ares
mk
I
Mt. w
Klneo *
1015 m \
o-..
.o
/ \
H %Ą>*
Legend
°-. /
° \ 8.	^
to—•" *v/	N
o Rain gage
	Stream
	Gaged Watershed o
	HBEF Boundary
A
3 km
FS = Forest Service; HBEh = Hubbard Brook Experimental Forest; km = kilometer; m = meter; Mt = mount; Mtn = mountain;
Natl = national.
Notes: shown are the network of rain gages, experimental watersheds(1-9), and Mirror Lake.
Source: Campbell et al. (2007).
Figure 16-3 Site map of Hubbard Brook Experimental Forest in the White
Mountains of New Hampshire.
Glacier-deposited materials vary greatly in degree of sorting and grain size, with
thicknesses ranging from zero on ridgetops and in stream valleys to 50 m near Mirror
Lake. The principal soils are acidic (pH about 4.5 or less) and relatively infertile. The
20- to 200-mm thick forest floor allows rapid infiltration of water and insulates against
soil freezing before snow accumulates. Streams vary from small ephemeral channels to as
large as perennial 5th order. Up to about 60 to 80% of storm precipitation flows by
stream. HBEF is entirely forested, predominately with deciduous northern hardwoods
including sugar maple (Acer sctcchanim), beech (Fagus grandifolia), and yellow birch
(.Betula cillegheniensis). High-elevation conifers include red spruce (Picea nibens),
balsam fir (Abies balsamed), and white birch (Betula papyri ferct). Logging operations
ended around 1915 to 1917. The second-growth forest is about 80 to 90% hardwoods and
10 to 20% conifers. HBEF hosts more than 90 species of birds, snowshoe hare, moose,
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fox, black bear, beaver, and white-tailed deer are present
(http://www.hubbardbrook.org/overview/sitedescription.shtml').
HBEF ornithological studies have found that the abundance of birds decreased from more
than 200 individuals per 10 ha in the early 1970s to 70 to 100 per 10 ha during the period
from the early 1990s to the present. This decrease is unexplained (R.T. Holmes
publications—Bird Abundances—http://\\\\\\ .hubbardbrook.org/data/datasct.php'.)id=S 1
and http://hubbardbrookfoundation.org/wp-
content/uploads/2010/12/long term effects.pdf).
Bear Brook Watershed
The BBW hosts a long-term, gaged, forested, first-order paired stream watershed. It is
located about 40 km from the Atlantic Ocean on the southeast slope of Lead Mountain. It
has a total relief of 210 m and a maximum elevation of 450 m. Two nearly perennial, low
dissolved organic carbon (DOC), low acid neutralizing-capacity (ANC) streams (East and
West Bear Brook [EB and WB]) drain 10.3 and 11.0 ha contiguous watersheds,
respectively. BBW is covered predominantly by hardwoods (Fagus grandifolia, Acer
rubrum, Acer saccharum, Betula alleghaniensis, Betulapapyrifera, and Acer
pensylvaticum) and red spruce. Soils are coarse, loamy, mixed, frigid Typic Haplorthods
developed on till averaging 1 m in thickness. Bedrock is mainly quartzites and
metapelites with local granitic intrusions
(https://umaine.edu/bbwm/about-bbwm/site-description-2/').
West Bear Brook has received bimonthly additions of ammonium nitrate since November
1989 (-1,800 eq/ha/yr—a 300% increase over ambient N deposition at the beginning of
the study). East Bear Brook has received no additions and serves as a reference [see
(Norton et al.. 1999)1. BBW has been a research site for three decades with a focus on
increasing decadal-scale understanding about northern forested ecosystems' response to
chemical and physical change, including whole ecosystem acidification, nitrogen
enrichment, and climate change. BBW research has focused on: alterations to nitrogen
dynamics, base cation decline, carbon cycling, phosphorus controls on nitrogen cycling,
response to decreasing ambient sulfate deposition, sustained elevated experimental
sulfate deposition, and changes in forest growth
(https://umaine.edu/bbwm/research/environmental-research/').
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16.1.1.2. Class I Areas
In an effort to preserve pristine atmospheric conditions, the Clean Air Act (42 USC 7470)
authorized Class I areas to protect air quality in national parks over 6,000 acres in size
and national wilderness areas over 5,000 acres.
Class I areas are subject to the "prevention of significant deterioration (PSD)" regulations
under the Clean Air Act. (42 USC 7470). PSD preconstruction permits are required for
new and modified existing air pollution sources, and air regulatory agencies are required
to notify federal land managers (FLMs) of any PSD permit applications for facilities
within 100 km of a Class I area.
ACAD is a PSD Class I Area. HBEF is not in a Class I area, but is near two New
Hampshire Class I areas, the Great Gulf and the Presidential Range—Dry River
Wilderness Areas. BBW is located southwest of the Moosehorn Wilderness Area in
Maine.
16.1.1.3. Regional Land Use and Land Cover
No large population centers are located near HBEF or BBW. Several small towns are
interspersed with ACAD lands, including Bar Harbor (pop. 5,235), Southwest Harbor
(pop. 1,765), and Trenton (pop. 1,481). The populations of these towns increase
substantially with tourists during summer. Bangor (pop. 32,000) lies about 40 miles north
from Mount Desert Island. All three study areas are mostly forested, as is much of the
Northeast region (Table 16-2. Figure 16-4).
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Table 16-2 Land use/land cover for northeastern case study areas.


Area Covered (km2)


ACAD
HBEF
BBW
Developed and barren
83.82
2.96
0.00
Hardwood forest
76.84
159.53
6.64
Conifer forest
592.56
61.88
0.60
Mixed forest and shrubland
300.27
111.49
2.53
Meadow/herbaceous
26.33
0.50
0.00
Wetland
137.05
1.47
0.16
ACAD = Acadia National Park; BBW = Bear Brook Watershed; HBEF = Hubbard Brook Experimental Forest; km = kilometer.
Northeast Case Study
Region: Land Cover
! Cast Study Locations
Native American
Reservations
Developed Land [increasing intensity)
Barren land
I Deciduous forest
H Evergreen Forest
Mixed Forest
	; Shrub/Scrub
Grassland/Herbaceous
¦ Cultivated Crops
m, Water/Wetlands
0	20	40	GO	80
Acadia .National
Park
Atlantic Ocean
Figure 16-4
June 2018
Land cover in the Northeast case study region.
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16.1.1.4. Organization of This Case Study
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Table 16-3 Literature cited by Northeast U.S. case study area.
Variable
Acadia
BBW
HBEF Northeast Regional
Summary of studies from NE case study for acidification
Base cation

Elvir et al. (2006)

BAI

SanClements et al.
(2010)

Mineralization

Fernandez et al.
(2003)

Stream dissolved
S042", pH, ANC, Al+

Norton et al. (2004)

Base cation, Ca, Mg

Norton etal. (2004)
Gbondo-Tuqbawa
and Driscoll (2003),
Gbondo-Tuabawa
and Driscoll (2002)
Summary of studies from NE case study for nutrient status
Aquatic N cycling

Mineau et al. (2014).
Simon et al. (2010)

Soil N cycling

Mineau et al. (2014)
Jefts et al. (2004)

Ca addition


Battles etal. (2014)
N addition

Bethers et al. (2009)
Hunt et al. (2008)

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Because this case study addresses the condition of three discrete locations in the
Northeast (ACAD, HBEF, and BBW), we developed Table 16-3 to summarize the
primary post-2000 research reported in the case study. In addition, we cite relevant
research in other areas of the northeastern region. Appendix 16.1.2 presents information
about N and S deposition in the Northeast. Discussions of critical load or dose-response
research (Appendix 16.1.3) are each organized around the three case study areas and the
Northeast region. Appendix 16.1.4 presents information available on long-term
ecological monitoring, and Appendix 16.1.5 contains information on the current status
and forecasts for recovery in the case study areas. Key research literature since January
2008 is highlighted in Table 16-4.
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Table 16-3 (Continued): Literature cited by northeast U.S. case study area.
Variable
Acadia
BBW
HBEF
Northeast Regional
N retention + multiple
disturbances


Aber et al. (2002)

N retention
climate/freezing


Groffman et al.
(2011). Judd etal.
(2011). Campbell et
al. (2010)

N retention and fire
Nelson etal. (2007)
Campbell et al.
(2004b)



N dep effect on
wetlands
Calhoun etal. (1994)



N loading to coastal
waters
Nielsen and Kahl
(2007)



N addition to lakes
Saros (2014)



Summary of studies from NE case study for terrestrial critical load or dose-response
Dose-response trees
Ellis etal. (2013)
Elvir et al. (2006). Elvir
et al. (2003)

Pardo et al. (2011c).
Thomas et al. (2010).
McNultv et al. (2005)
Does-response herbs



Pardo etal. (2011c)
Hurd etal. (1998)
Dose-response soil

SanClements et al.
(2010) Fernandez et
al. (2003)
Gbondo-Tuabawa
and Driscoll (2002)
Pardo etal. (2011c)
Dose-response
surface water


Gbondo-Tuabawa
and Driscoll (2002)

Modeling climate,
acid-base chemistry,
changes in plants

Phelan etal. (2016)
Phelan etal. (2016)
Campbell et al.
(2009)

Summary of studies from NE case study for aquatic critical load or dose-response
Macroinvertebrate
production



Chadwick and Hurvn
(2005)
Shift from deposition
to climate influence



Mitchell and Likens
(2011)
Flow volume and peak
flows



Kim etal. (2010)
NO3" concentration



Aber et al. (2003)
N limitation



Baron et al. (2011a)
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Table 16-3 (Continued): Literature cited by northeast U.S. case study area.
Variable
Acadia
BBW
HBEF
Northeast Regional
NO3" leaching



Driscoll et al. (2003a)
DOC



SanClements et al.
(2012)
Climate effects on
pools, concentrations,
and fluxes of major
elements



Pourmokhtarian et al.
(2012)
PH



Ouimet et al. (2006),
Dupont et al. (2005),
Ouimet et al. (2001)
ANC



Sutherland et al. (2015),
Nierzwicki-Bauer et al.
(2010), Tominaqa et al.
(2010), Ouimet et al.
(2006), Dupont et al.
(2005), Ouimet et al.
(2001)
Al+ = aluminum; ANC = acid neutralizing capacity; BAI = basal area increment; BBW = Bear Brook Watershed; Ca = calcium;
DOC = dissolved organic carbon; HBEF = Hubbard Brook Experimental Forest; Mg = magnesium; N = nitrogen; NE = northeastern;
N03" = nitrate; S042" = sulfate.
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Table 16-4 Key recent research literature focused on the case study region.
Publication
Focus
Baron et al. (2011a)
Regional patterns in N limitation
Battles et al. (2014)
Ca addition at HBEF
Bethers et al. (2009)
Effects of S and N on sugar maple foliar chemistry
CamDbell et al. (2009)
Influence of climate on biogeochemical cycling
Campbell et al. (2010)
Monitoring of climate conditions at HBEF
Cho et al. (2009)
Response of stream high-flow chemistry to Ca addition at HBEF
Ellis et al. (2013)
Empirical critical loads
Elviret al. (2010)
Tree growth and foliar chemistry at BBW
Fernandez and Norton (2010)
Overview of BBW study
Fernandez et al. (2010)
Responses to experimental acidification at BBW
Groffman et al. (2011)
Soil freezing at HBEF
Hunt et al. (2008)
N and P content of leaves
Judd et al. (2011)
Soil freezing at HBEF
Kim et al. (2010)
Stream flow at BBW
Laudon and Norton (2010)
Episodic stream chemistry at BBW
Lawrence et al. (2012)
Recovery of soil base status
Lona et al. (2009)
Sugar maple decline
Mineau et al. (2014)
Enzyme activity in soil and leaf litter
Mitchell and Likens (2011)
S budget at HBEF
Mitchell et al. (2011)
S mass balances for HBEF and BBW
BBW = Bear Brook Watershed; Ca = calcium; HBEF = Hubbard Brook Experimental Forest; N = nitrogen; P = phosphorus; S = sulfur.
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16.1.2. Deposition
Characteristics of nitrogen and sulfur deposition affecting the ACAD, HBEF, and BBW
in the study area are shown in Figure 16-5 and Figure 16-6. Figure 16-5 A and
Figure 16-6A show 3-year average total deposition of N and S for 2011-2013;
Figure 16-5B shows the partitioning between oxidized and reduced N; Figure 16-6B
shows the 25-year-long time series for wet deposition for NOs . NH4+, SO42 and H+
obtained at the NADP/NTN (National Atmospheric Deposition Program/National Trends
Network) monitoring site at HBEF (NH02). Surrounding areas in Maine and New
Hampshire and inserts showing the coterminous U.S. (CONUS) are shown to place the
depositional environment in context. See Appendix 2 A, Appendix 2.5 and Appendix 2JS
for more information on deposition in the U.S. Other maps showing the contributions of
individual species to dry and/or wet deposition are given in Appendix 2/7.
Data shown in the map Figure 16-5 and Figure 16-6A were obtained from the hybrid
modeling/data fusion product, TDEP (Total Deposition,
http://nadp.sws.uiuc.edu/committees/tdep/tdepmaps/ and described in Appendix 2.7).
However, the time series of wet deposition is taken directly from data on the NADP/NTN
(Figure 16-6B). This was done to track changes in deposition since the passage of the
Clean Air Act Amendment (CAAA) because the Community Multiscale Air Quality
(CMAQ) model simulations used in TDEP extend back only to 2000.
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Bear
Brook
Acadia
N Deposition (kg-N/ha)
*> J» & & •
Hubbard
Brook
O
O
Monitor NH# 02
Monitor ME # 98
Monitor Locations
Northeast Study Area
A
B
ha = hectare; kg = kilogram; N = nitrogen.
Figure 16-5 Total nitrogen deposition (A) and percentage of oxidized nitrogen
deposition (B) for the Northeast case study area estimated by the
National Atmospheric Deposition Program Total Deposition
Science committee.
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600
S00
400
ro
JZ
~o
*1 100
c
o
2 200
o
CL
V
Q
100
Annual Wet Deposition and 3-Year Moving Average at
Site NH02: 1990 - 2014
• 			
• N.

a	.. ~ V _
1988
1992
1996
2000	2004
Year
2008
2012
2016
B
H+ = hydrogen ion; ha = hectare; kg = kilogram; mol = mole; NH4+ = ammonium; N03 = nitrate; S = sulfur; S042 = sulfate; yr = year.
Figure 16-6 Total sulfur deposition (A) for the Northeast case study area
estimated by the National Atmospheric Deposition Program Total
Deposition Science committee. Time series of wet deposition (B)
from the National Atmospheric Deposition Program/National
Trends Network in the Hubbard Brook Experimental Forest, NH.
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Comparison of Figure 16-5 A and Figure 16-6A indicates that the general pattern of
deposition ofN and S is broadly similar; deposition tends to be higher near the coast and
decreases inland. Deposition at ACAD is an exception, with lower values along the coast
than elsewhere. As seen in Figure 16-5A. total deposition of N is relatively uniform
within the three portions of the study area. In addition, the area surrounding the three
sites shows a high degree of regional homogeneity but with higher values in New
Hampshire. Deposition of N is not as high as in many areas of the central U.S. or the rest
of the Northeast. Deposition of S is considerably lower than along the Ohio River Valley.
Figure 16-5B shows that the deposition of nitrogen is estimated to be mostly in oxidized
form throughout the entire domain. Although most of the area is subject to N deposition
in oxidized form with highest percentages surrounding ACAD and BBW, there are areas,
principally in central Maine, where N is deposited mostly in reduced form. In
Figure 16-6B. wet deposition of all species shows that downward trends in NO3 , NH4 .
S042 and H+ are consistently found over the past 25 years, although the rate of decrease
has been variable. In general, wet deposition typically exceeds dry deposition of N and S
in this case study area.
16.1.3. Critical Loads and Other Dose-Response Relationships
16.1.3.1. Terrestrial
This section presents post-2000 ACAD, HBEF, and BBW findings on the dose-response
relationships of N and S to terrestrial ecology, as well as the critical N and S loads for
maintaining ecosystem health. Additional relevant information for the northeastern region
is summarized. The section presents findings on both empirical research and modeling
analyses.
16.1.3.1.1. Empirical Studies
Post-2000 findings from empirical studies of terrestrial dose-response relationships and
critical loads are summarized in this section. Table 16-5 summarizes the body of
empirical and modeling research identified.
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Table 16-5 Terrestrial empirical and modeling research on the response of
nitrogen and sulfur deposition for the northeastern U.S.



Deposition/






Addition



Variable
Species
Response
(kg N/ha/yr)
Years
Site
Reference
Primary
NA
PnET-BGC Model. Predicted
Not specified
1999-2099
HBEF
Campbell et
productivity

increase due to future longer



al. (2009)


growing season




Base cation
NA
PnET-BGC Model. Historical
Not specified
1850-1995
HBEF
Gbondo-
Ca and Mg

forest cutting had little impact



Tuabawa and
depletion

on exchangeable cation soil



Driscoll


pools



(2003)
Base cation
Sugar
American beech and red
N addition: 8.4
1989-2003
BBW
Elvir et al.

maple,
spruce had lower foliar Ca,
(wet + dry)


(2006)

American
Mg, Zn concentrations;
(NH4)2S04




beech,
nutrient imbalance may offset
addition:




red
potential photosynthesis
WB 25.2




spruce
benefits






Sugar maple had higher






photosynthesis rates; no






decrease in Ca, Mg, Zn






concentrations




Tree growth Northern +
	hardwood —
Mortality
ND
8.4 N
¦ (wet + dry)/
+25 (NH4)2S04
1989-2002 BBW
Pardo et al.
(2011c)
Foliar %N
Foliar %Ca
NO3" leaching
Cation loss
Soil C:N
N
mineralization
Nitrification
Soil
respiration
Microbial
biomass
Tree growth Red 0
	spruce —
Mortality
ND
8.4 N	1989-2002 BBW Pardo et al.
¦ (wet + dry)/+25	(2011c)
kg (NH4)2S04
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Table 16-5 (Continued): Terrestrial empirical and modeling research on the
response of nitrogen and sulfur deposition for the
northeastern U.S.
Deposition/
Addition
Variable Species	Response	(kg N/ha/yr) Years	Site Reference
Foliar %N	+
Foliar %Ca
NO3 leaching	+
Cation loss	+
Soil C:N
N	+
mineralization
Nitrification	+
Soil	ND
respiration
Microbial	ND
biomass
Base cation Sugar
maple
Mineralization Not
specified
Little evidence of BC
depletion but confounded by
ice storm litter mineralization
Storm caused increased
litterfall, accelerated
mineralization, and obstructed
temporal trends in soil
chemistry (17 yr)
8.4 N	Not
(wet + dry)/ + 25 specified
kg (NH4)2S04
BBW SanClements
etal. (2010)
Fernandez et
al. (2003)
8.4	1989-1998 BBW
(wet + dry)/+25
(NH4)2S04
Soil base Not	ForSAFE-VEG model.	Not specified Not	BBW and Phelan et al
saturation specified Expected future climate	specified HBEF (2016)
change was simulated to
cause increase, especially at
BBW.
Soil acid-base Not	ForSAFE-VEG model.	Not specified Not	BBW and Phelan et al.
chemistry specified Simulated future climate had	specified HBEF (2016)
a lesser effect at HBEF than
BBW, likely due to the
overwhelming influences of
high S and N deposition.
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Table 16-5 (Continued): Terrestrial empirical and modeling research on the
response of nitrogen and sulfur deposition for the
northeastern U.S.
Variable Species
Response
Deposition/
Addition
(kg N/ha/yr)
Years
Site
Reference
Plant	Not	ForSAFE-VEG model,
communities specified Climate futures predicted by
and N	the IPCC of increased
enrichment	temperature and precipitation
will change plant communities
and N enrichment,
counteracting the acidifying
impacts of S and N deposition
on soil acid-base chemistry.
Not specified Not	BBW and Phelan et al.
specified HBEF (2016)
NO3 leaching NA
PnET-BGC Model. Increase
due to enhanced net
mineralization and nitrification
Not specified 1999-2099 NE
Campbell et
al. (2009)
Mineral	NA	PnET-BGC Model. Slight Not specified 1999-2099 NE	Campbell et
weathering	decrease due to reduced	al. (2009)
simulated soil moisture
(negative effect) and
increased temperature
(positive effect)
NO3"	Not	DayCent-Chem model,
leaching, specified Simulated future high N
GHG	deposition under climate
sequestration,	scenarios had increased
ecosystem C	ecosystem GHG
pools	sequestration. High N
increased stream NO3" fluxes
and ecosystem C pools.
Simulated N 2001-2075 ACAD Hartman et al.
deposition 3.9 to	and	(2014)
9.6 kg/ha/yr for	HBEF
HBEF and 4 to
11.3 kg/ha/yr for
ACADr
BBW = Bear Brook Watershed; BC = base cation; C = carbon; Ca = calcium; ForSAFE-VEG = Soil Acidification in Forest
Ecosystems; ha = hectare; GHG = green house gas; HBEF = Hubbard Brook Experimental Forest; IPCC = Intergovernmental Panel
on Climate Change; kg = kilogram; Mg = magnesium; N = nitrogen; NA = not applicable; ND = no data; NE = northeastern;
(NH4)2S04 = ammonium sulfate; N03" = nitrate; PnET-BGC = Photosynthesis and Evapotranspiration-Biogeochemical; S = sulfur;
WB = West Bear Brook; yr = year; Zn = zinc.
16.1.3.1.1.1. Acadia National Park
1	Ellis et al. (2013) estimated the CL exceedance for nutrient-N deposition to protect the
2	most sensitive ecosystem receptors in 45 national parks, based on CL ranges compiled by
3	(Pardo et al.. 2011c). Ellis et al. (2013) estimated the N CL for ACAD in the range of
4	3-8 kg N/ha/yr to protect hardwood forests and a similar range to protect lichens.
5	Cleavitt et al. (2015) conducted surveys and used FIA assessments of lichen species in
6	the Northeast, with particular focus on the four Class I wilderness areas in the region
7	(Lye Brook, VT; Great Gulf, NH; Presidential Range—Dry River, NH; and Acadia
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National Park, ME). In general, cumulative N and S deposition over the course of the
modeled period (2000-2013) was a more powerful explanatory factor than annual N or S
load for the lichen data. This work determined a critical load of 4.3-5.7 kg total N
(oxidized + reduced N)/ha/yr based on lichen species richness, the abundance of sensitive
species (higher species richness of cyanolichens and fruticose lichens below CL), and
thallus condition. Cumulative S and N deposition were both equally powerful predictors
of thallus condition, but no S critical load was determined (Cleavitt et al.. 2015).
Pardo etal. (2011c) estimated that ambient N deposition in parts of ACAD was higher
than estimated empirical CL values. Thus, these data suggest the possibility of
exceedance of nutrient-N CL in ACAD (see Table 16-6).
Table 16-6 Empirical critical loads for nitrogen in Acadia National Park, by
receptor, from Pardo et al. (2011c).
Critical Load (kg/ha/yr)
NPS	N Deposition Mycorrhizal	Herbaceous	Nitrate
Unit	Ecoregion	(kg/ha/yr)	Fungi Lichen Plant Forest Leaching
Acadia Eastern Temperate	5.2	5 to 12 4 to 8	17.5	3 to 8	8
NP Forests
ha = hectare; kg = kilogram; N = nitrogen; NP = national park; NPS = National Park Service; yr = year.
Ambient N deposition reported by Pardo et al. (20110) is compared to the lowest critical load for a receptor to identify potential
exceedance. A critical load exceedance suggests that the receptor is at increased risk for harmful effects.
16.1.3.1.1.2. Hubbard Brook Experimental Forest
Gbondo-Tuabawa and Driscoll (2002) used the PnET-BGC model to simulate the
responses of soil and surface water chemistry at the reference watershed at HBEF to
future emissions controls. Model performance was assessed using the normalized mean
absolute error and the efficiency as objective statistical criteria. The focus was on
comparing simulated chemistry with observed values between 1980 and 1998. Results
showed good agreement for stream Ca and SOr . However, stream NOs and Al
concentrations and soil solution Ca:Al ratios were over-predicted after 1990. Simulation
results suggested that emissions controls required by the CAAA will likely not result in
substantial changes in the future in "critical indicators such as soil base saturation, soil
solution Ca:Al, or stream ANC and Al concentrations."
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16.1.3.1.1.3. Bear Brook Watershed
Elvir et al. (2006) investigated the effects of increased N deposition on photosynthesis
and foliar nutrient content of sugar maple, American beech, and red spruce at the BBW.
Trees in the treated WB watershed (an addition of 25 kg N/ha/yr as [NFLJ2SO4) had
higher foliar N concentrations than EB reference trees. American beech and red spruce
displayed significantly lower foliar concentrations of Ca, Mg, and Zn. Sugar maple did
not show decreases in these nutrients and was the only species to have significantly
higher photosynthetic rates in the treated watershed as compared with the reference. This
result suggested that nutrient imbalances in American beech and red spruce in the
treatment watershed may have offset any potential photosynthetic benefits to these
species from higher N availability (Elvir et al.. 2006).
During the first 7 years of treatment at WB, sugar maple basal area increment (BAI) was
higher in WB compared with EB. After 8 years of treatment, however, the initial higher
sugar maple growth rate in WB decreased (Elvir et al.. 2010).
In the WB catchments, the BAI of sugar maple was enhanced 13 to 104% by the addition
of 25 kg N/ha/yr as (NFL^SC^. The BAI of red spruce was not significantly affected
(Elvir et al.. 2003).
Fernandez et al. (2003) reported results of quantitative soil excavations in 1998 that
measured soil pools of exchangeable base cations after 9 years of treatment at WB. The
treated watershed had lower concentrations of exchangeable Ca and Mg in all soil
horizons, with greater nutrient cation depletion in the O-horizon as compared with the
mineral soil. Depletion was also greater in conifer, as opposed to hardwood, stands. The
importance of treatment-caused base cation depletion was reinforced by MAGIC model
simulations. Estimates of watershed-wide exchangeable Ca and Mg (66 and 27 kg/ha)
were roughly comparable to the total cumulative excess stream Ca and Mg export in WB
after 9 years of treatment (55 and 11 kg/ha, respectively; Figure 16-7).
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5X
©
Qi
*
C3
3

Ca = calcium; ha = hectare; kg = kilogram; Mg = magnesium; yr = year.
Source: Fernandez et al. (2003).
Figure 16-7 Annual stream calcium and magnesium export (paired bars), and
cumulative excess export in West Bear Brook compared to East
Bear Brook (line), over the study period 1989-2000 at the Bear
Brook Watershed experiment.
16.1.3.1.1.4. Other Northeastern Regions
1	Pardo et al. (201 lc) compiled data on empirical CLs for protecting sensitive resources in
2	Level I ecoregions across the CONUS against nutrient enrichment effects caused by
3	atmospheric N deposition. Available data on empirical CL of nutrient-N in the Northeast
4	suggested that the lower end of estimates of the CL for resource protection was
5	3 kg N/ha/yr. Values at or above this level of deposition loading are considered to have
6	increased risk of harmful effects for mycorrhizal fungi, lichens, and forest vegetation
7	(Table 16-7). Keeping levels below this value also helps prevent NO, leaching into
8	drainage water.
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Table 16-7 Critical loads of nutrient nitrogen for the Northern Forests ecoregion.
Critical load or
Ecosystem N deposition
Component (kg/ha/yr) Reliability	Response	Comment	Reference
Tree	>3	#	Decreased growth of None	Thomas et al.
red pine, and decrease	(2010)
survivorship of yellow
birch, scarlet and
chestnut oak, quaking
aspen, and basswood
Lichens
4 to 6
(#)
Community
composition shift
Application of model
developed for marine
West Coast forest to
northern forests
Geiser et al.
(2010)
Ectomychorrizal
fungi
5 to 7
#
Change in fungal
community structure
None
Lilleskov et al.
(2008)
Herbaceous
species cover
>7 and <21
#
Loss of prominent
species
Response observed in
low-level fertilization
experiment
Hurd et al.
(1998)
Northern
hardwood and
coniferous forest
8
##
Increased surface
water NO3" leaching
None
Aber et al.
(2003)
Tree growth and
mortality
>10 and <26
#
Decreased growth
and/or induced
mortality
Response observed in
low-level fertilization
experiment in
old-growth montane
red spruce
McNultv et al.
(2005)
Arbuscular
mycorrhizal fungi
<12
(#)
Biomass decline and
community
composition change
Observed along a
Michigan N gradient
Pardo et al.
(2011c)
ha = hectare; kg = kilogram; N = nitrogen; N03 = nitrate; yr = year.
Reliability rate: ## reliable—a number of published papers show comparable results; # fairly reliable—the results of some studies
are comparable; (#) expert judgment—few empirical data are available, Critical Load based on expert judgment of those
ecosystems.
Source: Pardo et al. (2011c).
1	Thomas et al. (2010) analyzed Forest Inventory Analysis (FIA) data in the Northeast to
2	determine tree growth in response to a gradient of atmospheric N deposition from about 3
3	to 11 kg N/ha/yr. Some tree species showed increased growth across the N input gradient
4	(yellow poplar [.Liriodendron tulipifera], black cherry [Prunus serotina], and white ash
5	[Fraxinus americana]). Some showed highest growth at intermediate levels of N
6	deposition (quaking aspen [Populus tremuloides\ and scarlet oak [Quercus coccinea]).
7	Red pine (Pinus resinosa) exhibited growth decline across the gradient of increasing N
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deposition (Thomas et al.. 2010). Thus, it appeared that N deposition at ambient levels
can have both positive and negative effects on tree growth, depending on species and
deposition level.
In a high-elevation red spruce-balsam fir (Abies balsamea) forest in the Northeast, N
fertilization over 14 years led to a decrease in live basal area (LBA) with increasing N
additions. In control plots, LB A increased by 9% over the course of the study, while LB A
decreased by 18 and 40% in plots treated with 15.7 kg N/ha/yr and 31.4 kg N/ha/yr,
respectively (McNultv et al.. 2005).
There is little information available regarding the effects of N deposition on herbaceous
plants within northern hardwood forests in the Northeast. However, Hurd et al. (1998)
reported the results of experimental studies that added N at two and four times the
ambient N deposition level at several sites in the Adirondack Mountains. Herbaceous
plant coverage decreased after 3 years of fertilization, largely in response to shading
caused by enhanced growth of ferns. Additionally, information has been recently
published by Simkin et al. (2016) and is discussed in Appendix 62 of this ISA.
Modeling Studies
Evidence from research in northeastern forests indicates that direct and indirect effects of
climate change will cause changes in N and other biogeochemical cycling by altering
plant physiology, forest productivity, and soil processes. Campbell et al. (2009) reviewed
these relationships (Figure 16-8) and applied the PnET-BGC model in a northern
hardwood forest ecosystem at HBEF to test assumptions about interactions regarding
climate change. Model results suggested an increase in primary productivity due to a
longer growing season in the future, an increase in NOs leaching due to enhanced net
mineralization and nitrification, and a slight decrease in mineral weathering due to
reduced simulated soil moisture (negative effect) and increased temperature (positive
effect).
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Temperature, Precipitation, ""-v
and Extremu Weather Hvents ¦'
Fire
Insects-pathogens
Invasive species
A forest composition
Feedbacks
Soil trace gas fluxes
Plant respiration
Deforestation
Transpiration
Evaporation
Albedo
Soil abiotic: process *":l ¦ l""" ;	Soil biotic processes
ion exchange Hedox reactions	Decomposition
Adsorption-deserption	Mineralization
Weathering Stream export	Immobilization
Solute transport ^	Respiration
Ecosystem Services
Water quality and quantity
Carbon sequestration
Economics (wood products,
tourism, maple sugar)
Recreation and aesthetics
NPP = net primary production.
Source: Campbell et al. (2009).
Figure 16-8 Conceptual diagram showing the direct and indirect effects of
changes in temperature and precipitation on biogeochemical
processes in forests and on the services forests provide. Also
shown are feedbacks that further influence climatological effects.
16.1.3.2. Aquatic
1	This section presents post-2000 research findings at ACAD, HBEF, and BBW on the
2	dose-response relationships of N and S to aquatic ecology, as well as the critical N and S
3	loads for maintaining ecosystem health. Additional relevant information for the Northeast
4	region is summarized. Both empirical and modeling studies of aquatic dose-response
5	relationships and critical loads are included in this section.
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16.1.3.2.1. Empirical Studies
1	Post-2000 findings from empirical studies of aquatic dose-response relationships and
2	critical loads are summarized in this section. Table 16-8 compiles the body of empirical
3	research identified.
Table 16-8 Aquatic empirical research on the response of nitrogen and sulfur
deposition for the northeastern U.S.
Variable
Response
Deposition
Addition Years
Site
Reference
Fish
Shift in the upper
Not specified
0 1960S-2007
HBEF
Warren et al.
community
mainstem of HBEF



(2008)
including
from the presence of




brook trout
at least three fish




(Salvelinus
species to only brook




fori tin alis)
trout (Salvelinus





fontinalis)




S042"
W6 had a 32%
Not specified
0 1963-1994
HBEF
Driscoll et al.

decline in the annual



(2001b)

volume-weighted



Likens et al. (2001)

concentration




(-1.1 [jeq/L/yr)




S042" + NOs"
W6 had declines in
Not specified
0 1963-1994
HBEF
Driscoll et al.

stream



(2001b)

concentrations of





strong acids





(-1.9 [jeq/L/yr)




Sum of base
-1.6 [jeq/L/yr
Not specified
0 1963-1994
HBEF
Driscoll et al.
cations




(2001b)
PH
Small but significant
Not specified
0 1963-1994
HBEF
Driscoll et al.

increases In stream



(2001b)

pH, from 4.8 to 5.0




S042"
Biogeochemical
Not specified
0 1965-2008
HBEF
Mitchell and Likens

control of SO42"



(2011)

export from forested





watersheds has





shifted from





atmospheric S





deposition to climatic





factors that regulate





soil moisture.




pH and ANC Decrease	Not specified 1,800 eq/ha/yr 7 yr	BBW Norton et al. (2004)
(NH4)2SC>4 acidification
treatment
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Table 16-8 (Continued): Aquatic empirical research on the response of nitrogen
and sulfur deposition for the northeastern U.S.
Variable
Response
Deposition
Addition
Years
Site
Reference
Stream
dissolved
S042"
Decreased
Not specified
1,800 eq/ha/yr
(NH4)2S04
1989-2007
BBW
Norton etal. (2004)
Fatemi etal. (2012)
Base cations
Decreased
Not specified
1,800 eq/ha/yr
(NH4)2S04
1989-2007
BBW
Norton etal. (2004)
Fatemi etal. (2012)
Al+
Increased
Not specified
1,800 eq/ha/yr
(NH4)2S04
1989-2007
BBW
Norton etal. (2004)
Fatemi etal. (2012)
ANC
Regional surface
water ANC did not
change significantly in
New England during
the 1990s
Not specified
0
NA
NE
U.S. EPA (2003)
Evans and
Monteith (2001)
Release of S
from internal
storage pools
to drainage
water
Increased
Not specified
0
1965-2008
General
Mitchell and Likens
(2011)
Al+ = aluminum; ANC = acid neutralizing capacity; BBW = Bear Brook Watershed; ha = hectare; HBEF = Hubbard Brook
Experimental Forest; kg = kilogram; L = liter; |jeq = microequivalents; eq = equivalents; NE = northeastern;
(NH4)2S04 = ammonium sulfate; N03" = nitrate; S = sulfur; S042" = sulfate; W6 = Watershed 6; yr = year.
16.1.3.2.1.1. Acadia National Park
No empirical aquatic critical loads or dose-response studies have been identified in the
literature for the ACAD case study area since 2000.
16.1.3.2.1.2. Hubbard Brook Experimental Forest
The cycling of N, S, and other elements in the Northeast is closely linked with aspects of
climate. Many recent studies have explored some of those linkages, and many of the
empirical studies of aquatic dose-response relationships have included aspects of climate
change. The potential for a shift from atmospheric deposition to climatic regulation of
watershed S biogeochemistry at HBEF was examined by (Mitchell and Likens. 2011).
More than four decades of continuous long-term data collected from four watersheds
were used to evaluate S budgets. "Analyses focused on the role of changing water
availability in affecting the amount of S mobilized from internal watershed sources and
the resultant changes in the S budget discrepancies over time... The biogeochemical
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controls on annual SO42 export in stream water [appear to have] shifted from
atmospheric S deposition to climatic factors that affect soil moisture."
With declining inputs of atmospheric S, the higher outputs of SO42 in drainage waters
relative to precipitation inputs appear to be driven by the S stored in the soil. These
biogeochemical responses at the HBEF might be amplified with further climate change,
resulting in greater annual stream discharge and watershed wetness. Such climatic change
will potentially increase SO42 mobilization and hence may slow the recovery of both
aquatic and terrestrial ecosystems from acidification.
16.1.3.2.1.3.	Bear Brook Watershed
Chadwick and Hurvn (2005) assessed the effects of N additions on secondary stream
macroinvertebrate production in BBW. Production did not vary between streams
(reference and treatment) but was 35% higher for both streams in the second year of the
study. Results suggested that patterns of litter input and channel drying, rather than N
input, controlled secondary production in these intermittent streams by altering resource
availability and invertebrate community structure. It appeared that these variables
overrode the effects of N supply, perhaps because N is not limiting.
Ongoing climate change will likely affect a wide range of biogeochemical processes in
northeastern forested ecosystems. Of particular interest are the effects on flow volumes
and peak flows. Kim et al. (2010) analyzed a nearly two-decades-long stream flow record
at EB and a longer record for the Narraguagus River (a proximate watershed to the
BBW). The focus was on improving understanding of high-flow events that have a
disproportionate impact on biogeochemical processes and fluxes. A moving window
analysis to evaluate the changing flood potential over time suggested upward trends in
the occurrence of high-flow events during recent decades.
16.1.3.2.1.4.	Other Northeastern Regions
Considering trends across the Northeast, Aber et al. (2003) found that surface water NO3
concentrations exceeded 1 (j,eq/L mainly in northeastern watersheds receiving more than
about 9 to 13 kg N/ha/yr deposition. Above this range, mean NO3 export increased
linearly with increasing deposition at a rate of about 0.85 kg NO3 -N/ha/yr for every
1 kg N/ha/yr increase in deposition, although at higher rates of deposition there was
considerable variability in N retention among watersheds (Aber et al.. 2003).
N limitation across the Northeast was evaluated by Baron etal. (2011a) who estimated
that 34% of 4,361 New England lakes represented in the Eastern Lakes Survey were
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likely to be N limited, based on having dissolved inorganic N (DIN):total P (TP) ratio (by
weight) less than 4. Although eutrophication is a concern, there is no published evidence
that eutrophication is occurring to any appreciable extent in the lakes in ACAD.
Dissolved organic matter (DOM) is important to a range of processes critical to aquatic
ecosystem functioning, including providing a microbial food source, attenuating light,
buffering pH, binding metals, and controlling the cycling of C, N, Hg, and P. Shifts in the
quality of DOM may affect aquatic ecosystem functioning. A chemical signature of
terrestrial DOM was used by SanClements et al. (2012) to support the hypothesis that
increased dissolved organic carbon (DOC) concentrations in surface water in the
Northeast since about 1993 have been driven mainly by decreasing acidic deposition and
increasing solubility of soil organic matter. Fluorescence spectroscopy was used to
characterize the quality of DOM of archived samples from nine acid-sensitive lakes in
Maine. Decadal decreases in SOr in lake water were correlated with increased DOC
concentrations and a shift from microbial to terrestrially derived DOM during ecosystem
recovery from prior acidification.
Aquatic biota in the Northeast have been affected by acidification at virtually all levels of
the food web. Some species and some lifestages are especially sensitive. Decreases in
ANC and pH and increases in inorganic Al concentration contributed to declines in
species richness and abundance of zooplankton, macroinvertebrates, and fish as
documented in the Adirondacks (Sutherland et al.. 2015; Nierzwicki-Bauer et al.. 2010;
Keller and Gunn. 1995; Schindler et al.. 1985). Although some species are favored by
increased acidity, the overall species richness typically decreases as surface water acidity
increases.
Modeling Studies
Post-2000 findings from modeling studies of aquatic dose-response relationships and
critical loads are summarized in this section. Table 16-9 compiles the body of research
identified.
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Table 16-9 Critical and target load and exceedance modeling studies in the
northeastern U.S.
Reference
Location Model
Focus
Dupont et al. (2005)
NE region SSWC
CLs for acidity to protect aquatic biota


to pH 6.0
SanClements et al. (2010)	BBW	PnET-BGC	Interactions between climate change
parameters and atmospheric
deposition
Phelan et al. (2016)	BBW and ForSAFE-VEG	Atmospheric deposition to northern
HBEF	hardwood forests
Tominaqa et al. (2010)
HBEF
MAGIC, PnET-BGC,
Model output comparison


SAFE, VSD

Pourmokhtarian etal. (2012)
HBEF
PnET-BGC
N cycling and climate change
BBW = Bear Brook Watershed; CL = critical load; ForSAFE-VEG = Soil Acidification in Forest Ecosystems; HBEF = Hubbard
Brook Experimental Forest; MAGIC = Model of Acidification and Groundwater in Catchments; N = nitrogen; NE = northeastern;
PnET-BGC = Photosynthesis and Evapotranspiration-Biogeochemical; SAFE = Soil Acidification in Forest Ecosystems;
SSWC = Steady-State Water Chemistry; VSD = Very Simple Dynamic.
16.1.3.2.2.1.	Acadia National Park
No modeling studies on aquatic critical loads for ACAD have been identified in the
literature.
16.1.3.2.2.2.	Hubbard Brook Experimental Forest
Tominaga et al. (2010) evaluated the performance and prediction uncertainty of four
dynamic process-based watershed acidification models: MAGIC, PnET-BGC, SAFE, and
VSD. Model output was assessed by systematically applying each of the models to data
from the HBEF. The models were used to assess future soil and stream chemistry
response to reduced levels of atmospheric S deposition. Both hindcast and forecast
predictions were qualitatively similar across the four models. Nevertheless, the temporal
patterns of projected stream ANC and soil base saturation differed. Tominaga et al.
(2010) concluded that these differences can be accommodated by employing multiple
models. Nevertheless, these results have implications for individual model applications.
Pourmokhtarian et al. (2012) used "the PnET-BGC model to evaluate the effects of
potential future changes in temperature, precipitation, solar radiation, and atmospheric
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CO2 on the pools, concentrations, and fluxes of major elements at the HBEF." The
climate projections were based on downscaled climate output from atmospheric-ocean
general circulation models. These climate projections indicated that the average air
temperature will likely increase at the HBEF site by 1.7 to 6.5°C over the 21st century,
with simultaneous increase in annual average precipitation ranging from 4 to 32 cm
above the long-term mean. The PnET-BGC model simulations under expected future
climate showed shifts in hydrology from later snowpack development, earlier snowmelt,
increased evapotranspiration, and a slight increase in the annual water yield. The model
results suggested that net soil N mineralization and nitrification will markedly increase
under elevated temperature. This increase resulted in simulated acidification of soil and
stream water, altering the quality of water draining from the forested watershed.
16.1.3.2.2.3.	Bear Brook Watershed
No aquatic critical loads modeling studies have been identified for BBW in the literature
since 2000.
16.1.3.2.2.4.	Other Northeastern Regions
Studies in the Northeastern region have used a variety of steady-state and dynamic
modeling approaches. The Conference of the New England Governors and Eastern
Canadian Premiers (NEG/ECP) sponsored a modeling assessment of steady-state CLs for
protection of forest soils and lakes against acidification in the Northeast region and in
eastern Canada (Ouimct et al.. 2006; Dupont et al.. 2005; Ouimet et al.. 2001). Dupont et
al. (2005) reported the SSWC steady-state aquatic CLs of acidity and associated
exceedances for lakes. Atmospheric acid loads were assessed based on atmospheric
deposition of both S and N, using a critical limit of pH 6 to protect aquatic biota. This pH
level approximately corresponds with ANC = 40 |icq/L in this region (Small and Sutton.
1986). Estimated S critical loads were exceeded in 2002 for 12.3% of all studied lakes.
Lakes having lowest calculated CLs included many in southern Vermont, eastern and
northern Maine, northern New Hampshire, and Cape Cod. Exceedances of CLs, based on
estimated acidic deposition in 2002, were highest in central and coastal Massachusetts,
southern Vermont, much of Maine, and portions of New Hampshire. Eastern Maine and
southern Vermont were notable "hot spots" where ambient S + N deposition exceeded
CLs by more than 10 meq/m2/year (Dupont et al.. 2005).
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16.1.3.3. Integration
Table 16-10 contains a list of northeastern U.S. critical load determinations by multiple
researchers, ranging from 3-8 kg N/ha/yr (forests) to 17.5 kg N/ha/yr (herbaceous
plants).
Table 16-10 Empirical and modeled nitrogen critical loads applicable to the
northeastern U.S.
Critical Load
(kg N/ha/yr)
Ecosystem
Component
Response
N Deposition
(kg/ha/yr)
Site
Reference
5 to 12
Eastern
temperate
forest
Mycorrhizal Fungi
5.2
ACAD
Pardo et al.
(2011c)
4 to 8
Lichens
Epiphytic lichen community
change
5.2
ACAD
Pardo et al.
(2011c)
17.5
Herbaceous
Species
Increases in nitrophilic species,
declines in species-richness
5.2
ACAD
Pardo et al.
(2011c)
3 to 8
Hardwood
forest
Not specified
5.2
ACAD
Pardo et al.
(2011c)
Ellis et al. (2013)
8	Eastern	Increased surface water NO3	5.2 ACAD Pardo et al.
hardwood leaching	(2011c)
forests	Ellis et al. (2013)
ANC crucial Not specified Current legislated emissions
concentration	(= 13% reduction of SO42" by
= 20 peq/L	2015) results in limited response
Soil base	recovery
saturation	Maximum feasible technology
critical level of	reductions (= 78% reduction of
10%	SO42" by 2015) results in a more
rapid and greater extent of
chemical recovery
Not specified HBEF
Tominaqa et al.
(2010)
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Table 16-10 (Continued): Empirical and modeled nitrogen critical loads applicable
to the northeastern U.S.
Critical Load
Ecosystem

N Deposition


(kg N/ha/yr)
Component
Response
(kg/ha/yr)
Site
Reference
CO2
Not specified
N cycling and climate change.
Not specified
HBEF (year
Pourmokhtarian
fertilization

"Under elevated temperature, net

2070-2100)
etal. (2012)
plateau of

soil N mineralization and



600 ppm

nitrification markedly increase,





resulting in acidification of soil and





stream water... Invoking CO2





fertilization effect on vegetation





under climate change





substantially mitigates watershed





N loss...Showed recovery of up to





1 pH unit (assuming pH of 6.0 as





steady-state value) and...ANC of





6.9 to 15.5 peq/L in comparison to





-3.4 peq/L mean annual





measured values (1988-2000)"



Not specified
Eastern
Interactions among climate
Not specified
BBW
SanClements et

temperate
change parameters and


al. (2010)

forest
atmospheric deposition



Not specified
Eastern
Atmospheric deposition to
Not specified
BBW and
Phelan et al.

temperate
northern hardwood forests

HBEF
(2016)

forest




Not specified
NE
Surface water NO3"
9 to 13
NE
Aber et al. (2003)

watersheds
concentrations exceeded 1 peq/L





and mean NO3" export increased





linearly with increasing deposition



<20 kg/ha/yr
Aquatic biota
Protect aquatic biota to pH 6.0
Not specified
NE
DuDont et al.
S + N (= 9.6 kg




(2005)
S042"/ha/yr)





ACAD = Acadia National Park; ANC = acid neutralizing capacity; BBW = Bear Brook Watershed; C02 = carbon dioxide;
ha = hectare; HBEF = Hubbard Brook Experimental Forest; kg = kilogram; L = liter; |jeq = microequivalents; N = nitrogen;
NE = northeastern; N03" = nitrate; ppm = parts per million; S = sulfur; S042" = sulfate; yr = year.
16.1.4.	Long-Term Ecological Monitoring
1	This section summarizes research on the long-term effects of S and N deposition in the
2	case study areas. The focus is primarily on research published since about the Year 2000.
3	The acidification and nutrient enrichment subsections are each organized by three case
4	study areas followed by relevant information for the overall Northeast region. Key
5	publications are summarized in Table 16-11.
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Table 16-11 Summary table of observed terrestrial and aquatic acidification
long-term trends in Hubbard Brook Experimental Forest and Bear
Brook Watershed.
Variable
Species
Response
Deposition
(kg/ha/yr)
Addition
(kg/ha/yr)
Site
Reference
C:N
NA
Significant increase over
25 yr
Not
specified
Not
specified
HBEF
Campbell et al.
(2007)
Biomass
Sugar
maple
Decrease
Not
specified
Not
specified
HBEF
Campbell et al.
(2007)
Al+
NA
Decrease
Not
specified
Not
specified
HBEF
Campbell et al.
(2007)
Base cation
Ca and Mg
depletion
NA
Historical forest cutting had
little impact on
exchangeable cation soil
pools
Not
specified
Not
specified
HBEF
Gbondo-
Tuabawa and
Driscoll (2003)
SO42"	NA	Watershed 6, had a 32% Not	Not	HBEF
decline in the annual	specified specified 1963-1994
volume-weighted
concentration
(-1.1 [jeq/L/yr)
SCM2" + NOs" NA
Declines in stream
Not
Not
HBEF
Driscoll et al.

concentrations of strong
specified
specified
1963-1994
(2001b)

acids (-1.9 [jeq/L/yr)




Sum of base NA
-1.6 [jeq/L/yr
Not
Not
HBEF
Driscoll et al.
cations

specified
specified
1963-1994
(2001b)
pH	NA	Small but significant	Not	Not	HBEF	Driscoll et al.
increases in stream pH, specified specified 1963-1994 (2001b)
from 4.8 to 5.0
Driscoll et al.
(2001b)
Likens et al.
(2001)
Fish
community
Slimy
sculpin
(Cottus
cognatus),
blacknose
dace
(Rhinichthys
atratulus),
and brook
trout
(Salvelinus
fontinalis)
Shift in the upper mainstem
of Hubbard Brook from the
presence of at least three
fish species (slimy sculpin
[Cottus cognatus],
blacknose dace
[Rhinichthys atratulus], and
brook trout [Salvelinus
fontinalis]) to only brook
trout
Not	Not	HBEF	Warren et al.
specified specified (1960s-2007) (2008)
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Table 16-11 (Continued): Summary table of observed terrestrial and aquatic
acidification long term trends in Hubbard Brook
Experimental Forest and Bear Brook Watershed.



Deposition
Addition


Variable
Species
Response
(kg/ha/yr)
(kg/ha/yr)
Site
Reference
Base cation
Sugar
American beech and red
8.4 N
WB
BBW
Elvir et al.

maple,
spruce had lower foliar Ca,
(wet + dry)
25.2 kg

(2006)

American
Mg, Zn concentrations;

N/ha/yr



beech, red
nutrient imbalance may

(NH4)2S04



spruce
offset potential

WB




photosynthesis benefits

28.8 S




Sugar maple higher






photosynthesis rates/no






decrease in Ca, Mg, Zn






concentrations




Base cation
Sugar
Little evidence of BC
18.5 (1980)
WB
BBW
SanClements

maple,
depletion but confounded
4.74 (2010)
25.2 kg

et al. (2010)

American
by ice storm litter
Wet S042"
(NH4)2S04



beech, red
mineralization
2.8
WB



spruce

Inorganic N
28.8 S


Mineralization
Sugar
Storm caused increased
Not
WB
BBW
Fernandez et

maple,
litterfall and accelerated
specified
25.2 kg

al. (2003)

American
mineralization, obstructing

N/ha/yr



beech, red
temporal trends in soil

(NH4)2S04



spruce
chemistry (17 yr)

WB






28.8 S


pH and ANC
NA
Decrease
Not
WB
BBW
Norton et al.



specified
25.2 kg

(2004)




N/ha/yr






(NH4)2S04






WB






28.8 S


Stream
NA
Decrease
Not
WB
BBW
Norton et al.
dissolved


specified
25.2 kg

(2004)
S042"



N/ha/yr

Fatemi et al.




(NH4)2S04

(2012)




WB






28.8 S


Base cation
NA
Increased BC export.
Not
WB
BBW
Norton et al.
export in

Export rates declined after
specified
25.2 kg

(2004)
runoff

7 yr treatment.

N/ha/yr

Fatemi et al.




(NH4)2S04

(2012)




WB






28.8 S


Al+
NA
Increase
Not
WB
BBW
Norton et al.



specified
25.2 kg

(2004)




N/ha/yr

Fatemi et al.




(NH4)2S04

(2012)
WB
28.8 S
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Table 16-11 (Continued): Summary table of observed terrestrial and aquatic
acidification long term trends in Hubbard Brook
Experimental Forest and Bear Brook Watershed.
Variable
Species
Response
Deposition
(kg/ha/yr)
Addition
(kg/ha/yr)
Site
Reference
Release of S
from internal
storage pools
to drainage
water
NA
Increased
Not
specified
0
General
(1972-2008)
Mitchell and
Likens (2011)
S042"
NA
Biogeochemical control of
SO42" export from forested
watersheds in HBEF has
shifted from atmospheric S
deposition to climatic
factors that regulate soil
moisture
Not
specified
0
General
Mitchell and
Likens (2011)
Critical loads
Varied
Varied
Not
specified
Not
specified
NE
Pardo et al.
(2011c)
ANC
NA
Regional surface water
ANC did not change
significantly in New
England during the 1990s
Not
specified
0
NE
U.S. EPA
(2003)
Al+ = aluminum; ANC = acid neutralizing capacity; BC = base cation; BBW = Bear Brook Watershed; C = carbon; Ca = calcium;
ha = hectare; HBEF = Hubbard Brook Experimental Forest; kg = kilogram; L = liter; |jeq = microequivalents; Mg = magnesium;
N = nitrogen; NA = not applicable; NE = northeastern; (NH4)2S04 = ammonium sulfate; N03" = nitrate; S = sulfur; S042" = sulfate;
W6 = Watershed 6; WB = West Bear Brook; yr = year; Zn = zinc.
Acadia National Park is not a federally funded long-term monitoring study area.
16.1.4.1. Long-Term Monitoring of Acidification
Acidic deposition has been shown to be an important factor causing decreases throughout
much of the Northeast region in concentrations of exchangeable base cations in soils,
which were naturally low historically. Base saturation values less than 12% predominate
in the B-horizon in portions of the region where soil and surface water acidification from
acidic deposition have been most pronounced (Sullivan et al.. 2006a; Bailey et al.. 2004;
David and Lawrence. 1996).
Mitchell et al. (2011) evaluated S mass balances for 15 watersheds in the northeastern
U.S. and southeastern Canada, including watersheds at HBEF, BBW, and Cone Pond in
New Hampshire, for the period 1985 to 2002. Most study watersheds showed evidence of
internal watershed sources of SO42 . likely from mineralization of organic S stored from
decades of high S deposition. Mobilization of the mineralized S contributed an estimated
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1-6 kg S/ha/yr to the stream fluxes of SO42 . This mobilization has affected the observed
rates of recovery from acidification as S deposition has declined.
Surface water chemistry in the northeastern U.S. has been characterized, surveyed,
resurveyed, and monitored in many studies conducted over the last several decades
(Table 16-12V
Table 16-12 Example surface water acidification chemistry studies in the
Northeast case study region.




Time


Focus
Results
Deposition
Addition
Period
Site
Reference
N export
An unburned study
Not
Not applicable
Pre- vs.
ACAD
Nelson et al.

watershed exported 10 to
specified

post-1947

(2007)

20 times more inorganic N


fire



than the burned watershed;






in addition, retention of






inorganic N was 96% in the






burned watershed vs. 72%






in the unburned watershed.





N export
Total N + P exported by
Not
Not applicable
Not
ACAD
Nelson et al.

watersheds entirely within
specified

specified

(2007)

ACAD were significantly






lower than exports in






watersheds that were






partly or completely outside






the park.





Reference
Decline in stream SO42"
Not
Not specified
1963-1994
HBEF
Driscoll et al.
stream
(-1.1 [jeq/L/yr)
specified



(2001b)
chemistry
Decline in SO42" + NO3" in






stream (-1.9 peq/L/yr)





S budget
Biogeochemical control of
Not
Not specified
1965-2008
HBEF
Mitchell and

SO42" export has switched
specified



Likens (2011)

from atmospheric






deposition to climate






factors that regulate soil






moisture.





Stream
Stream dissolved SO42",
Not
1,800 eq/ha/yr
1987-2000
BBW
Norton et al.
trends for
pH, and ANC decrease
specified
(NH4)2S04


(2004)
reference
Base cation and Al+





and
increase





treatment






catchments






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Table 16-12 (Continued): Example surface water acidification chemistry studies in
the northeast case study region.
Time
Focus	Results	Deposition Addition Period	Site	Reference
Low-flow and Base cation, NH4+and CI" Not	1,800 eq/ha/yr 1987-2006 BBW	Navratil et al.
high-flow unchanged	specified (NH4)2S04	(2010)
stream
chemistry of
reference
and
treatment
catchments
Chronic
stream
chemistry for
reference
and
treatment
catchments
Increased BC export;
declined after 7 yr of
treatment.
Decrease pH and alkalinity
Increased dissolved Al+,
NOs", S042"
Not	1,800 eq/ha/yr 1987-2007 BBW	Norton etal.
specified (NH4)2S04	(2010)
Episodic
stream
chemistry for
reference
and
treatment
catchments
"18 yr of N and S addition
have not affected the
natural drivers of episodic
acidification. The
contribution of SO42" to the
ANC decline in WB has
been increasing linearly
since the beginning of
watershed treatment while
the role of NO3" has
remained relatively
constant after an initial
increase."
Not	1,800 eq/ha/yr 1988-2006
specified (NH4)2S04
BBW
Laudon and
Norton (2010)
Lake	Ubiquitous decreases in Not	Not specified 1986 and NE region Warbv et al.
resurvey the concentration of	specified	1998	(2009)
inorganic Al across the
region.
Organic monomeric Al also
declined region-wide in
New England.
In 2001, only seven study
lakes (representing
130 lakes in the Northeast
region) that had AN above
the toxic threshold of 2 |jM,
compared with 20 sampled
lakes (representing
449 lakes in the
population) in 1986.
Chronic and	Rates of change for	Not	Not specified 1990-2000 NE and	U.S. EPA
episodic lake individual water bodies	specified	Appalachian (2003)
and stream	ranged from about-1.5 to	Mtns.
chemistry	-3 peq/L/yr.
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5
6
7
8
9
10
11
12
Table 16-12 (Continued): Example surface water acidification chemistry studies in
the northeast case study region.
Focus
Results
Deposition Addition
Time
Period
Site
Reference
Regional NO3" concentrations in
lake trends New England and the
Adirondack Mountains,
which had no trend prior to
2000, declined at a rate of
-0.05 [jeq/L/yr
Not	Not specified 2000-2010 New England Strock et al.
specified	and	(2014)
Adirondacks
Response to Oa-horizon BS decreased Not
deposition from 56.2 to 33.0% and specified
reduction almost equivalent changes
in carbon-normalized
exchangeable Ca and
exchangeable Al
Not specified
1984 to
2001
NE region
Warbv et al.
(2009)
Suggested a nascent	Not	Not specified 1992-1993 General Lawrence et
recovery of soil acid-base specified	vs.	al. (2012)
chemistry at some	2003-2004
locations in the Northeast
where six red spruce
stands sampled.
Exchangeable Al in the
Oa-horizons decreased
(p < 0.05) by 20 to 40% at
all New England sites.
Evidence of base cation
depletion and decreased
Ca:AI ratio
ACAD = Acadia National Park; Al = aluminum; ANC = acid neutralizing capacity; BBW = Bear Brook Watershed; BC = base cation;
BS = base saturation; Ca = calcium; CI" = chloride; ha = hectare; HBEF = Hubbard Brook Experimental Forest; kg = kilogram;
L = liter; |jeq = microequivalents; |jM = micromole; N = nitrogen; NE = northeastern; NH4+ = ammonium; (NH4)2S04 = ammonium
sulfate; N03" = nitrate; Oa = well decomposed humus; P = phosphorus; S = sulfur; S042" = sulfate; WB = West Bear Brook;
yr = year.
These studies have focused on both chronic and episodic responses of reference and (in
the case of BBW and HBEF) experimentally treated catchments.
The U.S. EPA's Long-Term Monitoring (LTM) Program has been collecting monitoring
data since the early 1980s for many lakes and streams in acid-sensitive areas of the U.S.,
including within the Northeast region. These data allow evaluation of trends and
variability in key components of lake and stream water chemistry prior to, during, and
subsequent to 1990 CAAA Title IV implementation. Throughout the Northeast region,
the concentration of SO42 in surface waters has decreased substantially, often by
one-third to one-half, subsequent to the 1990 CAAA. These declines in lake and stream
S042 concentrations were considered consistent with observed declines in S wet
deposition and have been corroborated by other studies showing that SO42
concentrations in northeastern lakes have decreased steadily since about the late 1970s
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[e.g., (U.S. EPA. 2003; Driscoll et al.. 1995)1. Regional surface water ANC did not
change significantly in New England during the 1990s (U.S. EPA. 2003).
The impact of the 1990 CAAA relevant to aluminum releases to surface waters was
assessed by Warbv et al. (2008). who surveyed 113 lakes in the Northeast that had been
sampled in 1986 and again in 2001. They "found ubiquitous decreases in the
concentration of inorganic Al across the region." In 2001, there were only 7 study lakes
(representing 130 lakes in the Northeast region) that had inorganic Al above the toxic
threshold of 2 (j,M, compared with 20 sampled lakes (representing 449 lakes in the
population) in 1986.
Strock et al. (2014) analyzed recent trends in wet deposition and lake chemistry using
long-term monitoring data for lakes in New England and the Adirondack Mountains.
From 2000 to 2010, lake NOs concentrations decreased at a rate of-0.05 (j,eq/L/year (no
trend before 2000). There was a shift to nontoxic organic Al. Both ANC and pH
exhibited variable trends.
Table 16-13 summarizes the effects of N deposition on watershed nutrient in the three
case studies areas and relevant Northeastern regional studies.
Table 16-13 Summary of observed terrestrial and aquatic nutrient enrichment
long-term responses in Acadia National Park, Hubbard Brook
Experimental Forest, Bear Brook Watershed, and other northeastern
regions.
16.1.4.2. Long-Term Monitoring of Nitrogen Enrichment
Variable
Species
Response
Deposition Addition
(kg/ha/yr) (kg/ha/yr) Site Reference
N export
Not specified An unburned study	Wet
Not	ACAD	Nelson et
specified (pre- vs. al. (2007)
post-1947
watershed exported 10 to	deposition
20 times more inorganic	only (value
N than the burned	not specified)
watershed.
fire)
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Table 16-13 (Continued): Summary of observed terrestrial and aquatic nutrient
enrichment long-term responses in Acadia National
Park, Hubbard Brook Experimental Forest, Bear Brook
Watershed, and other northeastern regions.
Variable
Species
Response
Deposition
(kg/ha/yr)
Addition
(kg/ha/yr)
Site
Reference
N mass balance Sugar maple,
American
Beech,
yellow birch
Stream water export
decrease from 4 to
1 kg N/ha/yr
Shift to a net N sink
storing ~8 kg N/ha/yr;
unclear whether N is
accumulating or being
lost through
denitrification.
~7 kg N/ha/yr 0
HBEF
Yanai et
al. (2013)
DIN loss	Sugar maple, In W6, "losses were	"0.13g/m2/yr 0	HBEF	Aber et al.
American elevated in 1960s by a is 20% of	(2002)
Beech,	combination of recovery wet + dry
yellow birch from extreme drought and deposition"
a significant defoliation
event. N deposition
alone, in the absence of
other disturbances, would
have increased DIN
losses by 0.35 g N/m2/yr."
Biogeochemical Sugar maple, Modeled soil freeze-thaw Not specified 0	HBEF	Campbell
process	American to Year 2100. "Shortened	et al.
Beech,	frost covered period has	(2010)
yellow birch biogeochemical process
implications."
Biogeochemical Sugar maple, "A relatively mild freezing Not specified 0	HBEF	Groffman
process	American event induced significant	et al.
Beech,	increases in N	(2011)
yellow birch mineralization and
nitrification rates, solute
leaching, and soil N2O
production and caused
significant decreases in
soil methane uptake. Soil
freezing events may be
major regulators of soil
biogeochemical
processes and solute
delivery to streams in
forested watersheds."
NO3" export in Sugar maple, NO3" retention greater ~6 kg N/ha/yr	HBEF	Judd et al.
2006	American than expected. "Changes (1999-2008)	(2011)
Beech,	over last five decades
yellow birch have reduced impacts of
frost events on watershed
NO3" export."
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Table 16-13 (Continued): Summary of observed terrestrial and aquatic nutrient
enrichment long-term responses in Acadia National
Park, Hubbard Brook Experimental Forest, Bear Brook
Watershed, and other northeastern regions.
Deposition Addition
Variable	Species	Response	(kg/ha/yr) (kg/ha/yr) Site Reference
NPP and	Sugar maple, Ca nutrition promoted Not specified 0	HBEF	Battles et
photosynthetic American higher aboveground NPP	(15-yr study) al. (2014)
surface area Beech,	and increased
yellow birch photosynthetic surface
area.
Nitrification	Showed greatest	8.4 kg N/ha/yr WB 25.2 kg BBW	Jefts et al.
response to N treatments, (wet + dry) N/ha/yr	(2004)
(NH4)2S04
WB 28.8 S
(3-1,4
glucosidase
(3-1,4-N-
acetylglucos-
aminidase
Sugar maple, "Greatest leaf N + P
American contents can increase
beech, Red short-term
spruce	decomposition,
accelerate production of
more humic-like water
extractable organic
matter and thereby
potentially influence the
distribution of organic
matter within the soil C
pool."
Not affected by long-term
N enrichment in aquatic
and terrestrial habitats.
Not specified 25.2 N
BBW
since 1989
Hunt et al.
(2008)
Mineau et
al. (2014)
WB vs. EB Not specified Little difference found
streams: Leaf
breakdown
Not specified Not	BBW	Simon et
specified	al. (2010)
WB vs. EB Not specified "Virtually identical"
streams:
Invertebrate
production
Not specified Not	BBW	Simon et
specified	al. (2010)
WB vs. EB NA	N uptake responsive
streams:
Nutrient uptake
Not specified Not	BBW	Simon et
specified	al. (2010)
Base cation
Sugar maple, American beech and red 8.4 kg N/ha/yr WB 25.2 kg BBW
American spruce lower foliar Ca,
beech, red Mg, Zn concentrations;
spruce	nutrient imbalance may
offset potential
photosynthesis benefits.
Sugar maple higher
photosynthesis rates; no
decrease in Ca, Mg, Zn
concentrations.
(wet + dry)
N/ha/yr
(NH4)2S04
WB 28.8 S
Elvir et al.
(2006)
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Table 16-13 (Continued): Summary of observed terrestrial and aquatic nutrient
enrichment long-term responses in Acadia National
Park, Hubbard Brook Experimental Forest, Bear Brook
Watershed, and other northeastern regions.
Variable	Species	Response
Watershed N Varied	Varies widely. Not directly
retention	related to N loading.
Deposition Addition
(kg/ha/yr) (kg/ha/yr) Site Reference
Wet	0	Mid-Atlantic Campbell
deposition 1.8	and NE et al.
to 5.5 NOs"	forests	(2004b)
0.9 to
2.4 NH4+
Net nitrification Picea rubens Picea rubens density Not specified 0	NE: Ranch	Ross and
influences net nitrification.	Brook	Wemple
Watershed,	(2011)
Vermont
BAI
Sugar maple,
black cherry
Varied species with acid
deposition impact on soil
nutrient status.
Maple responds (growth)
to lime (Ca) addition.
Not specified
Not
specified
NE (7 sites)
(1937-1996)
Lonq et al.
(2009)
SOC response
Hardwood
Increased cumulative O-,
0.9 g N/m2/yr
5 to 15 g
NE: Harvard
Tonitto et

red pine
A-, and B-horizons C

N/m2/yr
Forest
al. (2014)


stocks of 211 g C/m2




ACAD = Acadia National Park; BAI = basal area increment; BBW = Bear Brook Watershed; C = carbon; Ca = calcium;
DIN = dissolved inorganic nitrogen; EB = East Bear Brook; g = gram; ha = hectare; HBEF = Hubbard Brook Experimental Forest;
kg = kilogram; m = meter; Mg = magnesium; N = nitrogen; N20 = nitrous oxide; NA = not applicable; NE = northeastern;
NH4+ = ammonium; (NH4)2S04 = ammonium sulfate; N03" = nitrate; NPP = net primary production; P = phosphorus; S = sulfur;
SOC = soil organic carbon; W6 = Watershed 6; WB = West Bear Brook; yr = year; Zn = zinc.
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16.1.5. Recovery
Research to measure and predict recovery in the Northeast indicates that while emission
control legislation has resulted in reduced S and N emissions from some stationary and
mobile sources, recovery of sensitive northeastern watersheds, including those located at
HBEF, BBW, and ACAD, has been limited. From 2000 to 2010, lake NOs,
concentrations in New England and the Adirondack Mountains, which had no trend prior
to 2000, declined at a rate of-0.05 (j,eq/L/year (Strock et al.. 2014). Research by Warbv
et al. (2008) in the Northeast suggested a nascent recovery of soil acid-base chemistry at
some locations where red spruce stands were sampled. In BBW, SanClements et al.
(2010) found little evidence of continued depletion or recovery of soil exchangeable base
cations. Mitchell and Likens (2011) observed that with declining inputs of atmospheric S,
the higher outputs of SO42 in drainage waters relative to precipitation inputs appear to be
driven by the S stored in the soil.
To simulate the future responses of soil and surface water chemistry to potential future
emissions controls at the reference watershed at HBEF, (Gbondo-Tugbawa and Driscoll.
2002) used the PnET-BGC (Photosynthesis and Evapotranspiration-Biogeochemical)
model. Simulation results suggested that emissions controls required by the CAAA will
likely not result in substantial future changes in critical indicators, such as soil base
saturation, soil solution Ca:Al, or stream ANC and Al concentrations. Tominaga et al.
(2010) assessed model outputs of four ecosystem models by systematically applying each
of the models to data from the HBEF. They predicted that current legislated emissions
reductions (13% reduction of SO42 by 2015) would result in limited recovery and that
maximum feasible technology reductions (78% reduction of SO42 by 2015) would result
in a more rapid and greater extent of chemical recovery.
The recovery of sensitive northeastern watersheds appears to be influenced by climate
change, which has become increasingly important to consider in understanding the
potential for recovery from decades of S and N deposition. For example, climatic change,
which increases annual stream discharge and watershed wetness, will potentially increase
S042 mobilization and hence may slow the recovery of both aquatic and terrestrial
ecosystems from acidification (Mitchell and Likens. 2011). Predictive modeling of
recovery should, therefore, consider climate change. Two modeling studies reported in
this case study are noteworthy for improving scientific understanding of recovery in the
Northeast. To test assumptions about interactions that are influenced by climate change,
Campbell et al. (2009) applied the PnET-BGC model in a northern hardwood forest
ecosystem at HBEF. Results suggested an increase in primary productivity due to a
longer growing season in the future, an increase in NO3 leaching due to enhanced net
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mineralization and nitrification, and a slight decrease in mineral weathering. In addition,
this warmer climate has contributed to significant declines in snow depth, snow-water
equivalents, and snow cover duration. This change has important implications for forest
ecosystem processes, such as tree phenology and growth, hydrological flow paths during
winter and spring, and soil biogeochemistry. Expected future climate change was also
simulated using ForSAFE-VEG (Soil Acidification in Forest Ecosystems model) and
found to cause an increase in soil base saturation, especially at BBW (Phclan ct al..
2016). The authors observed that climate had a lesser effect on simulated soil acid-base
chemistry at HBEF, likely due to the overwhelming influences of high S and N
deposition. They suggested that climate futures predicted by the Intergovernmental Panel
on Climate Change (IPCC) of increased temperature and precipitation will change plant
communities and N enrichment, counteracting the acidifying impacts of S and N
deposition on soil acid-base chemistry.
Finally, increasing attention is being given to developing and implementing practices to
restore forest productivity lost to impacts from acidification. Practices such as calcium
addition may increase recovery rates in northeastern forests. For example, Battles et al.
(2014) added calcium silicate to the HBEF paired watershed to determine whether Ca lost
from leaching due to acidification can be restored. Their results suggested tree biomass
recovery and reversal of forest decline through greater aboveground net primary
production and increased photosynthetic surface area. CO2 fertilization associated with
climate change is also being evaluated for its role in helping mitigate N loss in
watersheds (Pourmokhtarian et al.. 2012).
16.2. Adirondack Case Study: Adirondack Region of New York
16.2.1. Background
This case study is meant to identify the effects of nitrogen (N) and sulfur (S) in the
Adirondack region and the Adirondack state park in New York. This case study identifies
current acidification and nutrient status and empirical and modeled critical loads (CLs). It
is a supplement to the case study of acidification in the Adirondack region of New York
included in the 2008 NOx-SOx ISA [Section 3.2.2.4 of 2008 ISA (U.S. EPA. 2008a)l.
Further information about the effects in the Adirondack region can be found in
Appendix 4. Appendix 5. Appendix 7. and Appendix 8.
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16.2.1.1. Description of Case Study Region
1	The Adirondack Mountains are in northeastern New York State and are densely forested,
2	have abundant surface waters, and have 46 peaks that extend up to 1,600 m in elevation
3	(Figure 16-9). This area includes the headlands of five major drainage basins: Lake
4	Champlain and the Hudson, Black, St. Lawrence, and Mohawk rivers. There are more
5	than 2,800 lakes and ponds, and more than 1,500 miles of rivers that are fed by an
6	estimated 30,000 miles of brooks and streams. The Adirondack Park has long been a
7	nationally important recreation area for fishing, hiking, boating, and other outdoor
8	activities.
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Adirondack Park USGS National Land Cover
J Open Water (343,700 ac)
] Low Intensity Residential (6,100 ac)
3 High Intensity Residential (4,500 ac)
|Commerclal/lndustrial/Transportation (15,600 ac) |
] Bare Rock/Sand/Clay (500 ac)
| Quarries/Strip Mines/Gravel Pits (3,100 ac) [
| Transitional (Barren) (7,900 ac)
| Deciduous Forest (3,535,400 ac)
| Evergreen Forest (862,900 ac)
| Mixed Forest (606,700 ac)
3 Pasture/Hay (66,600 ac)
| Row Crops (45,900 ac)
| Urban/Recreational Grasses (1,800 ac)
3 Woody Wetlands (295.500 ac)
3 Emergent Herbaceous Wetlands (25,200 ac)
(TTT
A
rp-rn
0 4 8 Miles
Adirondack
Park Agency
April 2009
This map should
not be used for legal
jurisdictional determinations.
Figure 18-9 Map of Adirondack park land cover.
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The Adirondacks, particularly the southwestern Adirondacks, are sensitive to acidifying
deposition because they receive high precipitation, have shallow base-poor soils, and are
underlain by igneous bedrock with low weathering rates. The Adirondacks are among the
most severely acid-affected regions in North America (Driscoll et al.. 2003b; Landers et
al.. 1988) and have long been used as an indicator of the response of forest and aquatic
ecosystems to U.S. policy on atmospheric emissions of SO2 and NOx (NAPAP. 2011;
U.S. EPA. 1995a).
16.2.2. Deposition
Characteristics of nitrogen and sulfur deposition affecting the Adirondack region are
shown in Figure 16-10 and Figure 16-11. Figure 16-10A and Figure 16-11A show 3-year
average total deposition ofN and S for 2011-2013; Figure 16-10B shows the partitioning
between oxidized and reduced N; Figure 16-1 IB shows the 25-year-long time series for
wet deposition for NO3 , NH4+, SO42 and H+ obtained at the NADP/NTN (National
Atmospheric Deposition Program/National Trends Network) monitoring site at Whiteface
Mountain, NY (NY98). Surrounding areas in Vermont and New Hampshire and inserts
showing the coterminous U.S. (CONUS) are shown to place the depositional
environment in context. See Appendix 2.4. Appendix 2.5. and Appendix 2Jl for more
information on deposition in the U.S. Other maps showing the contributions of individual
species to dry and/or wet deposition are given in Appendix 2.7.
Data shown in the map Figure 16-10 and Figure 16-11A were obtained from the hybrid
modeling/data fusion product, TDEP (Total Deposition,
http: //nadp. sws .uiuc.edu/committees/tdep/tdepmaps/ and described in Annex 2,
Section 2.8). The time series of wet deposition is taken directly from data on the
NADP/NTN (Figure 16-1 IB). This was done to track changes in deposition since the
passage of the Clean Air Act Amendment (CAAA) because the Community Multiscale
Air Quality (CMAQ) model simulations used in TDEP extend back only to 2000.
Comparison of Figure 16-1 OA and Figure 16-11A indicates that the general pattern of
deposition of N and S is broadly similar; deposition estimates tend to be uniform
throughout the Adirondack Park. In addition, the area surrounding the park shows a high
degree of regional homogeneity but with higher values to the Southwest. Deposition of N
is not as high as in many areas of the central U.S. Deposition of S is considerably lower
than along the Ohio River Valley.
Figure 16-10B shows that the deposition of nitrogen is estimated to be mostly in oxidized
form throughout the entire park. In Figure 16-1 IB. wet deposition of all species shows
that downward trends in NO3 . NH4+, SO42 . and H+ are consistently found over the past
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1	25 years, although the rate of decrease was variable, hi general, wet deposition typically
2	exceeds dry deposition of N and S in this area.
B
ha = hectare; kg = kilogram; N = nitrogen.
Figure 16-10 Total nitrogen deposition (A) and percentage of oxidized nitrogen
deposition (B) for the Adirondack case study area estimated by
the National Atmospheric Deposition Program Total Deposition
Science committee.
O Monitor NY #98
0	Monitor NY #20
0 Monitor Locations
1	I Adirondack Park Boundary
0 30 60
Kilomelef s
O Monitor NY #98
0	Monitor NY #20
Q Monitor Locations
1	1 Adirondack Park Boundary
N Deposition (kg-N/ha)
% Oxidized N Deposition
WW
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A
B
Annual Wet Deposition and 3-Year Moving Average at
Site NY98:1990 - 2014
600
"i_ 500
>
400			
o
o
1988	1992	1996	2000	2004	2008	2012	2016
Year
•	nh4+
•	no3-
•	S04z-
•	H+ Lab
H+ = hydrogen ion; ha = hectare; kg = kilogram; mol = mole; NH/ = ammonium; N03 = nitrate; S = sulfur; S042" = sulfate; yr = year.
Figure 16-11 Total sulfur deposition (A) for the Adirondack case study area
estimated by the National Atmospheric Deposition Program Total
Deposition Science committee. Time series of wet deposition (B)
from the National Atmospheric Deposition Program/National
Trends Network Whiteface Mountain, NY
0	30 60
Kilometers
O Monitor NY #98
O Monitor NY #20
J Monitor Locations
II Adirondack Park Boundary
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16.2.3.	Critical Loads and Other Dose-Response Relationships
16.2.3.1.	Terrestrial
1	This section presents post-2008 Adirondack area findings on the dose-response
2	relationships of N and S to terrestrial ecology, as well as the critical N and S loads for
3	maintaining ecosystem health. Additional relevant information for the Adirondack region
4	is summarized. The section presents findings on both empirical research and modeling
5	analyses.
16.2.3.1.1.	Empirical and Modeling Studies
6	Post-2000 findings from empirical and modeling studies of terrestrial dose-response
7	relationships and critical loads are summarized in this section. Table 16-14 summarizes
8	the body of empirical and modeling research identified.
Table 16-14 Terrestrial empirical and modeling research on the response of
nitrogen and sulfur deposition for the Adirondack Mountains.



Deposition/






Addition



Variable
Species
Response
(kg N/ha/yr)
Years
Site
Reference
Community
Redback
Increasing trends in
14.57 to 19.7
2009
12 upland
Beier et al. (2012)
richness and
salamanders,
snail community
kg/ha/yr as wet

hardwood

abundance,
calciphilic
richness and
NOs"; 17.44 to

forests

live biomass
species of
abundance, live
29.09 kg/ha/yr




snails
biomass of
as wet SO42",





salamanders, with
modeled (Ito et





increasing soil Ca.
al.. 2002)






1990-1999



Canopy tree
Sugar maple
Sugar maple basal
14.57 to 19.7
2009
12 upland
Beier et al. (2012)
basal area

area was positively
kg/ha/yr as wet

hardwood



correlated to forest
NO3-; 17.44 to

forests



floor and mineral soil
29.09 kg/ha/yr





(B-horizon)
as wet SO42",



modeled (Ito et
al.. 2002)
1990-1999
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Table 16-14 (Continued): Terrestrial empirical and modeling research on the
response of nitrogen and sulfur deposition for the
Adirondack Mountains.
Deposition/
Addition
Variable Species	Response	(kg N/ha/yr) Years	Site	Reference
Tree basal
Sugar maple,
Sugar maple basal
NA
2008 Ha De-
McEathron et al.
area
black cherry,
area was positively

Ron-Dah
(2013)

American
correlated with

Wilderness


beech, red
mineral soil pH, and

Area in


maple, and
yellow birch basal

Adirondack


yellow birch
area was positively

Mountains



correlated with





mineral soil





exchangeable Ca.





Sugar maple basal





area was also





negatively correlated





with stream water





DOC.



Tree basal
Sugar maple,
Relative basal areas
NA
2004-2005 Adirondack
Paae and Mitchell
area
American
of sugar maple and

Park
(2008)

beech,
American basswood




American
were positively




basswood,
correlated with




and white
mineral soil




ash
exchangeable Ca;





American beech was





negatively correlated;





white ash was not





correlated.



Sugar maple
Sugar maple
Plots with lower soil
750 to
2009 Adirondack
Sullivan et al.
regeneration

base saturation did
1,120 eq/ha/yr
Park
(2013)


not have sugar maple
as N + S




regeneration (these
(NADP wet,




same plots also
CASTNET dry)




received higher N and





S deposition levels);





proportion of sugar





maple seedlings





dropped substantially





at base saturation





levels less than 20%.



Canopy
Sugar maple
Canopy vigor was
750 to
2009 Adirondack
Sullivan et al.
vigor and

positively correlated
1,120 eq/ha/yr
Park
(2013)
tree growth

with soil pH and
as N + S




exchangeable Ca and
(NADP wet,




Mg. Mean growth
CASTNET dry)


rates (BAI) were
positively correlated
with exchangeable
Ca and base
saturation at the
watershed level.
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Table 16-14 (Continued): Terrestrial empirical and modeling research on the
response of nitrogen and sulfur deposition for the
Adirondack Mountains.
Deposition/
Addition
Variable Species	Response	(kg N/ha/yr) Years	Site	Reference
Tree basal All major Significant positive Gradient of 1984-2004 Adirondack Bedison and McNeil
area and trees	overall effect of N 3.5—7 kg	Park	(2009)
woody	deposition on tree N/ha/yr
biomass	growth from
increments	1984-2004, but
positive growth
effects only for red
maple, balsam fir,
and red spruce at the
species level.
NO3"	NA	PnET-BGC Model. Not specified 1999-2099 NE	Campbell et al.
leaching	Increase due to	(2009)
enhanced net
mineralization and
nitrification
Mineral NA	PnET-BGC Model. Not specified 1999-2099 NE	Campbell et al.
weathering	Slight decrease due	(2009)
to reduced simulated
soil moisture
(negative effect) and
increased
temperature (positive
effect)
BBW = Bear Brook Watershed; BC = base cation; C = carbon; Ca = calcium; ForSAFE-VEG = Soil Acidification in Forest
Ecosystems; ha = hectare; HBEF = Hubbard Brook Experimental Forest; IPCC = Intergovernmental Panel on Climate Change;
kg = kilogram; Mg = magnesium; N = nitrogen; NA = not applicable; ND = no data; NE = northeastern; (NH4)2S04 = ammonium
sulfate; N03" = nitrate; PnET-BGC = Photosynthesis and Evapotranspiration-Biogeochemical; S = sulfur; WB = West Bear Brook;
yr = year; Zn = zinc.
16.2.3.1.1.1. Critical Loads
1	Target loads of S deposition were calculated for 44 watersheds and extrapolated to
2	1,320 acid-sensitive watersheds in the Adirondacks using MAGIC and different soil
3	chemical indicator and thresholds of base saturation (5 and 10%), soil solution Bc:Al, and
4	Ca:Al [1.0 and 10.0 (Sullivan etal.. 201 la) I. In acomparison of target loads (out to
5	Years 2050 and 2100) and the 2002 deposition, only 11.6 to 13.5% of the watersheds
6	were simulated to be in exceedance of target loads to protect soil base saturation to 5%.
7	For target loads to protect soils to a base saturation of 10%, 79.7 to 87.5% of the
8	watersheds were in exceedance. For soil solution Bc:Al, 7.8 and 98.1% of watersheds
9	were exceeded by the 2002 deposition for target loads to protect soil solution ratios of 1.0
10	and 10.0, respectively. For soil solution Ca:Al, 44.1 to 58.2% of watersheds experienced
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exceedances of target loads associated with a soil solution ratio of 1.0, and 98.1% of
watersheds with target loads to protect Ca:Al ratios of 10.0 were exceeded. Further
investigations revealed that 58.2, 85.7, and 93.6% of watersheds could not obtain
threshold values of 10% base saturation, Bc:Al = 10, or Ca:Al = 10, respectively, even if
acidic deposition was held at zero, thereby demonstrating that these chemical indicator
threshold values were not useful for target load calculations using MAGIC in the
Adirondack Mountains.
16.2.3.2. Aquatic
This section presents research findings at Adirondack Mountains on the dose-response
relationships of N and S to aquatic ecology, as well as the critical N and S loads for
maintaining ecosystem health. Additional relevant information for the Adirondack region
is summarized. Both empirical and modeling studies of aquatic dose-response
relationships and critical loads are included in this section.
16.2.3.2.1. Empirical Studies
Studies in the Adirondack Mountains reviewed in the 2008 ISA demonstrated the effect
of acidification on fish species richness. Of the 53 fish species recorded in Adirondack
lakes, about half (26 species) were absent from lakes with pH below 6.0. Those
26 species included important recreational species plus ecologically important minnows
that serve as forage for sport fish (Baker etal.. 1990b). There is often a positive
relationship between pH and number of fish species, at least for pH values between about
5.0 and 6.5, or ANC values between about 0 and 50 to 100 |icq/L (Cosby et al.. 2006;
Sullivan et al.. 2006a; Driscoll et al.. 2003b; Bulger etal.. 1999).
As summarized in the 2008 ISA, lakes and streams having an ANC < 0 (ieq/L generally
do not support fish (Figure 16-12). The analysis shown in this figure suggests that there
could be a loss of fish species with decreases in ANC below a threshold of approximately
50 to 100 (ieq/L (Sullivan et al.. 2006a').
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water's pH declines from pH 7.0 to 4.2. A loss of about 12 species occurred in streams
that had a pH between 7 and 4.2. Regression across all 36 streams showed a loss of
4.6 species per unit pH decrease (Figure 16-13). Inorganic A1 toxicity was likely the main
cause of the loss of macroinvertebrates. The A1 concentration is strongly correlated with
surface water pH (as pH decreases, the solubility of inorganic A1 increases) and acid-base
balance as measured by BCS.
Studies in the Adirondacks have considered effects of species assemblages associated
with acid-impacted water bodies. Nierzwicki-Bauer et al. (2010) presented baseline
midsummer chemistry and community composition of phytoplankton, bacteria, rotifers,
crustaceans, macrophytes, and fish in 30 lakes of the southwestern Adirondacks
(Figure 16-14). Species richness in the lakes was correlated with acid-base chemistry at
all trophic levels except bacteria. The decrease in number of taxa per unit pH was similar
among groups and ranged from 1.75 (crustaceans) to 3.96 (macrophytes).
Percent et al. (2008) assessed bacterioplankton community diversity and structure in
18 Adirondack lakes across a range of acid-base chemistry using sequencing and
amplified ribosomal DNA restriction analysis of constructed rRNA gene libraries. Based
on principal components analysis, pH was positively correlated with bacterioplankton
community richness and diversity. The richness of several bacterial classes, including
Alphaproteobacteria, was directly correlated with pH. However, other environmental
factors, such as lake depth, hydraulic retention time, dissolved inorganic C, and nonlabile
(organically bound) monomeric Al, were also important in explaining bacterioplankton
community richness and diversity, suggesting that acidity is only one factor that controls
community composition.
Post-2008 findings from empirical studies of aquatic dose-response relationships and
critical loads are summarized in this section. Table 16-15 compiles the body of empirical
research identified.
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u. 30
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Bactcrioplankton
~ ~
y=2.4972x - 0.1353
~ ~ ~
R'=0.049
~
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¦
y=1.4259x-1.9141 —
R'-0.122
•125 4 75 525 5 75 6.25 6,75 7 25 7 75
Rotifers
y^3,S603x - 12.249
R*-0.6863
•<25 4 75 5.25 5.75 6,25 675 7.25
~ ~
Macrophytes
~
y-3.957x • 9.3734
~ ~
R-—0.2128
~ ~ ~
—— i ¦ *—

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~ ~
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Phytoplanklon
y=3.9744x - 9.0623
~ ~
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42)	4 75	5 25	575	625	6,75	7 25
Crustaceans
30
25
20
15 |
10
5 1-
y-l,7529x- 3.4791
R-0.5873
4,25 4.75 5.25 5.75 6.25 6.75 7.25
Fish

y=3.7218x - 17-641
R!=0.5316
~
~
~
A r~
	—
425 4.75 5.25 5,75 625 6.75 725
Lake pH
Notes: they-axis denotes number of groups (bacteria), taxa (phytoplankton) or species (rotifers, crustaceans, macrophytes, fish).
The x-axis denotes the pH range that occurred in the 30 study lakes. In the bacterioplankton graph, two regressions were run on the
basis of genera (diamonds) and classes (squares).
Source: Nierzwicki-Bauer et ai. (2010't.
Figure 16-14 Species richness of biotic groups in 30 Adirondack study lakes
relative to midsummer epilimnetic pH during sample years.
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Table 16-15 Aquatic empirical research on the response of nitrogen and sulfur
deposition for the Adirondacks.
Study
Time
Period
Focus
Notable Results
Changes to Acidity
Strock et al. 1980-2010 Trends in recovery for SO42" concentration decreased slightly faster in the
(2014)	lakes in the Northeast, 2000s than the 1990s—possibly the result of a decline
including 43 sites in the in S emissions that was twice as large in the 2000s
Adirondack Mountains. than it was in the previous decade. NO3" concentration
did not significantly fall in the 1990s (a decade which
had more modest N emissions reductions) but did
significantly fall in the 2000s, which saw large N
emission reductions.
Mitchell et al. 1984-2010 S deposition response in SO42" concentration has declined significantly along
(2013)	16 Adirondack long-term with total S deposition. However, there is a discrepancy
monitoring Lakes.	in the balance of deposition inputs and discharge
outputs. The authors conclude that internal S supplies
are likely contributing to high sulfur concentrations and
the slow recovery of lakes in the region, and that
internal S will become a more important factor as
emissions decline.
Waller et al. 1991-2007 The effectiveness of the
(2012)	Acid Rain Program and
Nitrogen Budget Program
in the 1990 CAAA in
reducing Adirondack lake
acidity for 42 lakes.
Sulfate concentration declined by 23.47% over the
study period, but nitrate concentration did not
significantly decline. ANC has increased over the study
period, with the number of lakes with ANC <0 dropping
by 46%. ANCg was higher than ANCcaic, suggesting
contribution to acidity from naturally occurring organic
acids that are not factored into ANCcaic.
Baron et al.	1997-2006 Empirical N loads in	Minimally disturbed Northeast lakes were estimated to
(2011b)	216 northeastern U.S.	have thresholds of 3.5-6.0 kg N/ha/yr for avoidance of
lakes including	acidic episodes.
Adirondack lakes.
Lawrence et al. 1985-2008 Focuses on	The pH in streams with lower ANC was more
(2011)	12 Adirondack streams responsive to the 50% reduction in S deposition over
rather than lakes.	the study period than pH in streams with high ANC.
Authors noted "a more muted recovery response" in
streams compared to what has been reported in lakes.
Inamdar and 2002-2004 Point Peter Brook	Results suggested that groundwater sources have
Mitchell (2008)	watershed in Western NY	greater impact on changes to sulfate flux in observed
and fluxes of sulfate	lakes than does deposition through precipitation,
following storms.
Warbv et al. 1986-2001 Speciation and
(2008)	concentration of Al in
113 northeastern U.S.
lakes (including the
Adirondacks) and the
effect of decreased acid
deposition.
The Adirondack lakes had the largest decline in Al
concentration of all lakes studied (Adirondacks:
2.01 pmol/L; all: 0.44). For reference, the decline in
SO42" deposition in the Adirondacks was about the
same as the decline in the region as a whole.
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Table 16-15 (Continued): Aquatic empirical research on the response of nitrogen
and sulfur deposition for the Adirondacks.
Study
Time
Period
Focus
Notable Results
Changes to Biota
Sutherland et al.
(2015)
1984-2012
Brooktrout Lake
acidification and biotic
recovery.
Likely as a result of the 1990 Clean Air Act
Amendments, SO42" concentrations have dropped 56%
or 2 pequiv/L/yr since the 1980s. Plankton species
richness has also blossomed, though crustacean
species richness has had less impressive growth
(possibly due to pressures exerted by the predator-free
glassworm, Chaoborus). Brook trout (Salvelinus
fontinalis) are now successfully surviving and
reproducing after their reintroduction in 2005.
Josephson et al.
(2014)
1960-2011
Honnedaga Lake
watershed recovery.
There has been significant decline in SO42" and NO3"
since 2001, when consistent measurements in surface
water chemistry began. There were no notable trends
in zooplankton species richness as lake acidity dropped
over the observation period, but there was a significant
resurgence of adult brooktrout in the 2000s compared
to catches in the 1970s (a nearly sixfold increase). The
authors noted the relationship between acidity and
aluminum toxicity for fish. Recovery of adult trout could
be improved by lower acidity tributaries that would
make a friendlier environment for young-of-year trout.
Yu et al. (2011)
2003-2004
Focused on Hg
concentrations in biota of
44 lakes (not NOxSOx but
relevant to acidity).
Lower ANC in surface water was correlated with higher
Hg concentrations in the tissues offish, loons, and
zooplankton.
Nierzwicki-Bauer
et al. (2010)
1994-2006
Correlation between
species richness and
acid-base chemistry of
30 Adirondack lakes.
For each unit of pH (toward increased acidity), species
richness dropped from 1.4-1.8 (bacterial classes,
crustaceans) to above 3.5 (fish, rotifers, phytoplankton,
macrophytes). Authors classified species into a system
of brackets describing acid sensitivity or tolerance, with
most fish species being considered acid sensitive.
Baldiao et al.
(2009)
2003-2005
Macroinvertebrate
response to acidification
of 36 streams in the
Adirondack region.
Species richness was measured by acid
bioassessment profile (acidBAP) scores. In about half
of streams, acidBAP scores were moderately or
severely impacted (44%) from low pH—this increased
to about two-thirds of streams when including slight
impacts (69%).
Percent et al.
(2008)
2002-2002
Acidity and bacterial
communities in
18 Adirondack lakes.
Authors reported a significant decline in bacterial
species richness and species diversity with increasing
acidity. However, they did not find the overall bacterial
community composition to be significantly affected.
Notes: ANC = acid neutralizing capacity.
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16.2.3.2.2.
Modeling Studies
Post-2008 findings from modeling studies of aquatic dose-response relationships and
critical loads are summarized in this section. Table 16-16 compiles the body of research
identified.
Table 16-16 Critical and target load and exceedance modeling studies in
Adirondack Mountains.
Reference	Model	Focus	Notable Results
Sullivan et al. (2012a) MAGIC and	TL for lakes in 2050 To achieve ANC = 50 peq/L in 2100, about
regional	and 2100	30% of lakes had simulated TL of S
extrapolation of	deposition <500 eq/ha/yr and about 600 lakes
model	were in exceedance.
Zhou et al. (2015b)	PnET-BGC	TL link to fish and The magnitude of simulated historical
zooplankton richness acidification represented by ANC loss ranged
from about 26 to 100 peq/L. The amount of
historical acidification of the lakes was related
to the total ambient deposition of
SO42" + NO3", the Ca weathering rate, and
the simulated preindustrial ANC in the
Year 1850. Adirondack lakes have lost fish
and total zooplankton species richness
beginning with the onset of acidic deposition.
Modeling results suggested that complete
recovery to preindustrial conditions may not
be possible for most acidified lake
ecosystems.
Zhou et al. (2015c)	PnET-BGC	Effects of biophysical Model simulations suggested that future
factors on the TL in decreases in SO42" deposition would be more
lakes	effective in increasing the lake water ANC
than equivalent decreases in NO3"
deposition. The difference was a factor of 4.6
during the period 2040 to 2050, but
decreased to a factor of 2 by the Year 2200.
Lakes that had longer hydrologic residence
exhibited less historical acidification and
therefore could achieve a higher ANC in
response to chemical recovery compared with
lakes that had short hydrologic residence
times.
Fakhraei et al. (2014) PnET-BGC	TMDLs for 128 acid- Model simulations suggested that an S TL
impaired lakes	equal to 79 eq S/ha/yr (representing a 60%
decrease from ambient deposition) would
lead to ANC recovery at a rate of
0.18 [jeq/L/yr through 2050, with reduced rate
of recovery thereafter.
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Table 16-16 (Continued): Critical and target load and exceedance modeling
studies in Adirondack Mountains.
Reference	Model	Focus	Notable Results
NAPAP (2011)	SSWC	Combined deposition To achieve ANC = 50 peq/L on average,
load of S and N to critical load of sulfur and nitrogen for lakes in
which a stream and the Adirondack Mountains is 1,620 eq/ha/yr
its watershed could
be subjected and still
have a surface water
concentration ANC of
50 peq/L on an
annual basis
Sullivan (2015)	MAGIC and Development and	To achieve ANC values of 50 and 20 peq/L in
regional extra- application of tools to	the Year 2050 and 2100, the TL to protect
polation of model document and	against acidification of surface waters was
quantify TL and their	exceeded throughout the Adirondack
exceedances	Mountains.
TL = target load; CL = critical load; ForSAFE-VEG = Soil Acidification in Forest Ecosystems; TMDL = Total Maximum Daily Load;
MAGIC = Model of Acidification and Groundwater in Catchments; N = nitrogen; S = sulfur; PnET-BGC = Photosynthesis and
Evapotranspiration-Biogeochemical; ANC = acid neutralizing capacity; SSWC = Steady-State Water Chemistry; VSD = Very Simple
Dynamic.
16.2.4. Long-Term Monitoring
This section summarizes research on the long-term effects of S and N deposition in the
Adirondacks.
16.2.4.1. Long-Term Monitoring of Acidification
Acidic deposition has been shown to be an important factor causing decreases throughout
much of the Northeast region in concentrations of exchangeable base cations in soils,
which were naturally low historically. Base saturation values less than 12% predominate
in the B-horizon in portions of the region where soil and surface water acidification from
acidic deposition have been most pronounced (Sullivan et al.. 2006a; Bailey et al.. 2004;
David and Lawrence. 1996).
At the time of the 2008 ISA, the region of the U.S. most thoroughly studied to ascertain
changes in acid-base surface water chemistry overtime was the Adirondack Mountains.
This has not changed. Several new studies are highlighted here that focused on
Adirondack surface waters, three on lakes and one on streams. The Adirondack Lakes
Survey Corporation (ALSC) has been monitoring Adirondack lakes for about 30 years.
This work has mainly focused on acid-base chemistry, but has also involved some fish
monitoring and measurement of parameters relevant to nutrient enrichment. Monitoring
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data have shown some chemical recovery from lake acidification, reflected in increased
pH and ANC and decreased inorganic A1 concentrations.
A mass balance study by Mitchell et al. (2011) for 15 watersheds located in the
northeastern U.S. and southeastern Canada suggested substantial sources of SO42 in
watershed soils. The internal S sources were attributed mainly to mineralization of S
stored in soils in response to decades of atmospheric S deposition. Mitchell et al. (2013)
studied 16 of the original Adirondack Long-Term Monitoring lakes that were monitored
between 1984 and 2010. Total S deposition significantly declined in all of the study lake
watersheds during the monitoring period. Correspondingly, significant decreases were
observed overtime (-2.14 |iinol/L/vcar) in lake SO42 concentrations. The authors
observed discrepancies in the calculated S mass balances that were associated with
discharge. These results suggested that internal S sources have become more important to
the watershed S budget as atmospheric S deposition has decreased. The watershed supply
of S042 was lower in those watersheds that contained lakes that had lower ANC. Limited
contributions from internal sources of SO42 to the low-ANC lakes will allow ANC
recovery. Mitchell et al. (2013) further concluded that the effects of future increases in
precipitation might become more important in regulating the amount of SO42 mobilized
from internal watershed sources.
Surface water ANC and pH recovery has been documented by Lawrence et al. (2013) and
Lawrence et al. (2011) in studies of Adirondack streams and lakes. Lawrence et al.
(2011) reported results of Adirondack Mountain stream monitoring from the early 1980s
to 2008. During that period, there was an approximate 50% reduction in atmospheric
deposition of S in the Adirondack region. On average, the pH increased by only 0.28 pH
units and ANC by 13 (j,eq/L in 12 streams that were monitored over a period of 23 years.
Larger increases in ANC and pH occurred in the monitored streams that had lower initial
ANC. In the north tributary of Buck Creek that had a relatively high DOC concentration,
the S042 concentration decreased between 1999 and 2008 at a rate of 2 |imol/L/ycar. In
the adjacent south tributary, which had a lower DOC concentration, the decrease in SO42
concentration was substantially smaller, only 0.73 |imol/L/ycar. Driscoll et al. (2016)
observed increases in ANC and pH and marked decreases in dissolved inorganic Al in 45
of 48 study lakes in the Adirondack region from 1982 to 2015, corresponding to
decreases in acidifying deposition.
Michelenaetal. (2016) reported changes in the water chemistry of 30 Adirondack lakes,
in response to reductions in acidic deposition from 1994 to 2012. The water quality of the
study lakes generally improved during the study period, but the responses were sporadic
and complex. Lake pH values increased until about 2002 and then fluctuated. Inorganic
Al concentrations generally decreased throughout the period of record. During the early
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years of monitoring, the average pH increased dramatically from about 5.5 to 6.0. This
increase was followed by a period of reacidification for about 5 years, followed by
another period of increased pH. This 5-year cycle was then repeated to a muted extent,
even though acidic deposition continued to decline.
Chemical recovery from surface water acidification, and associated CL exceedance, in
the Adirondack Mountains has been accompanied by increasing concentrations of DOC
and organic acids, which have complicated and restricted recovery from acidification.
Lawrence et al. (2013) reported the chemistry of 42 Adirondack lakes from samples
collected between 1994 and 2011. They evaluated long-term changes in DOC and BCS
(which reflects the calculated ANC and includes an adjustment for strong organic acid
anions). Increases in DOC concentration during the study period contributed to organic
complexation of Al. This reduced the fraction of monomeric Al that was in the inorganic
(and potentially toxic) form from 57% in 1994 to 23% in 2011. Thus, the higher DOC
found during the latter sampling period increased the organic acidity of the lake water
and limited ANC recovery, but at the same time it decreased the toxicity of dissolved Al
to aquatic biota. Similarly, Strock et al. (2014) reported trends in recovery from
acidification of northeastern U.S. surface waters associated with Al chemistry. They
analyzed recent trends in lake chemistry using long-term data from lakes in the
Adirondack Mountains and New England. During the 2000s, the wet deposition of NO3
declined more than 50%. The lake NOs concentration, which showed no trend prior to
2000, declined subsequent to 2000 at a rate of-0.05 (j,eq/L/year. There was a shift to
nontoxic (organic) forms of Al. Despite the Al recovery, both the ANC and pH in the
study lakes failed to show evidence of appreciable chemical recovery; rather, the lakes
continued to exhibit variable trends in ANC and pH.
Waller et al. (2012) also evaluated the response of lake watersheds in the Adirondack
Mountains to changes in SO2 and NOx emissions and deposition, in this case using data
from the TIME monitoring program collected during the period 1991 to 2007. The
percentage of Adirondack lakes that were acidic decreased from an estimated 15.5 to
8.3%. Decreases in lake water SO42 concentrations, and to a lesser extent NO3
concentrations, generally were accompanied by increases in lake ANC. However, the
calculated ANC increased more than twice as much as the Gran titrated ANC. This
discrepancy is important for assessing lake recovery from acidification. It appeared to be
due mainly to increases in DOC (and associated organic anions) that accompanied
decreases in concentrations of the strong mineral acid anions, SO42 and NO3 . The
difference between calculated ANC and Gran titrated ANC was generally consistent with
trends of increasing DOC. Waller et al. (2012) concluded that assessments of surface
water recovery from previous acidification should consider differences in DOC
concentration and in the manner in which ANC is calculated and quantified.
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Interpretation of long-term trends in Adirondack surface water chemistry, as summarized
above, has also been augmented by results of repeated surveys of the chemistry of lakes
initially surveyed several decades previously. Warbv et al. (2008) resurveyed 113 lakes
throughout the Northeast in 2001 that had previously been sampled by the U.S. EPA in
1986 to assess chemical recovery from lake acidification. Decreases in total Al and
inorganic Al concentrations were widespread and largest in the Adirondack Mountains
(-2.59 to -4.60 (imol/L). In 2001, only 7 study lakes (representing 130 lakes in the
population) had inorganic Al concentrations above a toxic limit of 2 (imol/L, compared
with 20 sampled lakes (representing 449 lakes in the population) in 1986. Thus, it was
estimated that in 2001 more than 300 northeastern lakes no longer had summer inorganic
Al concentrations at levels considered harmful to aquatic biota.
16.2.4.2. Recovery
Biological recovery from past acidification is a process affected by chemical, climatic,
biological, and hydrologic influences overtime. It may, under certain conditions, follow
chemical recovery of such water quality constituents as pH, ANC, and the concentrations
of SO42 , NO3 . inorganic Al, and DOC. Both chemical and biological recovery can, and
often does, lag behind changes in the levels of S and/or N emissions and deposition
because of chemical, hydrological, and biological processes and constraints. Studies at
Honnedaga Lake and Brooktrout Lake in the Adirondacks show evidence for return of
biota to levels approaching preacidification levels (Sutherland et al.. 2015; Josephson et
al.. 2014). In Brooktrout Lake, biological recovery of the food web structure has begun,
in part, due to reintroduction and re-establishment of brook trout in the lake. However,
ongoing biological recovery cannot necessarily be expected to conclude with the return of
the biological community to preacidification conditions. The reason is due to factors such
as differences in the rate of recovery of aquatic organisms, permanent loss of some
acid-sensitive species, and irreversible chemical and physical alterations to aquatic
environments during acidification (NAPAP. 2011).
Since the 2008 ISA, some paleolimnological studies have been conducted that
documented historical acidification and subsequent recovery (see Appendix 8). In one
study from the U.S., diatom shifts were linked to historical changes in pH at Brooktrout
Lake in the Adirondacks. Fragilariforma acidobiontica, a diatom that is often abundant
at pH <5.0, was present in lake sediments deposited since the 1950s, and shifts in
Mallomonas and Synura were also observed (Sutherland et al.. 2015). In this study,
phytoplankton and rotifer taxonomic richness showed substantial increases
(Figure 16-15) in association with pronounced decreases in lake SO42 . H+, and inorganic
Al concentrations. In contrast, species richness of crustaceans changed little. Sutherland
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et al. (2015) suggested that the absence of crustacean recovery may have been due to the
severity of the initial stress and continuing toxic pH conditions during some seasons,
and/or to a large population of predatory Chaoborus larvae present since the early 1980s
when fish were absent from the lake. In another study, Arseneau et al. (2016) concluded
that Adirondack lakes that were not previously acidified by acidic deposition will likely
not recover to predisturbance chrysophyte community structure because of the influence
of other stressors, including changes in climate.
Despite observed reductions in acidic deposition and improvement in water quality of
New York lakes, Baldigo et al. (2016) found no evidence of widespread or substantial
biological recovery of brook trout populations or broader fish communities in the
Adirondack Mountains. The study focused on 43 lakes sampled by the Adirondack Lakes
Survey Corporation during three time periods (1984-1987, 1994-2005, and 2008-2012).
Metrics reflecting fish species richness, abundance of fish species, and abundance of
brook trout did not change significantly over the 28-year period across the group of study
lakes despite a significant average ANC increase and a decrease in inorganic Al over
time. Fish species richness and catch of all fish species per net-night were positively
related to lake chemistry reflecting some limited degree of biological recovery. The
authors speculated that additional time may be needed for fish recolonization.
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160 -|
140
- 120
>•
1100
i
E
2." 80
ar
9/
| 60
Ł 40 -\
20
0
20
18
16
I"
^12
<
J10
O"
i 8
c
Ł 6
4
2
0
40
35
1 30
JC
u
= 25
.2
120
I 15
(a)
8 °
8
R2 = 0.88
(b)
(c)
| 10
o Mid-summer Epilimnetic S042' (peq I"1)
• Wet Deposition S042 (meq/m2/yr)
-¦O... ° § Q n
o o °	o
O o	o
o	° e;
® « a 8
® § 8 s
1 8 e
R2 = 0.78

~ Mid-summer Epilimnetic IMA (pM I1)
¦ Mid-summer Epilimnetic [H*] (peq I *)
R2 = 0.56
D o
D ~
I ¦
R2 = 0.55
¦ a
-s.t
~ .'«!¦ i J ¦ ' I j n
. R2 = 0.60
i '
A A
AAA
* i 4 4 R2 = 0.60
a Phytoplankton
* Total Plankton
84 86 88 90 92 94 96 98 00 02 04 06 08 10 12
Year
H' = hydrogen ion; IMA = inorganic monomeric aluminum; (jeq = microequivalent; |jM = micromolar; m = meter; S042" = sulfate;
yr = year.
Source: Sutherland et al. (2015).
Figure 16-15
(A) Midsummer sulfate concentration in the epilimnion of
Brooktrout Lake (o) and in annual wet deposition (•) at local
National Atmospheric Deposition Program/National Trends
Network Station NY52 from 1984-2012. (B) Midsummer
epilimnetic concentrations of inorganic monomeric aluminum (~)
and hydrogen ion (¦) in Brooktrout Lake from 1984-2012.
(C) Midsummer phytoplankton (A) and total plankton
(phytoplankton, rotifers, crustaceans) (A) species richness in
Brooktrout Lake from 1984-2012.
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16.3.
Southeastern Appalachia Case Study
16.3.1. Background
This case study focuses on the Great Smoky Mountain National Park and on several
closely located Class I areas in North Carolina (Joyce Kilmer Forest, Linville Gorge,
Shining Rock). These areas are representative of the southern Appalachian region and
also have been the focus of research on the ecological impacts of historical and current
NOx and SOx deposition. The case study summarizes literature published since 2008,
although older studies are included when relevant (e.g., conducted in Class I areas for the
purpose of investigating critical loads). The southern Appalachians are in the Eastern
Temperate Forests ecoregion, so critical loads for the broader Level I ecoregion are also
presented.
16.3.1.1. Description of Case Study Region
This case study focuses on the southern Appalachian Mountains, with special emphasis
on Great Smoky Mountains National Park (GRSM), located on the border between
eastern Tennessee and western North Carolina. The park consists of 2,100 km2 of
mountain ridges and valleys and ranges over an elevation of 267 to 2,025 m (Thornberrv-
Ehrlich. 2008). with peaks at the highest elevations in the eastern U.S. The wide range in
elevation, annual precipitation [140 cm in valleys and 215 cm on ridges and peaks,
(Thornbcrrv-Ehiiich. 2008)1. and temperature [growing season temperatures are 10-15°C
cooler on ridge tops and peaks than in valleys, (Lesser and Fridlev. 2016)1 within the
park has allowed distinct terrestrial and aquatic communities to form at different
elevations.
Geologically, the park is in the Blue Ridge physiographic province, where Cambrian
crystalline quartzite forms ridges, and Precambrian metamorphic rocks underlie the
valleys, particularly in the eastern part of the park. Carbonates and shales of the Valley
and Ridge province underlie valleys in the western part of the park where deposits of
dolomite provide substantial base cations to drainage water. The Anakeesta Formation
sits at the highest elevation of the Great Smoky Group, which forms the crest of the
mountains. The Anakeesta Formation contains iron sulfide minerals that contribute to soil
solution and surface water sulfate; it has outcropppings at Clingmans Dome, Newfound
Gap, and Chimney Tops, as well as at other high-elevation locations in the north-central
section of the park.
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Soils of the Blue Ridge tend to be thin and shallow, overlying steep terrain (Thornbcrrv-
Ehrlich. 2008). Park soils are highly weathered and, particularly on mountain tops and
ridges, have substantial soil S adsorption capacity and limited base cation supply
(Elwood. 1991). Linville Gorge Wilderness is also within the Blue Ridge Province, with
soils that formed with low Ca, Mg, and K concentrations, in which base cation
concentrations have further declined in response to acid deposition (Elliott et al.. 2013).
There are 3,200 km of streams within the park's boundaries (Nichols and Langdon.
2007). Streams in the park are first through sixth order, and their chemistry is strongly
influenced by geology and storm events, as well as by acidic deposition. In 2006,
Tennessee listed 12 streams (67 km stream length) within GRSM as impaired under the
Clean Water Act Section 303d, because mean stream pH was below 6.0 (Fakhraei et al..
2016). Streams in GRSM that have ANC between 50 and 200 |icq/L (41%) are mostly
influenced by limestone weathering and tend to be located in the western portion of the
park (Neff et al.. 2009). Most stream ANCs in the park (54.2%) are in the 0-50 j^ieq/L
range. Streams that have ANC near or below 0 j^icq/L (2.4%) are influenced by
weathering of the Anakeesta Formation and are mainly located in the north-central
portion of the park (Neff et al.. 2009; Flum and Nodvin. 1995). Within the park,
hindcasting by PnET-BGC model simulations suggested that ANC in park streams in
1850 ranged from 28 to 107 (j,eq/L, with lower ANC at higher elevations (Zhou et al..
2015a). In the broader southern Blue Ridge province, MAGIC model simulations
hindcasted to 1860 suggested that 66 modeled streams had preindustrial
ANC > 30 (j,eq/L; in 2005, 30% of the streams had ANC below that level (Sullivan et al..
2011c). Streams in the Blue Ridge region with ANC < 20 (j,eq/L were underlain by
siliceous bedrock, including sandstone and quartzite (Sullivan et al.. 2007b).
Storm events cause episodic acidification in streams, with decreases in pH of 0.5 to
1.6 pH units in streams with ANC of <20 (j,eq/L under high-flow conditions (Lawrence et
al.. 2015b; Devton et al.. 2009). Stream chemistry responses to storm events varies by
elevation within the park. At high elevations, stream pH decreased by as much
0.5-2.0 pH units during storm events in 1990-2008 (Neff et al.. 2009; Cooketal.. 1994).
and streams above 975 m in elevation had significantly lower pH and ANC and higher
nitrate concentrations during storm events than did lower elevation streams (Neff et al..
2013). High-elevation streams (823-966 m elevation) in the Little Pigeon River
watershed had lower ANC and pH during storm events and higher sulfate, nitrate, and
organic acid anion concentrations than at low flow (Devton et al.. 2009). Streams in
watersheds influenced by the Anakeesta Formation also had significantly lower pH and
higher Al concentrations during storm flow (Neff et al.. 2013).
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GRSM historically received some of the highest deposition loads of NOx and SOx in the
U.S. In the late 1980s, a monitor at 1,740 m elevation within the park estimated S
deposition of approximately 2,200 kg S/ha/yr, about half of which was occult deposition
(DW and SE. 1992); long-term trends in deposition from a monitor placed at a lower
elevation within the park are shown in Figure 16-18 and Figure 16-19. Much of the
historically deposited N and S remains stored within park ecosystems, with N stored in
soils, biomass, and coarse woody debris (Cai et al.. 2011b; Creed et al.. 2004). and S
adsorbed to soil (Fakhraci et al.. 2016; Cai et al.. 2011 b). The capacity of these
compartments for storing N and S are not infinite, and the capacity to retain N has been
exceeded in some areas (ecosystems have reached N saturation), as evidenced by
increased nitrate leaching and acidification of some surface waters. This suggests that
ecosystems are impaired by N deposition, and that there may be a lag time in ecosystem
recovery if deposition is reduced (Fakhraci et al.. 2016; Cai etal.. 201 la).
The range of elevation and precipitation within the park has allowed distinct terrestrial
and aquatic communities to form at different elevations, making the park one of the most
biologically diverse areas in North America. Furthermore, because the park remained
unglaciated during recent ice ages, it served as a refugium for northern plant and animal
species, as well as an evolutionary generator of new species. The All Taxa Biodiversity
Initiative is an ongoing project that seeks to document all species within the park's
boundaries, and as of March 2016, there were 19,260 species across all domains
documented within the park (https://www.dlia.org/smokies-species-tallv').
Land cover in this case study region is mostly hardwood forest; coniferous forest occurs
primarily at the higher elevations in and around GRSM (Figure 16-17). Distributed in
elevational zones from highest elevation to lowest, forest types in the park include
spruce-fir forest, beech-yellow birch forest, pine-oak forest, hemlock forest, and cove
hardwood forest. Different forest types may support endemic animal, plant, fungal, and
lichen species, as well as more broadly distributed species. There are 831 species of
lichens documented within the park (https://www.dlia.org/smokies-species-tallv); lichens
are some of the most sensitive organisms in terrestrial ecosystems to atmospheric
deposition (see Appendix 6.3.7). and work in Eastern Temperate forests (Cleavitt et al..
2015; Will-Wolf et al.. 2015) suggests that high atmospheric deposition can negatively
impact their condition, species richness, and community composition. The forests in the
park support rich understory plant communities, which are sensitive to deposition effects
such as shifts in mycorrhizal communities, competitive exclusion, and increases in
herbivory, all of which can lead to declines in herbaceous species richness (Simkin et al..
2016).
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A number of threatened and endangered species occur within the park (Table 16-17).
GRSM is a key preserve for North American amphibian species (Nickerson et al.. 2002;
Hyde and Simons. 2001). including salamander species which have seen their North
American populations decline over the past half century (Caruso and Lips. 2013). The
park's high biodiversity has been recognized by its designation as a UNESCO World
Heritage Site and International Biosphere Reserve. The broader Tennessee-Cumberland
region contains the highest diversity of freshwater mussel and crayfish species and the
highest levels of aquatic species endemism in North America (Abell et al.. 1999). Within
the park, sampling between 1993 and 2001 found 90-168 genera of benthic
macroinvertebrates in 13 different streams (Milncr et al.. 2016). A national analysis of
federally listed species identified 49 aquatic and 4 terrestrial threatened or endangered
species impacted by anthropogenic N in the U.S. Fish and Wildlife Service Region 4,
which encompasses 10 southeastern states, including Tennessee and North Carolina
(Hernandez et al.. 2016).
Table 16-17 Species in the Southeast case study region that are listed as
threatened or endangered or as a species of concern.
Threatened and endangered species known or believed to occur in Great Smoky Mountains NP
Mammals
•	Myotis septentrionalis, northern long-eared bat—threatened
•	Myotis sodalis, Indiana bat—endangered
•	Glaucomys sabrinus coloratus, Carolina northern flying squirrel—endangered
Birds
•	Picoides borealis, red-cockaded woodpecker—endangered
Fish
•	Etheostoma sitikuense, Citico darter—endangered
•	Noturus baileyi, smoky madtom—endangered
•	Noturus flavipinnis, yellowfin madtom—threatened
Arthropods
•	Microhexura montivaga, spruce-fir moss spider—endangered
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Table 16-17 (Continued): Species in the southeast case study region that are
listed as threatened or endangered or as species of
concern.
Threatened and endangered species known or believed to occur in Great Smoky Mountains NP
Plants
•	Geum radiatum, spreading avens—endangered
•	Spiraea virginiana, Virginia spiraea—threatened
•	Gymnoderma lineare, rock gnome lichen—endangered
Species believed to have been extirpated from the park
•	Canis lupus, gray wolf—endangered mammal
•	Canis rufus, red wolf—endangered mammal
•	Felis concolor couguar, eastern puma or cougar—endangered mammal
•	Erimonax monachus, spotfin chub—threatened fish
Federal Species of Concern found in the park
Mammals
•	Myotis leibii, eastern small-footed bat
•	Sorex palustris, water shrew
•	Sylvilagus obscurus, Appalachian cottontail
Birds
•	Ammodramus henslowii, Henslow's sparrow
•	Contopus cooperi, olive-sided flycatcher
•	Dendroica cerulea, cerulean warbler
•	Loxia curvirostra, red crossbill
•	Poecile atricapillus, black-capped chickadee
•	Sphyrapicus varius, yellow-bellied sapsucker
•	Vermivora chrysoptera, golden-winged warbler
Amphibians
•	Cryptobranchus alleganiensis, eastern hellbender
•	Desmognathus aeneus, seepage salamander
•	Eurycea junaluska, Junaluska salamander
Fish
•	Percina squamata, olive darter
•	Phoxinus tennesseensis, Tennessee dace
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Table 16-17 (Continued): Species in the southeast case study region that are
listed as threatened or endangered or as species of
concern.
Threatened and endangered species known or believed to occur in Great Smoky Mountains NP
Plants
•	Abies fraseri, Fraser fir
•	Calamagrostis cainii, Cain's reed-bent grass
•	Cardamine clematitis, mountain bittercress
•	Glyceria nubigena, Smoky Mountain manna grass
•	Silene ovata, Blue Ridge catchfly
NP = national park.
Source: National Park Service https://home.nps.gov/qrsm/learn/nature/te-species.htm
A number of disturbances in the park can interact with acidic deposition to affect
biodiversity. Invasive species alter the distribution of native species within the park; for
example, native brook trout (Salvelinns fontinalis) have been displaced from
approximately 75% of their historical range within park streams by rainbow trout
(Oncorhynchus mykiss) introduced to the park in the early twentieth century for
recreational fishing (Kanno et al.. 2017). Recent eradication campaigns of rainbow trout
and reintroduction of brook trout within a set of park streams showed that brook trout
quickly re-established populations (Kanno et al.. 2016). Both introduced and native
pathogens are also changing plant communities within the park. Fraser fir is endemic to
the Southeast and sensitive to acidic deposition, and 74% of remaining Fraser fir (Abies
fraseri) forests are within GSMNP boundaries (Kavlor et al.. 2017). The invasive balsam
woolly adelgid (Adelges piceae) killed mature stands of Fraser fir which historically
dominated the highest elevations of the park, (Kavlor etal.. 2017; Van Miegroet et al..
2007). The most recent survey of Fraser fir was conducted in the park in 2010 and found
the highest elevation stands regenerating despite the continuing presence of the adelgid,
suggesting some population recovery following reductions in acid deposition (Kavlor et
al.. 2017). Ongoing damage from the invasive hemlock woolly adelgid (Adelges tsugae)
may kill the mature hemlock (Tsuga spp.) stands which grow along streams in the park
and which regulate stream temperature by the dense shading they provide (Martin and
Goebel. 2012). Recent research suggests that N deposition may alter beech (Fagus
grandifolia) bark chemistry, making the trees more susceptible to mortality caused by
beech bark disease (attack by beech scale, Cryptococcus fagisuga, followed by infection
by fungal pathogens in the Neonectria genus), which was first detected in the park in
1993 (Cale et al.. 2017). In addition to disturbances caused by plant pathogens, shifting
wildfire frequency affects the distribution of plant communities within the park and the
wider region. In late November 2016, a fire burning at Chimney Tops spread rapidly
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north across the park and into the nearby cities of Gatlinburg and Pigeon Forge. The fire,
aided by drought conditions and strong winds, burned 17,140 acres
(https://inciweb.nwcg.gOv/incident/5112/). Wildfire suppression was the policy of the
Park from its founding in 1931 until 1996, when a program of prescribed burns was
instituted in order to regenerate forests of fire-dependent oak and pine species on dry
ridges (Schwartz ct al.. 2016).
Precipitation and temperature alter the responses of terrestrial and aquatic ecosystems to
N and S deposition. Between 1900 and 2011, annual precipitation increased 10 cm while
there was no overall trend in temperature over the same period of time at stations within
GSMNP (Lesser and Fridlev. 2016). At nearby Grandfather Mountain, NC, summer
temperature increased by more than 1.4°C between 1956 and 2011 (Soule. 2011).
Changes in temperature may alter species distributions. For example, native brook trout
inhabit about 970 km of stream length in GRSM, most of which is above 900 m elevation
(Neff et al.. 2009). Brook trout have disappeared from six high-elevation streams in the
Little Pigeon River watershed (Devton et al.. 2009; Neff et al.. 2009). Simulation
modeling by McDonnell et al. (2015) showed that acidification of streams may hinder the
movement of southern Appalachian aquatic species (including brook trout) to colder,
higher elevation habitats if stream temperatures rise.
Like GSMNP, Linville Gorge Wilderness (LGW) contains plant communities structured
by elevation. LGW comprises 43.9 km2 of mountain ridges and gorges that range over an
elevation of 426 to 1,250 m. Forest cover in the gorge is 43% coniferous tree species and
55% deciduous tree species, in plant communities from low to high elevation of acidic
cove, mesic oak-hickory, and xeric pine-oak or oak heath zones (Kantola et al.. 2016).
Soils in LGW formed from parent material with very low concentrations of Ca, Mg, and
K; and historic acidic deposition has further depleted these nutrients while decreasing soil
pH and increasing soluble aluminum (Elliott et al.. 2013). Experimental addition of
dolomitic limestone resulted in transient improvements in soil Ca:Al ratios, while
wildfire, which can release base cations from biomass into the soil, had no measurable
impact on Ca pools in soils (Elliott et al.. 2013). Between 1992 and 2011, forest structure
in LGW shifted towards trees in younger age classes, as outbreaks of beetles and
hemlock wooly adelgid, drought, and wildfire caused mortality in older and larger age
classes of trees (Hagan et al.. 2015).
16.3.1.2. Legal Authorities
GRSM is a Prevention of Significant Deterioration (PSD) Class I Area. The CAA
(42 USC 7470) authorized Class I areas to protect air quality in national parks over
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6,000 acres and national wilderness areas over 5,000 acres in an effort to preserve pristine
atmospheric conditions and air quality related values (AQRVs).
Wilderness areas in North Carolina located in the vicinity of GRSM that also have been
designated PSD Class I include Shining Rock, Linville Gorge, and Joyce Kilmer
Memorial Forest. All are managed by the USDA's U.S. Forest Service. Shining Rock
consists of 74 km2 of high elevation (1,450-1,550 m) hardwood forests dominated by
yellow birch and red maple on soils formed in gneissic bedrock. Linville Gorge is
43.9 km2 of acidic cove and slope forests at intermediate elevations (1,090-1,160 m)
dominated by chestnut oak and red maple on quartzite parent material. Joyce Kilmer
Memorial Forest is 68.1 km2 of low elevation (250-450 m) cove hardwoods, dominated
by tulip poplar, oaks, and eastern hemlock stands, on metasandstone parent material.
Additional information on these three wilderness areas is in (Elliott et al.. 2008). The
geography of the Class I wilderness areas described in this case study are shown in
Figure 16-16; further details on this study can be found in Appendix 16.2.3.2.
Class I areas are subject to the PSD regulations under the CAA. PSD preconstruction
permits are required for new and modified existing air pollution sources. Air regulatory
agencies are required to notify federal land managers (FLMs) of any PSD permit
applications for facilities within 100 km of a Class I area.26 The FLMs are entitled to
review and comment on PSD Class I permit applications with the authorized permitting
agency. Information on air quality monitoring by NPS within GRSM is available at the
GRSM Air Quality website (https://www.nps.gov/grsm/learn/nature/air-qualitv.htm).
26 http://webcam.srs.fs.fed.us/psd.
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• Long-term monitored stream sites (42)
	States
I I Class 1 wilderness areas
National forests
i Simulated watersheds (30)
Cherokee j
I For^fet
Plsgah
National Forest
Lost Bottom Creek
Shutts Prong
Noiand Divide Creeks
ionai Forest
ining Roclf'
/ilderne^S
Great Smoky Mountains
National Park
Joyce raimer-Slickrock
Wilderness
Cherokee
National Fori
Nantahala
National Forest
Stimter
NaHonal Forest
Cohutta
Wilderness
Chattahoochee
National Forest
Source: Fakhraei et al. (2016).
Figure 16-16. Great Smoky Mountains National Park and nearby Class I
wilderness areas, with emphasis on water sampling locations
within GSMNP and critical loads for watersheds described in
Appendix 16.2.3.2.
16.3.1.3.	Regional Land Use and Land Cover
1	Land cover in this case study region is mostly hardwood forest; coniferous forest occurs
2	primarily at the higher elevations in and around GRSM (Figure 16-17). Although located
3	generally downwind of populous areas, GRSM has only a few human population centers
4	of any magnitude nearby. They include Knoxville, TN, 30 mi from the park (pop.
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181,000); Asheville, NC, 30 mi (pop. 87,000); Johnson City, TN, 70 mi (pop. 64,000);
and Greenville, SC, 90 mi (pop. 61,188).
Southern Appalachian Case
Study Region: Land Cover
Case Study Locations
Native American
Reservations
Developed Land (increasing intensity)
| Barren Land
] Deciduous Forest
m Evergreen Forest
| Mixed Forest
	] Shrub/Scrub
I Grassland/Herbaceous
H Cultivated Crops
I Water/Wetlands
jtat ShiokV Sliiulitain-
'.SatifsiliV! Park
Figure 16-17 Land cover in the southern Appalachian Mountains case study
region.
16.3.2.	Deposition
3	The highest elevations of the park have received some of the highest rates of acidic
4	deposition in the U.S. (Weathers et al.. 2006; Herlihv et al.. 1993) due largely to the
5	location of upwind power plants, major agricultural regions, and the substantial amount
6	of annual precipitation.
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Characteristics of nitrogen and sulfur deposition affecting the Great Smoky Mountains
Study Area are shown in Figure 16-18-Figure 16-21. Data shown in
Figure 16-18-Figure 16-23 were obtained from the hybrid modeling/data fusion product,
TDEP, http://nadp.sws.uiuc.edu/committees/tdep/tdepmaps/ and described earlier in
Appendix 2.1. However, the time series of wet deposition is taken directly from data on
the NADP/NTN website. This was done to use the long-term record of wet deposition to
illustrate trends in deposition since the passage of the Clean Air Act Amendment,
because the CMAQ dry deposition simulations involved in estimating TDEP total
deposition extend back only to 2000. Figure 16-18 shows the 25-year-long time series for
wet deposition for NO.-; . NHU+, SO42 . and H+ measured at the NADP/NTN monitoring
site Great Smoky Mountain NP (TNI 1), as well as the partitioning between oxidized and
reduced N Deposition of nitrogen, which is estimated to be mostly in oxidized form in the
study area. Although most of the area in Figure 16-18 is subject to N deposition in
oxidized form, there are areas, principally in northern Georgia, where N is deposited
mainly in reduced form. The graph of wet deposition of all chemical species shows that
downward trends in NO3 . NFU+, S042 , and H+ are consistently found over the past
25 years, although the rate of decrease has been irregular, with occasional increases.
Annual Wet Deposition and 3-Year Moving Average at
Site TN11:1990- 2014
Z
E
r
a
0 mo
• W.'
•W
•M,1-

.v**v
m tw iw
m xm au *u>
Yetr
~ Knoxvtlle. IN
O Monitor TW» 11
0 Monitor Lourticn
Appalvchia Study A/
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Figure 16-19 shows the 25-year-long time series for wet deposition of N and S in terms
of units commonly used to determine critical loads within GSMNP. Critical loads for
eutrophication are commonly expressed in terms of kg N/ha/yr. Critical loads for
acidification are commonly expressed in terms of kg S/ha/yr, or because both N and S
deposition can contribute to acidification, in terms of eq/ha/yr.
s.
2 «
s
8
Annual Wet Deposition and 1-Year Moving
Average at Site TN11:1990 - 2014
•	hf* kwn M4.'
•	4 fr om iO,'"
v«Vy* v
V * *
Annual Wet Deposition and 3-Year Moving
Average at Site TN11:1990 - 2014
z
I
a *
s
a.
9
o
' *
\	/X
I /	\
V *
' 'l '-fX
1M* 11M-
jwa jo u joit
t frftl U>b
JOM JQOC JOLI J016
eq = H+ equivalents; ha = hectare; kg = kilogram; N = nitrogen; NH4+ = ammonium; N03 = nitrate; S = sulfur; S042 = sulfate;
yr = year.
Notes: on the left, N and S deposition are shown in units of kg/ha/yr of N or S. On the right, N and S deposition are combined into
units of acidifying deposition, eq/ha/yr of S + N.
Source: NADP/NTN.
Figure 16-19 Trends in wet deposition of nitrogen and sulfur in Great Smoky
Mountain National Park, 1990-2014.
Figure 16-20 shows the 3-year average total deposition of N and S for 2011-2013.
Surrounding areas in southern states and inserts showing the coterminous U.S. are shown
to place the depositional environment in context. Comparison of maps indicates that the
general pattern of deposition of N and S is broadly similar, with higher values within the
study area than outside of it. Current rates of deposition of S are considerably lower than
along the Ohio River Valley, which might be expected given that the Ohio River Valley
is a major industrial region, and GRSM is a national park. Wet and dry deposition of S
were roughly equal across the study area. See Appendix 2.4. Appendix 25. and
Appendix 2J) for more information on deposition in the U.S. Other maps showing the
contributions of individual species to dry and/or wet deposition, based on TDEP are
given in Appendix 2.7.
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i«y V /

Ylrtt^*3
Tfro"****®
• ~
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Q Monitor TN ¦ 11
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ApplteClM Study Am*
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10 km
Deposition in GRSM,
2000 {kg ha^-yr1)
Sulfur	Nitrogen
6.5-13.5
13.5-20.2
20.2-26.9
26.9-33.6
33.6-41.5
4.6-10.0
10.0—15.0
15.0-20.0
20.0-25.0
25.0-30.9
GRSM = Great Smoky Mountain National Park; ha = hectare; kg = kilogram; yr = year.
Source: Weathers et al. f2006).
Figure 16-21 Modeled sulfur and nitrogen deposition to the Great Smoky
Mountain National Park for the Year 2000.
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Emission sectors
NH3 (livestock)
[ NH3 (fertilizer)
^ NH3 (natural)
U NOx (surface inventory)
| NOx (electric generating units)
| NOx (non-electric generating
industrial stacks)
~| NOx (aircraft)
	NOx (lightning)
NOx (soil)
Geographic footprint of Nr deposition
"I	I 	^	I	I	M
0.005 0.01 0.02 0.03 0.04 0.06 0.08 0.1
[kg N/ha/yr]
SM 10.4
Source: adapted from Figure 5 in Lee et al. (2016).
Figure 16-22. Annual-averaged monthly footprint of reactive N deposition in
Great Smoky Mountain National Park (10.4 kg N/ha/yr), and pie
chart of fractional contribution from emission sectors, as
estimated by GEOS-Chem adjoint model.
16.3.3.	Critical Loads and Other Dose-Response Relationships
1	The following sections describe critical loads determined for the Great Smoky Mountains
2	National Park, for the larger southern Appalachian region, or for the Level I ecoregion in
3	which the park resides, the Eastern Temperate Forests.
16.3.3.1. Empirical Studies
4	Thresholds of deposition are quantified for GRSM and the broader region. Sampling of
5	soil and streams at high elevations in the park in the mid 1990s found elevated
6	extractable inorganic N concentrations at high elevations (Garten. 2000). as well as
7	substantial nitrate leaching in streams from these watersheds (Van Miegroet et al.. 2001).
8	Because NOx deposition was determined by (Van Miegroet et al.. 2001) in the
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watersheds to be 32 kg N/ha/yr, Gilliam et al. (201 la) stated that the nitrate leaching
critical load would be <32 kg N/ha/yr. Additional water quality critical loads for the park
are based on models (see below). The comparable critical load set for leaching in the
Eastern Temperate Forests is 8 kg N/ha/yr, which is lower than the Gilliam et al. (201 la)
value. This critical load is set to protect hardwood forest soils from NOs leaching to
surface water and is based on a synthesis of data across northeastern forests (Aberetal..
2003). This quantity is in agreement with an earlier recommendation for Class I
wildernesses in the Forest Service's eastern region that N deposition should not exceed
5-8 kg N/ha/yr to protect terrestrial biota (Adams et al.. 1991). That report also
recommended an annual load for total sulfur deposition not to exceed 5-7 kg S/ha/yr to
protect terrestrial biota (Adams et al.. 1991). The critical loads set by Adams et al. (1991)
were based on observations across class I wildernesses in the Forest Service eastern
region; the deposition limits were based on the deposition loads of three Class I areas
where eutrophication and acidification were not observed.
A number of additional empirical critical loads have been set for the Eastern Temperate
Forests ecoregion based on protection of sensitive biota from the effects of nitrogen
deposition (Gilliam et al.. 201 la; Pardo etal.. 2011a). A critical load of >3 kg N/ha/yr
protects growth rates or survivorship of sensitive tree species, including red pine, yellow
birch, scarlet and chestnut oaks, quaking aspen, and basswood. A critical load of
4-8	kg N/ha/yr was set for Eastern Temperate Forests based on extrapolation from data
collected in Northwest forests to preserve sensitive lichen species and prevent lichen
community composition shifts towards more eutrophic species. A critical load of
5-10	kg N/ha/yr protects ectomychorrizal fungi associated with coniferous tree species
from community composition change. A critical load of <12 kg N/ha/yr protects
arbuscular mycorrhizal fungi associated with hardwood tree species from decreases in
abundance and community composition change. A critical load of <17.5 kg N/ha/yr
protects forest understory communities from herbaceous species loss and shifts in plant
community composition. These critical loads (3-17.5 kg N/ha/yr) represent protection for
a broad range of sensitive terrestrial species within the Eastern Temperate Forest
ecoregion (Gilliam et al.. 201 la; Pardo etal.. 2011a).
Two more recent studies document empirical critical loads for the broad Eastern
Temperate Forest ecoregion. Cleavitt et al. (2015) conducted surveys and used FIA
assessments of lichen species in the Northeast, with particular focus on the four Class I
wilderness areas in the region (Lye Brook, VT; Great Gulf, NH; Presidential Range—Dry
River, NH; and Acadia National Park, ME). In general, cumulative N and S deposition
over the course of the modeled period (2000-2013) was a more powerful explanatory
factor than annual N or S load for the lichen data. This work determined a critical load of
4.3-5.7 kg total N (oxidized + reduced N)/ha/yr based on lichen species richness, the
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abundance of sensitive species (higher species richness of cyanolichens and fruticose
lichens below CL), and thallus condition. Cumulative S and N deposition were both
equally powerful predictors of thallus condition, but no S critical load was determined
(Cleavitt et al.. 2015). A study by Simkin et al. (2016) used FIA data to determine critical
loads for herbaceous species in a national assessment. The critical load for this analysis
was placed to prevent any loss of herbaceous species richness, and was determined to be
7.8-19.3 kg N/ha/yr for closed-canopy forest herbaceous communities in the Eastern
Temperate Forest ecoregion. The critical load determined for open-canopy (grassland,
shrubland, and woodland) herbaceous communities in this ecoregion was lower,
6.6-9.7 kg N/ha/yr (Simkin et al.. 2016). These critical load estimates may be too high
for GRSM, as Simkin et al. (2016) determined that critical loads tended to decrease under
conditions of low soil pH, high precipitation, and high temperatures, all of which affect
park ecosystems (Appendix 6.2.3.2).
16.3.3.2. Modeling Studies
Some of the critical loads modeled for GSMNP are aimed to protect terrestrial
ecosystems from N saturation and associated acidification. The Integrated Forest Study
sampled forests across North America, and several plots were located at high elevations
within GSMNP. A critical load of 2.5-9 kg N/ha/yr was determined for the park using
SMB modeling (OiaandArp. 1998). A decade later, Pardo and Duarte (2007) used SMB
and data from GRSM to determine critical loads for different forest types within the park.
Critical loads of 2.7-2.8 kg N/ha/yr to prevent terrestrial ecosystem eutrophication were
determined for low- and mid-elevation hardwood forests within the parks, while the
critical load for high-elevation spruce-fir forests was 3.4-7 kg N/ha/yr. Pardo and Duarte
(2007) also developed critical loads to protect GSMNP forests from acidification based
on pH, Al, and base cation threshold values. In low-elevation hardwood forests, the most
protective critical load was 1,080 eq S + N/ha/yr; in mid-elevation hardwood forest, the
most protective critical load was 1,650 eq S + N/ha/yr; and in high-elevation spruce-fir
forest, the most protective critical load was 2,000 eq S + N/ha/yr (Pardo and Duarte.
2007).
Critical loads of N and S deposition to protect aquatic ecosystems depend upon
underlying geology, elevation, and sensitivity of streams (i.e., preindustrial ANC) in the
southern Appalachian Mountains (Sullivan et al.. 201 lb). A recent application of the
PnET-BGC model was used to determine target loads to protect aquatic ecosystems in
12 streams within GSMNP from acidification. The target load to achieve a target ANC of
50 |icq/L by 2050 was deposition of 270-1,400 eq (S + N)/ha/yr in streams at <1,000 m
elevation and which did not drain watersheds within the Anakeesta Formation (Zhou et
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al.. 2015a). In an acid-sensitive, high-elevation watershed with drainage from the
Anakeesta Formation, no reduction in deposition was sufficient to reach a target of
ANC = 50 (j,eq/L, but a target load of 270 eq (S + N)/ha/yr resulted in ANC recovery to 0
by 2050 (Zhou et al.. 2015a). In a follow-up analysis, Fakhraei et al. (2016) and Fakhraei
et al. (2017a) determined total maximum daily loads of acid deposition for each of the
twelve 303(d)-listed stream watersheds at high elevations within the park. The target date
for recovery to pH of 6.0 was 2150 for these model runs, and critical loads ranged
between 240 and 960 eq/ha/yr of SOr + NOs + NH4+ deposition to eight of the twelve
watersheds. For the remaining four watersheds, no reduction in deposition was sufficient
to achieve pH of 6 by 2150; recovery in these streams is projected to take centuries
(Fakhraei et al.. 2017a). These target loads are in good agreement with critical loads set
for S only deposition based on MAGIC and Steady-State Water Chemistry (SSWC)
modeling for the broader southeastern Appalachian region (McDonnell et al.. 2014b).
The S only critical load was based on achieving a target stream ANC of 50 |icq/L. and for
high elevations in the park, the critical load was 0-250 eq S/ha/yr. Most of the
low-elevation area within the park had a critical load of 250-500 eq S/ha/yr by this
method (McDonnell et al.. 2014b).
16.3.4. Characterization and Long-Term Monitoring
The majority of studies conducted in this case study region have evaluated the effects of
acidification on ecosystems rather than the effects of eutrophication (see Table 16-18).
Acidification has been considered to be the dominant result of air pollution stress in the
region. Biogeochemical processes that govern watershed responses to acidification are
generally similar between the Northeast and the southern Appalachian Mountains in most
respects. The major exception is that soils in the southern Appalachians are unglaciated
and therefore more weathered, with greater sulfate adsorption capacity, than soils in
northeastern forests. Cai et al. (2010) used long-term monitoring data to generate an
input-output budget to identify processes that have influenced stream pH and ANC at the
Noland Divide watershed in Great Smoky Mountains NP during the period 1991-2006.
The majority (about 61%) of the net SO42 entering the watershed was retained,
confirming that S adsorption on soil is an important biogeochemical process in this
watershed. However, during large precipitation events, SO42 in wet deposition moved
more directly and rapidly to streams, contributing to episodic stream acidification.
Fakhraei et al. (2016) project that GRSM streams in watersheds with a low capacity to
adsorb SO42 and low N retention will respond positively (increase stream pH and ANC)
to additional reductions in S and N deposition, while streams in watersheds with high
S042 adsorption and low N retention (typically higher elevation watersheds) are less
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1	responsive (or unresponsive) to reduced S and N deposition. In these watersheds,
2	increases in soil pH due to reduced deposition are projected to increase desorption of
3	accumulated SO42 from soils, resulting in a SO42 pulse to streams that will slow
4	recovery of stream ANC and pH (Fakhraci et al.. 2016). Monitoring data from the Great
5	Smoky Mountains NP indicated that the high S absorption in watershed soils delayed
6	recovery from previous stream acidification. Stream chemistry at 42 monitoring sites in
7	the park did not show substantial changes over the recent period of long-term monitoring,
8	1991-2014 (Fakhraci etal.,2016V
Table 16-18 Example soil, terrestrial biota, and surface water acidification
characterization and long-term monitoring studies in the southern
Appalachian Mountains region.
Study
(HERO ID)
Location
Time
Period
Focus
Results
Soil studies
Rice et al. (2014)
Southeastern
U.S.
2006-2021
Sulfate mass balances
for 27 forested
watersheds to estimate
cross-over years
(change from retaining
to releasing SO42").
Documented extensive S
adsorption on soils
Lawrence et al.
(2015b)
AT corridor,
including
through
GRSM
2010-2012
Soil sampling along AT
corridor and compilation
of other existing soil
chemistry data
Documented occurrence of low
soil base saturation along the
AT corridor, including within
GRSM
Cai et al. (2011a)
Noland Divide,
GRSM
2008
Laboratory soil columns
and in situ lysimeters to
determine response of
soil solution to reduced
acidic deposition.
Recovery of soil water, as
represented by renewed Ca2+
and Mg2+ retention, was
estimated to occur at S
deposition reduction of 61% of
ambient.
Cai et al. (2012)
Noland Divide,
GRSM
2009
Four sites along
elevational gradient.
Base saturation values <7% and
Ca:AI ratio <0.01 indicated high
acid sensitivity.
Elliott et al. (2008)
Three FS
Class I areas
in North
Carolina

NuCM model
Low Ca:AI (-0.3) in A-horizon at
Shining Rock and Linville Gorge
suggested stress to forests from
soil acidification.
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Table 16-18 (Continued): Example soil, terrestrial biota, and surface water
acidification characterization and long term monitoring
studies in the southern Appalachian Mountains region.
Study

Time


(HERO ID)
Location
Period
Focus
Results
Terrestrial biota studies
Hames et al. (2002)
Assessment of
1995-1999
Wood thrush breeding
Low breeding success in the

650 study sites

success
wood thrush strongly correlated

across the


with low soil pH

range of the




wood thrush in




the eastern U.S.



McNultv and Boqas
Western NC
1999-2002
Red spruce and southern
CL for protecting forest
(2010)


pine beetle
ecosystems may not accurately




reflect risk to ecosystem health




due to multiple stresses,




including climate change, insect




infestation, and N supply.
Surface water studies
Scheffe et al. (2014)
U.S., including

Critical load by
AAI model

southern

ecoregion, as reflected


Appalachian

in AAI


Mts. region



Sullivan et al.
SAMI region

MAGIC model
Streams exhibited broad range
(2004): Sullivan et



of responses to changes in
al. (2002)



future S deposition, including




pronounced base cation




depletion. Recovery from past




acidification expected to be slow




and gradual.
Sullivan et al.
Southern
-2,000
Delimited high-interest
Lithology and elevation
(2007b)
Appalachian

area for water
explained locations of low-ANC

Mts.

acidification based on
streams.



geology and elevation,




which included almost




all known low-ANC




(<20 peq/L) streams in




the region.

Sullivan et al.
Southern

MAGIC model
Estimate target loads of S
(2011b)
Appalachian


deposition to protect stream

Mts.


ANC to multiple levels (0, 20,




50, 100 peq/L) at multiple points




in time (2020, 2040, 2100). TLs




ranged from 0 to values many




times higher than ambient S




deposition.
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Table 16-18 (Continued): Example soil, terrestrial biota, and surface water
acidification characterization and long term monitoring
studies in the southern Appalachian Mountains region.
Study
(HERO ID)
Location
Time
Period
Focus
Results
Sullivan et al.
(2011c)
Southern Blue
Ridge
province of
southern
Appalachian
Mts. region
MAGIC model
Modeled 66 stream watersheds,
all of which were simulated to
have had preindustrial
ANC > 30 peq/L. In 2005, 30%
of streams had ANC below this
level. Median stream lost
25 peq/L of ANC between 1860
and 2005.
Hops and McMahon Southeastern
(2009)	U.S.
SPARROW model	Examined N transport relative to
landscape characteristics. N
transport coincided with Level III
ecoregion boundaries. Lower
fraction of N input delivered to
stream at locations where
primary flow path is shallow.
Cai etal. (2011b) Noland Divide, 1991-2007 Stream water chemistry Volume-weighted NO3"
GRSM	trends.	concentration decreased
0.56 [jeq/L/yr; stream pH and
SO42" concentration did not
change.
Cai et al. (2010) Noland Divide, 1991-2007 Developed input-output About 61% of incoming SO42"
GRSM	budget.	was retained; during rain events,
SO42" moved more readily to
streams.
Lawrence et al.
(2015b)
AT corridor,
including
through
GRSM
2010-2012
Sampled more than
200 streams along the
AT corridor, including
under both high- and
low-flow conditions.
Documented wide variability in
stream ANC along the AT
corridor.
Robinson et al.	Noland Divide, 1990-1999 Developed MLR model
(2004)	GRSM	of water chemistry.
Showed decrease in ANC over
time.
Robinson et al.
(2008)
GRSM
1993-2002
Conducted trends
analysis for 90 stream
sites, which showed
decreases in pH and
SO42" concentrations.
Extrapolation over 50 yr
suggested pH <6.0 for nearly all
study streams in the future.
Neff et al. (2013)
GRSM
2008-2009
Examined relationships
between stream
chemistry at base flow
and storm flow vs. basin
characteristics.
Following precipitation, pH
decreased and Al increased.
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Table 16-18 (Continued): Example soil, terrestrial biota, and surface water
acidification characterization and long term monitoring
studies in the southern Appalachian Mountains region.
Study	Time
(HERO ID)	Location	Period	Focus	Results
Peyton et al. (2009) Little Pigeon	2006-2007 Characterized chemistry	During storm flow, stream pH
River	of three high-elevation	and ANC decreased: SO42",
watersheds,	streams during	NO3", and organic acids
GRSM	episodes.	increased.
AAI = Aquatic Acidification Index; Al = aluminum; ANC = acid neutralizing capacity; AT = Appalachian Trail; Ca = calcium;
CL = critical load; FS = Forest Service; GRSM = Great Smoky Mountains National Park; L = liter; |jeq = microequivalents;
MAGIC = Model of Acidification of Groundwater in Catchments; Mg2+ = magnesium; MLR = Multiple Linear Regression models;
N = nitrogen; N03" = nitrate; NuCM = Nutrient Cycling Model; S = sulfur; SAMI = Southern Appalachian Mountains Initiative;
S042" = sulfate; SPARROW = Spatially Referenced Regressions on Watershed Attributes; TL = target load; yr = year.
16.4. Tampa Bay Case Study
16.4.1. Background
This case study provides an overview of the Tampa Bay estuary's ecology and research
related to nutrient impairments and nitrogen (N) deposition. It also reports on the
improvements to the bay's ecology being realized from strategies to reduce nutrient
loading to the bay. Years of historical data going back to the 1950s and unique
partnerships among government, industry, and citizen groups make Tampa Bay an
interesting example of the varied issues involved in monitoring and managing water
quality and N loading in an urban estuary. During the 1970s and early 1980s eutrophic
conditions were commonly observed in Tampa Bay. Contrary to many other estuaries
suffering from eutrophication worldwide, Tampa Bay appears to be recovering since the
mid 1980s as a result of the nutrient management strategy defined by the community
(Greening et al.. 2014; Morrison et al.. 2011). These approaches have included upgraded
sewage treatment, reduction of atmospheric N emissions, and controls on stormwater and
point sources. Atmospheric deposition continues to be a source of N to Tampa Bay and,
with the decreases in N loading from point sources, represents a greater proportion of
total N loading to the bay (Greening et al.. 2014). Estimates from both modeling and
measurement indicate that direct atmospheric N deposition to Tampa Bay is between 14
and 30% while total (direct plus indirect) atmospheric deposition is between 35 and 71%
(Poor etal.. 2013b; Poor et al.. 2013a; Greening and Janicki. 2006).
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Information regarding sulfur deposition is not included in this case study because N is a
primary contributor to nutrient enrichment and the focus of the monitoring studies used to
compile this case study.
16.4.1.1. Description of Case Study Region
The Tampa Bay estuary (Figure 16-23) is located on the eastern shore of the Gulf of
Mexico in Florida. At more than 1,000 km2, it is Florida's largest open water estuary,
with an average water depth of just 4 m and annual average rainfall rate of more than
125 cm (TBEP. 2015; Poor et al.. 2013a; TBEP. 2012b). It lies in the Northern Gulf of
Mexico marine ecoregion near the South Florida/Bahamian Atlantic marine ecoregion
(Wilkinson et al.. 2009). This transition zone between warm-temperate and tropical
ecoregions, combined with its shallow waters, large size, and gradient of fresh water to
salt water, allows the bay to support a diversity of organisms and habitats (Greening et
al.. 2014; USGS. 2011; USFWS. 1990. 1988).
Seagrass, a type of submerged aquatic vegetation (SAV), and mangrove forests are the
most prominent estuarine habitats within Tampa Bay (TBEP. 2015). Tidal marshes and
mud flats are also found throughout the estuary (USGS. 2011). Collectively, these
habitats support a large biodiversity of flora and fauna and play a role in nutrient cycling.
Seagrasses stabilize sediments, are a food source for wildlife, provide habitat, and serve
as a nursery for large numbers of fish and shellfish species. (Greening et al.. 2014).
Mangrove roots help to stabilize the shoreline and provide nursery habitat for important
fishery species. Mud flats and salt marshes in the bay provide habitat for wading birds,
while oyster reefs near river mouths serve to naturally filter the water and attract
recreational fish species, making them popular fishing spots (TBEP. 2015). Federally
endangered West Indian manatees (Trichechus manatus), a migratory species, visit
Tampa Bay seasonally to feed in the seagrass meadows. Other endangered species in the
bay documented to be impacted by N enrichment include green turtle (Chelonia mydas)
and Atlantic sturgeon [Acipenser oxyrinchus; (Hernandez et al.. 2016)1. Three national
wildlife refuges (Egmont Key, Passage Key, and Pinellas) are located in Tampa Bay and
support populations of nesting birds.
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Hillsborough
OldNft .
Tampa ^
Bay |fli8-P
^Gandyl
Bridge J"
Hillsborough
Bay
Alalia
Pinellas
,g Middle
* Tampa
f Bay
lower
Tampa
Bay ¦
Manatee River
Manatee
Pasco
20
ZD Kilometers
Source: Poor et al. (2013a).
Figure 16-23 Tampa Bay mainstem segments arid watershed. Also shown are
the National Atmospheric Deposition (NADP) National Trends
Network wet deposition monitoring site FL41 at Verna wellfield in
Sarasota County and the NADP Atmospheric Integrated Research
Monitoring Network site FL18 in Hillsborough County.
1	The Tampa Bay watershed, located in the Eastern Temperate Forest ecoregion, is highly
2	urbanized. It is now the second largest metropolitan area in Florida with a population of
3	over 2 million people (TBEP. 2015). Pinellas, Hillsborough, and Manatee counties border
4	directly on Tampa Bay while Pasco, Polk, and Sarasota counties are also within the
5	estuary's 6,000-km2 watershed (Poor et al.. 2013a). Activities in Tampa Bay linked to the
6	local and regional economy include shipping, tourism, and commercial and recreational
7	fishing (Poor et al.. 2013a; TBEP. 2006; Tomasko et al.. 2005). More than 80 miles of
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deep-water shipping channels carry water traffic to the three seaports along the bay's
borders, in Tampa, St. Petersburg, and in northern Manatee County. The largest of these,
the Port of Tampa, consistently ranks among the busiest ports in the nation.
Mangroves, salt marshes, and SAV habitats in Tampa Bay have all experienced
significant reductions in extent since the 1950s, due to physical disturbance (dredge and
fill operations) and water quality degradation (TBEP. 2015). Excess N enrichment in the
bay is one of the main factors impacting water clarity and quality. The excess N leads to
increased algal biomass and reduced light availability for shallow-water (less than a 3-m
depth) SAV (TBEP. 2015). Numerous studies have shown that the waters of Tampa Bay
are strongly N limited, due to a high concentration of phosphorus (P) from natural
leaching and from active P mining in the watershed (Greening et al.. 2014). Therefore,
the primary focus for managing eutrophication in Tampa Bay has been the reduction of N
loading to the bay (Greening et al.. 2014).
Tampa Bay was designated an "estuary of national significance" by Congress in 1990. In
1998, the Tampa Bay National Estuary Program (TBNEP) partnership (consisting of the
Florida Department of Environmental Protection [FDEP]; the U.S. Environmental
Protection Agency; the counties of Hillsborough, Manatee, and Pinellas; the cities of
Tampa, St. Petersburg, and Clearwater; and the Southwest Florida Water Management
District) signed a formal Inter-local Agreement in support of a Comprehensive
Conservation and Management Plan (CCMP) for Tampa Bay. Upon adoption of the
Inter-local Agreement, the TBNEP became simply the Tampa Bay Estuary Program
(TBEP). TBEP continues to finance research projects in support of the CCMP and
coordinates the overall protection and restoration of the bay with assistance and support
from its many formal and informal partners (TBEP. 2015; Morrison et al.. 2011). TBEP
also coordinates the Tampa Bay N Management Consortium (TBNMC), a public-private
partnership working to reduce nitrogen loading to the bay and whose members include
local governments (Tampa, St. Petersburg, and others) along with key industries
bordering the bay, such as electric utilities, fertilizer manufacturers, and agricultural
operations.
The Tampa Bay watershed's human population is expected to increase over the next
decade and a 7% increase in N loading is projected to occur (TBEP. 2012a). To account
for these increases, the TBEP and TBNMC have adopted anew 17-ton/year reduction
target for total N loading, to offset expected increases in total nitrogen (TN) loading and
maintain TN loading rates at average annual rates for 1992-1994.
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16.4.1.2. Class I Areas
1	The Tampa Bay area is not a Clean Air Act Prevention of Significant Deterioration (PSD)
2	Class I area.
16.4.1.3.	Regional Land Use and Land Cover
3	Land use within the Tampa Bay watershed is mixed among undeveloped, agricultural,
4	urban, and mining uses, including phosphate mining and fertilizer manufacturing.
5	Figure 16-24 shows the land coverage within the bay's border communities and
6	watershed.
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Source: Sherwood et al. (2016).
Figure 16-24 Tampa Bay overview map highlighting watershed development
and land use.
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16.4.2. Deposition
Both direct atmospheric deposition of N to surface waters and indirect deposition to the
watershed with subsequent transport to the Tampa Bay represent important sources of N
to the case study area. In a 25-year time series with deposition data from the nearest
National Atmospheric Deposition Program (NADP) National Trends Network (NTN)
monitoring site (Figure 16-23. Verna wellfield in Sarasota, FL), wet deposition of NO, .
NH4+, SOr . and H+ show strong interannual variability, but downward trends in wet
deposition of all these species are consistently found over the past 25 year (Figure 16-25).
Annual Wet Deposition and 3-Year Moving Average at
Site FL41:1990-2014
• /.
\ /\
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« 550
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1	impact of air quality regulations (Poor et al.. 2013a). Although total nitrogen (TN)
2	loading to Tampa Bay has trended downward over the past 20 years, the percentage of N
3	from different sources has fluctuated. Direct atmospheric deposition has accounted for a
4	greater average percentage of the TN loading in recent years (from 2007-2011) than in
5	previous years KGreening et al.. 2014; Poor et al.. 2013a); Figure 16-261. Mainstem
6	segments of the bay vary widely in the percentage of total N loading attributable to direct
7	atmospheric deposition (Janicki Environmental. 2013).
Worst case: (~1976) 1985-1989	1990-1999
Time Period
2000-2011
Source: Greening et al. (2014).
Figure 18-26
Estimated annual loads of total nitrogen from various sources to
Tampa Bay summarized from 1976 to 2011.
BRACE also investigated the primary sources of N deposition to the bay, which included
power plants and mobile sources such as cars and trucks. Power plants had recently
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reduced N emissions through upgrades made as a result of TBEP's work with the
TBNMC (Poor et al.. 2013a'). The sensitivity results suggested that, per unit of emission,
mobile sources had a disproportionately higher contribution than power plant sources to
atmospheric N deposition to Tampa Bay (over the watershed, the mobile NOx emissions
were responsible for four times more oxidized-N deposition than the power plant
emissions; over the bay, the mobile NOx emissions were responsible for twice as much
oxidized-N deposition as the power plants). According to BRACE, reductions in N
emissions from mobile sources must be a part of the strategy to reduce N loading to the
bay, and that control of atmospheric N emissions both within and outside the Tampa Bay
watershed is important to restore and maintain good water quality and a healthy
ecosystem within the bay (Poor et al.. 2013a).
The U.S. EPA recently modeled atmospheric deposition in Tampa Bay using TDEP with
monitoring data from 2000 to 2013 (see Appendix 2.7). The general pattern of wet + dry
deposition of N (NOy and NHx) for 2011-2013 shows that fluxes are higher in the
northern half of the study area than the southern half (Figure 16-27A). The deposition of
N varies by roughly a factor of two across the study area with higher deposition of N to
the north and west of Tampa Bay and a maximum in the northeast corner of the study
area. Fluxes are typically higher than those for the rest of Florida. However, they are not
as high as in many areas of the central and northeastern U.S. Dry deposition of NO2
accounts for approximately 1/6 of total deposition of oxidized N. At Tampa Bay (and
some other NADP sites), dry fluxes are directed upward using the Community Multiscale
Air Quality (CMAQ) model, due to the implementation of bidirectional exchange of NH3.
(However, the net flux of NHx in CMAQ is still directed downward from wet
deposition.).
Most N deposition in the case study area except for the northeastern corner is estimated
to be mostly in oxidized form throughout the study area and in most of Florida, which is
consistent with many areas in the eastern and southern U.S. (Figure 16-27B). This is
generally in agreement with findings from BRACE modeling where oxidized forms of N
made up 60% of the total atmospheric loading to Tampa Bay, compared to 40% for
reduced forms such as ammonia [NH3; (Poor et al.. 2013a; TBEP. 2012bVI. In the study
area dry deposition dominates over wet deposition of N, which contrasts with many areas
in southern Florida where wet deposition dominates (Poor et al.. 2013a).
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A. Total Wet and Dry N Deposition
in Southern Florida 2011-2013
(12 km resolution)
B. Percent Oxidized N Deposition in
Southern Florida 2011-2013
(12 km resolution)
0	Monitor FL# 41
Monitor Locations
1	I Tampa Bay Study Area
ha = hectare; kg = kilogram; km = kilometer; N = nitrogen.
Notes: the Verna Wellfield National Atmospheric Deposition Program/National Trends Network monitoring site, FL41, was chosen to
characterize long-term wet deposition of N species, because it is closest to the Tampa Bay study area. The site is located at the
southern edge of the study area in Sarasota, FL and is shown as the grey dot on the maps.
Source: data shown in the figures were obtained from the hybrid modeling/data fusion product, total deposition,
http://nadp.sws.uiuc.edu/committees/tdep/tdepmaps/. and described earlier in Appendix 2.7.
Figure 16-27 (A) Wet and dry nitrogen deposition in Tampa Bay and the
surrounding area. (B) Percentage of oxidized nitrogen deposition
in Tampa Bay and the surrounding area.
16.4.3.	Long-Term Ecological Monitoring
1	Tampa Bay has been intensively studied over the years due to its significance as a natural
2	resource and its susceptibility to degradation from eutrophication. There is monthly
3	ambient water quality data going back to 1972 (Sherwood et al.. 2016). Phytoplankton
4	biomass and primary production have been assessed regularly in the bay from the 1980s
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N Deposition (kg-N/ha)
0	Monitor FL#41
0 Monitor Locations
1	I Tampa Bay Study Area
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and benthic organisms since 1993 (Greening et al.. 2014). Aerial photography of the
Tampa Bay shoreline is available from the early 1950s showing the extent of seagrass
coverage prior to rapid human population increases in the watershed (Greening et al..
2014). A consistent monitoring program of seagrass extent (every 2 year) was started in
1988 (Greening et al.. 2014).
Eutrophication symptoms were observed in Tampa Bay as early as the 1970s. These
included phytoplankton and macroalgal blooms and areas of low dissolved oxygen
(Greening et al.. 2014). According to Bricker et al. (2007). algal blooms were common in
the 1970s and caused foul odors and aesthetic impairment. In the late 1970s, hypoxic and
anoxic conditions along the shoreline of the most urbanized segment of Tampa Bay
caused extensive mortality of benthic organisms in the late summer months (Greening et
al.. 2014). Fish kills have also been observed in and near Tampa Bay. Historically, the
most obvious symptom of eutrophication in Tampa Bay was the loss of light at depth due
to elevated algal biomass. The loss of SAV from reduced light availability was dramatic.
In 1950, approximately 16,000 hectares of SAV were present. By the early 1980s, over
half of this area had been lost rFigure 16-28 (Bricker et al.. 2007; Haddad. 1989)1.
Tampa Bay is somewhat unique in that the bay waters have historically exhibited a high
level of P from local phosphate deposits, P mining, and fertilizer manufacturing and
shipping operations. The high phosphorus concentrations have resulted in an unusually
low N:P ratio in comparison to other estuaries, making N the limiting nutrient for
phytoplankton growth (Greening et al.. 2014; Greening and Janicki. 2006). Numerous
nutrient limitation studies using natural Tampa Bay phytoplankton populations and
cultured test organisms have confirmed N limitation (Greening et al.. 2014).
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f

t c—S
j ^ 1 rl
\V ^ # n
% ,
\ .\
*3r
4
Mi vv
¦
Ł
V. * TTpa
\2 a, Bay f>
\J0y J)
u Jt{
*?c
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( ^ ' ' VNyv/
v f 1 ^
. *»J-V.'
*>t N 	 i
H
¦4*
10 km

25
iim
km = kilometer; mi = mile.
Notes: red indicates area of seagrass iost from 1950 to 1990.
Source: Bricker et al. (2007).
Figure 16-28 Seagrass Loss in Tampa Bay.
16.4.3,1. Indicators of Enrichment and Eutrophication in Tampa Bay
Indicators of historical water quality trends in the bay include chlorophyll a concentration
and seagrass coverage (Greening et al.. 2014).
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16.4.3.1.1. Chlorophyll a
1	Phytoplankton biomass as measured by chlorophyll a concentration is one of the most
2	effective indicators of eutrophication in the bay (Sherwood et al.. 2016). Chlorophyll a is
3	directly linked to nutrient inputs and has been measured in the bay since 1972. Increased
4	algal biomass results in reduced light penetration in the water, thereby harming seagrass
5	and ultimately resulting in the loss of SAV area.
6	The TBEP has developed water quality models to quantify linkages between N loads and
7	bay water quality. N is generally the primary limiting nutrient, and chlorophyll a
8	variation in the bay responds most significantly to watershed TN loads ITigure 16-29
9	(Greening and Janicki. 2006)1.
Tampa Bay Estuary Program
Nitrogen Loading - Chlorophyll a Relationship
Comparison of Predicted and Observed Chlorophyll a Concentrations

Predicted



40


86-98
99-07
REF
30

+

20

4*4 *o yf -f o ~
D o i>p9, ~ Q,e b
*¦ a 00 % ' * P ~ ~ 10
~ % at J* - t , . +*
\ * ° o

10
• >
' f*n> '3 •
ikdwEBKafip**- >*o 0
iB& v-1 *

0



3
10 20 30
Observed
40
Source: Janicki Environmental (2011).
Figure 16-29 Comparison of observed chlorophyll a and that predicted from
the total nitrogen load—chlorophyll a relationships for all four
mainstem Tampa Bay segments, for 1986-1998 and 1999-2007.
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Data show that, with the exception of the Old Tampa Bay segment in 2009 and 2011,
chlorophyll a thresholds were not exceeded in all four major bay segments over the
2007-2011 period (TBEP. 2012a). The chlorophyll a threshold values are used as
indicators of impairment and vary slightly for each different segment of the bay, ranging
from 5.1 |ig/L for Lower Tampa Bay, to 8.5 |ig/L for Middle Tampa Bay, 9.3 |ig/L for
Old Tampa Bay, and 15 |ig/L for Hillsborough Bay [expressed as annual averages;
(TBEP. 2012aYI.
Seagrass Coverage
Five species of seagrasses, turtle grass (Thalassia testudinum), manatee grass
(Syringodium filiforme), shoal grass (Halodule wrightii), star grass (Halophila
engelmannii), and widgeon grass (Ruppia maritima), are found in Tampa Bay (TJSGS.
2011). These shallow-water grasses need sufficient light for photosynthesis and are
sensitive to reduced water clarity associated with increased algal growth. Scientists have
been tracking seagrass coverage as far back as the 1950s in Tampa Bay. Between the
1950s and 1980s seagrass coverage declined by about 50% or 9,000 ha in conjunction
with decreased water quality and physical impacts to the bay such as dredging and filling
(Poor etal.. 2013a; Haddad. 1989). Between 1999 and 2010, bay water clarity improved,
chlorophyll a concentrations decreased, and seagrass coverage area steadily increased,
coinciding with the reductions to N inputs to the bay rFigure 16-30; (TBEP. 2015; Poor et
al.. 2013a; TBEP. 2011; Greening and Janicki. 2006)1. The most recent surveys of
seagrasses (2014) exceeded the recovery goal set in 1995 [95% of estimated seagrass
coverage of the 1950s; (Sherwood et al.. 2016)1.
Through management actions, including an agreement with a power company to install
air pollution control systems and the completion of more than 250 projects to reduce N
inputs from the watershed, TN loadings to the bay have declined from previous annual
average estimates despite an ever-increasing population in the Tampa Bay metropolitan
area (Figure 16-31). Data compiled by the TBNMC indicate that continuing efforts to
reduce N loading are resulting in more-than-sufficient water quality for the expansion of
seagrasses (TBEP. 2012a).
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CO 18
O
"C—
x 16
CD
03 14 H
>
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(/)
CO
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2000 -I	T	1	T	,	T	|	T	,	.	,	T	,	T	,	T	1	.	,	T	,	T	,	T	,	T	,	.	1" 1.5
1984 1986 1988 1990 1992 1994 1996 1998 2000 2002 2004 2006 2008 2010 2012
Year
3750
-Hydrologically Normalized TN Loads
-Tampa Bay Metropolitan Area Population
2.75
3000
2750
O
2.25 -O
TN = total nitrogen; yr = year.
Source: TBEP (2012a).
Figure 16-31 Trend in hydrologically normalized total nitrogen load to Tampa
Bay relative to population increases in the Tampa Bay
metropolitan area.
1	Russell and Greening (2015) estimated a potential relative cost savings of $22 million per
2	year in avoided wastewater treatment plant costs due to nutrient reductions associated
3	with restoration and recovery of seagrass, marsh, and mangrove habitats in Tampa Bay.
4	Ecosystem services benefits of over $365 million estimated over an 18-year period from
5	1990 to 2008 include C sequestration and nutrient reductions via denitrification.
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16.4.4. Nitrogen Management
16.4.4.1. Numeric Nutrient Criteria
In 1998, the U.S. EPA published the National Strategy for the Development of Regional
Nutrient Criteria to promote the use of nutrient concentration levels in state water quality
standards (U.S. EPA. 1998b). Historically, Florida had a narrative nutrient water quality
criterion in place to protect waters against nutrient enrichment. In 2011, the state adopted
the first set of statewide numeric nutrient standards for Florida's waters. By 2015, almost
all of the remaining waters in Florida had numeric nutrient standards. Numeric nutrient
criteria in Florida are established for all estuary segments and include criteria for TN,
total phosphorus (TP), and chlorophyll a. For open ocean coastal waters, numeric criteria
are derived from satellite remote sensing techniques and established for chlorophyll a
only. Numeric nutrient criteria covering all of the mainstem segments and estuarine
segments in Tampa Bay were approved by the U.S. EPA in November 2012 (U.S. EPA.
2012c). A map of the bay segments is provided in Figure 16-23. The numeric nutrient
criteria for chlorophyll a for the mainstream segments of Tampa Bay are shown in
Table 16-19.
Table 16-19 Numeric nutrient criteria for chlorophyll a for the four mainstem
segments of Tampa Bay adopted by the Florida Department of
Environmental Protection.
Bay Segment
Chlorophyll a(|jg/L)
Old Tampa Bay
9.3 as annual mean
Hillsborough Bay
15.0 as annual mean
Middle Tampa Bay
8.5 as annual mean
Lower Tampa Bay
5.1 as annual mean
L = liter; |jg = microgram.
Modified from Sherwood et al. (2016).
Tampa Bay's numeric nutrient criteria for TN (Table 16-20) is expressed as tons/million
cubic meters of water as an annual total not to be exceeded more than once in a 3-year
period and represents an unorthodox approach to developing nitrogen nutrient criteria
(Florida DEP. 2016; Sherwood et al.. 2016). These values are based on previously
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adopted total maximum daily loads (TMDLs) or nutrient targets developed by TBEP.
Usually, the U.S. EPA's numeric nutrient criteria are expressed as a concentration, but
TMDLs are typically expressed as loads. The N management strategy for Tampa Bay
linked seagrass growth to external N loadings based on evidence that TN loads, resultant
chlorophyll and light attenuation, and seagrass response were closely correlated
(Sherwood et al.. 2016; Janicki Environmental. 2011). The TBNMC also found that when
freshwater inputs are greater, TN moves through the system more quickly; therefore,
residence time was an important consideration. TN loads and hydrologic loads affect both
the chlorophyll within the bay, and the amount of TN delivered per unit water was used
to set the numeric nutrient criteria for Tampa Bay. The FDEP and the U.S. EPA both
agreed that this framework, although unique for numeric nutrient criteria, would provide
reasonable assurance that the state water quality criteria would be met in Tampa Bay
(Sherwood et al.. 2016; Greening et al.. 2014).
Table 16-20 Numeric nutrient criteria for total nitrogen for the four mainstem
segments of Tampa Bay.
Segment
Nitrogen Delivery Ratio Threshold (tons/million m3)
Old Tampa Bay
1.08
Hillsborough Bay
1.62
Middle Tampa Bay
1.24
Lower Tampa Bay
0.97
m = meter; TBEP = Tampa Bay Estuary Program; TN = total nitrogen.
TBEP estuarine numeric nutrient criteria expresses as nitrogen delivery ratio (tons TN per million m3 hydrologic load) based on
1992-1994 conditions.
Source: Florida PEP (2016). Janicki Environmental (2011).
16.4.4.2. Nitrogen Management Consortium Efforts
The TBEP was established in 1992 as a part of the U.S. EPA's National Estuary Program
(Sherwood et al.. 2016). TBEP coordinates TBNMC, a public-private partnership
working to reduce N loading to the bay and whose members include 40+ local
governments (Tampa, St. Petersburg, and others) along with key industries bordering the
bay, such as electric utilities, fertilizer manufacturers, and agricultural operations.
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In recent years, one focus of the TBNMC's strategy has been to allocate N loads for
major point and nonpoint sources in order to meet state and federal requirements. The
governments and industries participating in the TBNMC voluntarily committed to cap
their N loads at average annual levels recorded in 2003-2007, which also meets the 1998
federally recognized TMDL for N. These capped allocations have been adopted by the
state of Florida and incorporated into NPDES discharge permits (TBEP. 2014). One
example of the Consortium's work is the reconfiguration of coal-fired plants in the
Tampa Bay watershed to reduce NOx emissions (Poor et al.. 2013a).
Every 5 years, the TBNMC must submit a Reasonable Assurance Update document to
FDEP for approval to document progress toward water quality and seagrass management
goals. The most recent reasonable Assurance Update (2012) presented data to show that
reasonable progress has been made towards the attainment of designated uses of Tampa
Bay. Based on this conclusion, FDEP upgraded Hillsborough Bay and Old Tampa Bay
segments to assessment Category 4b for nutrients (adequate management in place) from
Category 5 (impaired), and segments in Lower Tampa Bay were moved from
Category 4b to Category 2 (attains standards), signifying the state's recognition of
significant improvements in water quality and seagrass expansion (TBEP. 2014). The
Consortium members share the cost of the data collection and analysis to prepare the
Reasonable Assurance Update document, and the next update will cover progress made
during the 2013-2017 time period.
16.4.4.3. Response to Nitrogen Management Strategies
A recent paper reviewing the response of Tampa Bay to N management strategies
presented evidence of the ecosystem's recovery constituting a "regime shift" trajectory
(Duarte et al.. 2015; Duarte et al.. 2009). Since the mid 1980s Tampa Bay has shifted
back to a clear water seagrass system from a phytoplankton-dominated system that was
characteristic of the bay under previously eutrophic conditions. Following recovery from
an El Nino heavy rainfall period (1997-1998), water clarity in Tampa Bay increased
significantly, and seagrass expanded at a rate significantly different than before the event,
suggesting a feedback mechanism observed only in a few other systems. The authors
observed that too often these estuarine ecosystems are not able to recover even when
eutrophication stressors are mitigated because of the concurrent changes in environmental
conditions, that affected ecosystem dynamics occurring over the time period from the
onset of eutrophication to the reduction of nutrient levels. (Greening et al.. 2014) noted
four specific elements of the Tampa Bay management strategy that have resulted in a
successful recovery trajectory: (1) development of numeric water quality targets in
consultation with stakeholders, (2) citizen involvement in calling for action from
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regulatory agencies, (3) collaborative actions such as the TBNMC, and (4) state and
federal regulatory programs. BRACE scientists concluded that declines in atmospheric N
deposition to the estuary and its watershed were likely part of the historical improvement
in water clarity and increases in seagrass acreage (Poor et al.. 2013a; TBEP. 2011).
16.5. Rocky Mountain National Park Case Study
16.5.1. Background
This case study describes the environment and ecology of Rocky Mountain National Park
(ROMO) and summarizes research that provides insight into the impacts of atmospheric
nitrogen and sulfur deposition on terrestrial and aquatic ecosystems in ROMO. This
includes atmospheric deposition research from within ROMO, the surrounding area of the
Rocky Mountains, as well as other portions of the Northwestern Forested Mountains
ecoregion. Although the focus is on research published since 2008, publications prior to
2008 are used to establish context and identify long-term trends.
16.5.1.1. Description of Case Study Region
Rocky Mountain National Park (ROMO) was established by Congress in 1915 and is
located in the Front Range of the Rocky Mountains in north-central Colorado, about
80 km northwest of Denver. Straddling the Continental Divide, the western side of
ROMO serves as the headwaters for the Colorado River and eastern slope feeding the
South Platte River Basin (Wolfe et al.. 2003). The land within ROMO contains steep
elevation gradients and diverse ecosystem types, ranging from talus slopes at high
elevations and permanent glaciers above 4,000 m to grassy meadows at 2,400 m. Over
25% of ROMO is above the tree line (about 3,500 m), and ROMO contains large areas of
alpine tundra. In addition, there are more than 60 peaks above 3,600 m in elevation
(Porter and Johnson. 2007). Ecologically, ROMO is part of the Northwestern Forested
Mountains ecoregion (Figure 16-32).
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Source: Clarke (2013V
Figure 16-32 Rocky Mountain National Park ecosystems.
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The ecoregion hosts extremely diverse vegetation and well-defined community zones
along elevation gradients. Winter snow cover averages 1.5 m annually in the Rocky
Mountains, and is a strong determinant of both plant and soil biodiversity (Clow ct al..
2016). More than 1,000 species of vascular plants have been documented in ROMO
(NPS. 2013). Within the alpine tundra, plant communities contain willow (Salix spp.) and
a variety of grasses, sedges, and forbs such as alpine sunflower (Rydbergia grandiflora)
and alpine forget-me-nots (Polemonium viscosum). The subalpine zone is composed
primarily of forests, dominated by Engelmann spruce (Picea engelmannii), subalpine fir
(Abies lasiocarpa), and limber pine (Pinus flexilis), whereas forests in the montane zone
vary from lodgepole pine (Pinus contorta) at cold high elevation sites, trembling aspen
(Populus tremuloides) in moist sites, and Douglas fir (Pseudotsuga menziesii) and
ponderosa pine (Pinus ponderosa) at more moderate and xeric sites, respectively
(Beidleman et al.. 2000).
More than 270 species of birds, including migratory species, have been reported in this
area over the last 100 years. Many are unique to mountainous habitats—aspen, ponderosa
pine, high-elevation willow, and spruce-fir vegetation types—found in the southern
Rocky Mountains. Seven native fish and four exotic fish inhabit the aquatic systems of
ROMO. Sixty-seven mammal species are known to be native to the area, but grizzly
bears (Ursus arctos), gray wolves (Canis lupus), and bison (Bison bison) were locally
extirpated in the 19th and early 20th centuries. The lynx (Lynx canadensis) and wolverine
(Gulo gulo) are either extirpated or extremely rare. Moose (Alces alces) are now
common, but were not historically recorded as part of this area of the Rocky Mountains.
A select group of four amphibian and two reptile species survive the high elevations and
cold temperatures in ROMO. They are all considered species of concern due to apparent
low numbers, lack of information about their status, and/or declining population trends.
There are 141 confirmed species of butterflies, but arthropod (insects, spiders, centipedes,
etc.) communities in ROMO have not been as well documented as other taxonomic
groups.
The NPS maintains a list of species known to occur in ROMO that are considered
endangered, threatened, or candidates for protection by the Endangered Species Act
(NPS. 2013):
•	Boreal toad (Bufo boreas boreas)—candidate
•	Yellow-billed cuckoo (Coccyzus americanus)—candidate
•	Canada lynx, (Lynx canadensis)—threatened
•	Greenback cutthroat trout (Oncorhynchus clarki stomias)—threatened
Pardo et al. (2011c) reported that responses to elevated N deposition in the Northwestern
Forested Mountains ecoregion include alteration of soil carbon:nitrogen (C:N) ratios,
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base cation composition, and accelerated N cycling rates, including increased
mineralization and nitrification. Further changes described in this case study and in other
Appendices include plant, lichen, and algal chemistry; surface water chemistry (including
N concentration and acid neutralizing capacity); catchment N leaching rate; and changes
in the community composition of plants, lichens, and phytoplankton.
16.5.1.2. Class I Areas
Rocky Mountain National Park is designated as a federal Class I area. The Clean Air Act
(42 USC 7470) authorized these Class I areas to protect air quality in national parks over
6,000 acres and national wilderness areas over 5,000 acres in an effort to preserve pristine
atmospheric conditions. Class I areas are subject to the "prevention of significant
deterioration (PSD)" regulations under the Clean Air Act (42 USC 7470). New and
modified existing air pollution sources are required to have PSD preconstruction permits
and air regulatory agencies are required to notify federal land managers (FLMs) of any
PSD permit applications for facilities within 100 km of a Class I area.
16.5.1.3. Regional Land Use and Land Cover
The land area within the ROMO region falls predominantly into four land cover
classifications: evergreen forest, barren land, herbaceous cover, and perennial snow/ice
(Figure 16-33). The remainder of the land coverage categories account for less than 7%
of the total area. The area immediately surrounding ROMO is largely forested.
Regionally, the land to the north, west, and south of ROMO is predominantly forested,
with some grasslands used as grazing land. The region to the east and southeast contains
more grassland, agricultural, and urban land use. The smaller towns of Estes Park,
Allenspark, and Grand Lake are close to the boundaries of ROMO (Figure 16-34). while
the larger population centers of Fort Collins and the Denver metropolitan area are further
to the east and southeast, respectively (Figure 16-33 inset).
There are 29 total watersheds at the U.S. Geological Survey level-12 hydrological unit
code (HUC12s) scale that fall completely or partially inside ROMO (Figure 16-34).
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Rocky Mountain National Park
Case Study Location
Land Cover Classification

•Evergreen Forest (52.7%)
	| "Barren Land (13,7%)
1 1
Hay/Pasture
I Deciduous Forest
1 1
"Herbaceuous (13.8%)
Developed, Low Intensity

Open Water
| Developed, Medium Intensity
i i
•Perennial Snow/Ice (12.6%)
| Developed, Open Space
i—in
Shrub/Scrub
Emergent Herbaceuous Wetlands
I I
Woody Wetlands
"Over 90% of coverage from these four
categories
m
Estes Park
m
Fort Collins
80 Miles
J a c k s o n
County
Routt National
Forest
Glen Haven
Grand Lake
COL
Jamestown
Boulder
Gr
Co u n
Aliens park
Lakewood
Denver
Figure 18-33 Rocky Mountain National Park land coverage using the land cover
classifications as mapped by the National Land Cover Dataset.
Percentage of cover is shown for the four dominant cover types.
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Rocky Mountain National Park
Case Study Location
HUC12 Boundaries
Inside and outside of the park boundary
©
Allenspark
g Arapaho National
Forest
D Miles
Jackson
County
Grand Lake
HUC = hydrologic unit code.
Figure 16-34 Rocky Mountain National Park hydrologic unit code 12
watersheds.
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16.5.2.
Deposition
Based on TDEP calculations (see Appendix 2 of the ISA), most of ROMO is estimated to
receive deposition between 3-9 kg N/ha/yr and 3-6 kg S/ha/yr (Figure 16-35). In
contrast to the low rates observed throughout much of the western U.S., rates of N
deposition within ROMO are relatively high, and areas to the east of ROMO experience
deposition rates similar to areas in the north-central U.S. (see Figure 16-35 N deposition
inset). However, the estimated rates of S deposition are consistent with the low rates
observed elsewhere in the western U.S. and considerably lower than the S deposition
rates observed within the Ohio Valley. Within these TDEP estimates, the inorganic N
deposition in ROMO is relatively evenly balanced between oxidized and reduced forms,
with the northeastern portion of the region receiving predominantly reduced forms of N
(Figure 16-36). There are three National Atmospheric Deposition Program (NADP)
measurement sites within ROMO (Figure 16-36). including measurements stretching
back more than 30 years at the Loch Vale and Beaver Meadows sites. Over the past
25 years, sulfate (SO42 ) and hydrogen ion (H+) deposition have declined slightly at the
Beaver Meadows site (CO 19), whereas deposition of ammonium (NFLO and nitrate
(NO3 ) has been relatively constant (Figure 16-35). See Appendix 2.4. Appendix 25_ and
Appendix 2jS for more information on deposition in the U.S. Other maps showing the
contributions of individual species to dry and/or wet deposition are given in
Appendix 2.7.
Atmospheric N deposition generated from anthropogenic sources is a significant
influence on many ecosystems within ROMO (Wolfe et al.. 2003; Baron et al.. 2000;
Williams and Tonnessen. 2000; Williams et al.. 1996a; Caine. 1995). For this reason,
atmospheric N deposition has been the focus of considerable research, including the
Rocky Mountain Airborne Nitrogen and Sulfur (RoMANS) study (Beem et al.. 2010).
The RoMANS study was designed to characterize the sources as well as the transport,
transformation, and deposition processes of oxidized S and oxidized and reduced forms
ofN (Beem et al.. 2010).
Primary atmospheric transport to ROMO is from the west and northwest. However,
atmospheric circulation patterns along the Colorado Front Range can be complex, and
significant quantities of precipitation can fall on the east side of ROMO when easterly
upslope events occur (Porter and Johnson. 2007; Mast et al.. 2003; Baron and Denning.
1993). For instance, although prevailing pattern of atmospheric circulation brings
relatively clean air from west of ROMO, anthropogenic N can be transported to ROMO
from urban and agricultural areas along the Front Range when convective heating over
the plains produces easterly upslope winds (Figure 16-37). Higher than 3,000 m above
sea level (masl), westerly air flow dominates over the upslope pollution (Sievering et al..
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1996). but coal-fired power plants located in Hayden and Craig, CO approximately
150 km west of RQMO, emit NOx that may reach ROMO with prevailing westerly
winds.
Annual Wet Deposition and 3-Year Moving Average at
Site C019: 1990-2014
200
180
160
hi 140
-C
"5 120
E
\ 100
c
,2 80
s «>
(X
OJ 40
Q
20
•		-	• T
1988	1992	1996	2000	2004
Year
2008
2012
2016
H+ = hydrogen ion; ha = hectare; kg =ki!ogram; mol = mole; N = nitrogen; NhV = ammonium; N03 = nitrate; RMNP = Rocky
Mountain National Park; S = sulfur; S042" = sulfate; yr = year.
Figure 16-35
Spatial patterns of atmospheric nitrogen and sulfur deposition in
the Rocky Mountain National Park region based on TDEP
calculations averaged from 2011-2013 (see Appendix 2) and
long-term trends in wet atmospheric deposition from the Beaver
Meadows National Atmospheric Deposition Program Monitoring
site within Rocky Mountain National Park.
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~ Ouflv^r. CQ
O Monitor CO * 19
0 Monitor Location*
HMNF* Sludry Area
m
-	tri&
CoibCiKfo-
m,
%. 0»^.x*d W 0^pc-»ir<.n
¦¦mi r Hm
fJ*J*J* ** jf
~	D»flv*r,CO
O Moniflor CO * 19
#	Monitor Location*
Li RMNP Slucfy Area
m
0
_
Co4of»ab
% W«E Ckp-xi' .:-'.
//////
N = nitrogen; RMNP = Rocky Mountain National Park.
Figure 16-36
Spatial patterns in the percentage of atmospheric nitrogen
deposited in oxidized forms and through wet deposition in the
Rocky Mountain National Park region based on TDEP calculations
averaged from 2011-2013 (see Appendix 2).
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IIKIShSl O'cLu
^^^litroger^ource^
chemical
conversion
and dispersion
C Transport I Transformation
(i
^^Dcposttio^^Foedbaci^TB
Particle and Gas
Transformations
Effects
^Vis|b|lit^ ^Otofw^ ^Cjjmate^Ownj^
Wet Deposition
Dry DcposHiOf
lightning
VOWWWI
ami olbor
area sources
ptents
urban and
sources
Surface and Ground
Water Pollution t-, ^ ^
*%SfdB
Ecosystem
Effects
Soil and
Natural
Vegetation
Fertilizer and
Feedlot Chemistry
Intro.
HN03 = nitric acid; NH3 = ammonia; NH4+ = ammonium; NO = nitric oxide; N02 = nitrogen dioxide; NCV = nitrate; N0X = NO + N02
Source: Clarke (2013).
Figure 16-37 Rocky Mountain National Park nitrogen cycle.
1	Numerous organic and inorganic N species contribute to total N deposition in ROMO.
2	Lee et al. (2016) estimated wet and dry deposition of reactive N species to ROMO using
3	a GEOS-Chem adjoint model, while Benedict et al. (2013) conducted a year-long survey
4	of the chemical speciation of N deposition at ROMO (Figure 16-38). Wet deposition of
5	inorganic N, particularly NH/, was the dominant form of N delivered to the surface at
6	ROMO. Wet deposition of organic nitrogen (ON) was also a major component of
7	deposition. Combined, reduced forms of N (wet and dry N H4 . dry NHj, ON) comprised
8	a large majority of the total N deposition flux in these measurements (Benedict et al..
9	2013).
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14






Dry HNO.
Dry NU, V
Dry NO ¦
Dry PON ¦
Dry NO, W	










DO 0.2 0.4 0.6 OB 10 1.2 1.4
Total deposition (kg N/ha)
ha = hectare; HN03 = nitric acid; kg = kilogram; N = nitrogen; NH3 = ammonia; NH4+ = ammonium; N02 = nitrogen dioxide;
N03 = nitrate; ON = organic nitrogen; PON = particulate organic nitrogen.
Source: Benedict et al. (2013).
Figure 16-38 Total wet and dry deposition of nitrogen components measured at
Rocky Mountain National Park from November 2008 to November
2009, including organic nitrogen and particulate organic nitrogen.
Modeling conducted by Gebhart et al. (2011) suggests that the majority of the ammonia
measured in ROMO during the RoMANS study was emitted within Colorado. Modeling
conducted by (Lee et al.. 2016) suggest that reduced N deposition comes mostly from
sources east of the park, and oxidized N deposition comes primarily from sources west of
the park. Major emission sources of N to deposition in ROMO include ammonia from
livestock and oxidized N from mobile sources (Figure 16-39).
Aeolian dust deposition from the Colorado Plateau mitigates some of the acidifying
effects of S andN deposition in ROMO. Between 1993 and 2014, dust deposition
increased by as much as 81% in the southern Rocky Mountains. Clow et al. (2016)
estimated that deposition of dust has altered snowpack chemistry by increasing its
alkalinity (annual change of 0.32 |acq Ca2+/L/yr) over the same time period in which
smaller magnitude decreases in acid deposition to the snowpack (annual changes of
-0.10 |_ieq S0.f7L/yr and -0.07 |acq NO;, /L/yr) have occurred.
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Emission sectors
NH3 (livestock)
[ NH3 (fertilizer)
^ NH3 (natural)
U NOx (surface inventory)
| NOx (electric generating units)
| NOx (non-electric generating
industrial stacks)
~| NOx (aircraft)
	NOx (lightning)
NOx (soil)
Geographic footprint of Nr deposition
» 0.03 0.(
[kg N/ha/yr]
RM (x2) 4.
Source: adapted from Figure 5 in Lee et al. (2016).
Figure 16-39 Annual-averaged monthly footprint of reactive N deposition in
Rocky Mountain National Park(4.0 kg N/ha/yr), and pie chart of
fractional contribution from emission sectors, as estimated by
GEOS-Chem adjoint model.
16.5.3.	Critical Loads and Other Dose-Response Relationships
1	This section focuses on studies that describe dose-response functions and/or critical loads
2	(CL) for atmospheric deposition effects on ecological endpoints. Terrestrial effects are
3	discussed first, followed by aquatic effects and a discussion on integration.
16.5.3.1. Terrestrial Critical Loads and Dose-Response Relationships
4	Among terrestrial ecosystems impacted by N deposition, alpine ecosystems are
5	particularly sensitive to increased availability of N due to inherently low rates of N
6	cycling, low rates of primary production, and thin, poorly weathered soils (Fenn et al..
7	1998). Among taxonomic groups in this ecoregion, the impact of atmospheric deposition
8	on epiphytic lichens is concerning because their strong dependence on atmospheric
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deposition for nutrients and unique symbiotic physiology makes them highly vulnerable
to N and S pollution and because lichens are important contributors to ecosystems
services (Brodo et al.. 2001). Among lichens, the structure and physiology of oligotrophic
species makes them especially important components of winter food webs, hydrologic
and nutrient cycles, and wildlife habitat (McCunc and Geiser. 1997).
16.5.3.1.1. Empirical Studies
In the Pardo et al. (201 lc) critical loads assessment, most of the studies examining the
responses of alpine ecosystems were conducted in the Colorado Front Range of the
Rocky Mountains, either in ROMO or in adjacent ecosystems such as the Niwot Ridge
Long-Term Ecological Research site. Critical load values for these alpine ecosystems
ranged from 1.2 to >20 kg N/ha/yr, excluding two values from Fenn etal. (2008) which
were determined for southwestern forested ecosystems in California (Table 16-21).
Table 16-21 Terrestrial empirical critical loads of nutrient nitrogen for the
Northwestern Forested Mountains ecoregion.
CL





Reliability

(kg N/



Deposition/

in Pardo et

ha/yr)
Species
Response
Method
Addition
Site
al. (2011c)
Reference
1.2 to
Epiphytic
Lichen
Application of
Total
Coastal
(#)
Geiser et al.
3.7
lichens
community
western
deposition: 0.8
Alaska

(2010)

(150 species)
change in
Oregon and
to





mixed-conifer
Washington
8.2 kg N/ha/yr





forests
model
(CMAQ)



2.5 to
Epiphytic
Lichen
Fenn:
Fenn:
Sierra
##
Fenn et al.
7.1
lichens
community
empirical CLs
Inorganic N
Nevada and

(2008)

Fenn: Letharia
change in
and DayCent
throughfall: 1.4
San

Geiser et al.

vulpina
mixed-conifer
modeling
to
Bernardino

(2010)

Geiser:
forests
Geiser:
71.1 kg N/ha/yr
Mountains



150 species

application of
Geiser: Total





western
deposition: 0.8






Oregon and
to 8.2 kg






Washington
N/ha/yr






model
(CMAQ)



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Table 16-21 (Continued): Terrestrial empirical critical loads of nutrient nitrogen
for the Northwestern Forested Mountains ecoregion.
CL
(kg N/
ha/yr)
Species
Response
Method
Deposition/
Addition
Site
Reliability
in Pardo et
al. (2011c) Reference
Subalpine
forest
(Engelmann
spruce [Picea
engelmannii])
Increase in
organic horizon
N, foliar N,
potential net N
mineralization,
and soil
solution N,
initial increases
in N leaching
below the
organic layer
Baron:
Baron:
CENTURY
modeled total
model
deposition: 0.2
Rueth and
to 16.0 kg
Baron: field
N/ha/yr
sampling east
Rueth and
and west
Baron: total
slope forests
deposition: 3.2

to 5.5 kg

N/ha/yr (for

1992-1997)
N addition
Ambient:
experiment
6 kg N/ha/yr;

Addition: 20,

40, or

60 kg N/ha/yr
ROMO east
and west of
Continental
Divide
##
Baron et al.
(1994) Rueth
and Baron
(2002)
4 to 10 Alpine dry
meadow
(Carex spp.,
including Carex
rupestris)
Plant species
composition
change
Niwot
Ridge, CO
##
Bowman et
al. (2006)
Carex rupestris
Increase in
vegetation
cover
N addition
experiment
Total
deposition:
6 kg N/ha/yr
N additions:
20, 40, or
60 kg N/ha/yr
Niwot
Ridge, CO
##
Bowman et
al. (2006)
5 to 10
Ectomycorrhizal
Ectomycorrhizal
Expert
2002: bulk N
Kenai
fungi
fungi
judgment
deposition
Peninsula,
(-40 species)
community
extrapolated
across five
AK

structure
from marine
sampling sites


change in
west coast
ranged from


spruce (Picea)
spruce forest
0.15 to


forests

2.3 kg/ha/60
days
2008: Wet
deposition 2.8
to 7.9 kg
N/ha/yr

Alpine
Soil NOs"
N addition
Total
Niwot
terrestrial
leaching and N
fluxes

deposition:
6 kg N/ha/yr
N additions:
20, 40, or
60 kg N/ha/yr
Ridge, CO
(#)
Lilleskov
(1999)
Lilleskov et
al. (2001)
Lilleskov et
al. (2002)
Lilleskov et
al. (2008)
>20
Bowman et
al. (2006)
CL = critical load; CMAQ = Community Multiscale Air Quality model; ha = hectare; kg = kilogram; N = nitrogen; N03 = nitrate;
ROMO = Rocky Mountain National Park; yr = year.
Reliability rating: ## = highly reliable—a number of published papers show comparable results; # fairly reliable—the results of some
studies are comparable; (#) expert judgment—few empirical data are available, Critical Load based on expert judgment of those
ecosystems.
Source: (Pardo et al.. 2011c).
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1	Since the publication of Pardo etal. (2011c) assessment, there have been several new
2	studies on terrestrial CLs (Table 16-22). New studies find CLs for soil N concentration
3	changes and leaching at 9.0-14.0 kg N/ha/yr, CLs to protect lichens at 4.0-4.1 kg
4	N/ha/yr, and CLs to protect vascular plant biodiversity (species richness, abundance, or
5	community composition) at 1.0-8.7 kg N/ha/yr.
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Table 16-22 Montane forest and alpine ecosystem critical loads for nitrogen
deposition research published since the critical load assessment by
Pardo et al. (2011c).
CL (kg
N/ha/yr) Vegetation Response	Method Deposition/Addition Site	Reference
1.0
Composite
5% change in
ForSAFE-VEG
Deposition:
Northern and
SverdruD et al.

montane plant
alpine and
modeled
~4 kg N/ha/yr,
central
(2012)

communities
subalpine plant
change in
1 kg S/ha/yr
Rocky



community
future (to 2500)

Mountains



composition
vegetation

region

1.9
Salix Candida,
10% change in
ForSAFE-VEG
Deposition:
Loch Vale
McDonnell et al.

Carex spp.,
treeline
modeled
3.5 kg N/ha/yr
watershed,
(2014a)

Abies
(sub/alpine)
change in

ROMO


lasiocarpa,
plant
future (to 2100)




Geum rossii
community
vegetation





composition




3.0
Carex
Alpine
Field addition
Ambient deposition:
ROMO
Bowman et al.

rupestris
vegetation
study
4 kg N/ha/yr

(2012)


abundance

Addition: 5, 10, and






30 kg N/ha/yr.


4.0
Epiphytic
Degradation of
Empirical CLs
1.83 to
Northern
McMurrav et al.

lichens
lichen

3.45 kg N/ha/yr
Rocky
(2015)


communities

(CMAQ)
Mountains

4.1 Epiphytic Poorerthallus Empirical CLs Throughfall	Wind River McMurrav et al
lichens	condition	<0.9 to 4.1 kg N/ha/yr Range, WY, (2013)
L vulpina	including the
place I
U' ^ppon/ca	Brjdger
Wilderness
8.7
Forbs and
grasses—
open canopy
ecosystem
Species
richness loss
Empirical CLs
from vegetation
surveys
1 to 19 kg N/ha/yr,
wet (NADP) + dry
(CMAQ) deposition.
Over
15,000 plots
over the
continental
U.S.
Simkin et al. (2016)
9.0
Alpine dry
Soil solution
Field addition
Ambient deposition:
ROMO
Bowman et al.

meadow
NOs"
study
4 kg N/ha/yr

(2012)


concentrations

Addition: 5, 10, and






30 kg N/ha/yr


14.0
Alpine dry
meadow
Soil dissolved
inorganic N
concentrations
Field addition
study
Ambient deposition:
4 kg N/ha/yr
Addition: 5, 10, and
30 kg N/ha/yr
ROMO
Bowman et al.
(2012)
CL = critical load; CMAQ = Community Multiscale Air Quality model; ForSAFE-VEG = a dynamic forest ecosystem model; ha = hectare;
kg = kilogram; N = nitrogen; NADP = National Atmospheric Deposition Program; N03" = nitrate; ROMO = Rocky Mountain National
Park; S = sulfur; yr = year.
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35
36
Modeling Studies
Elevated lake and stream NOs concentrations have been observed in ROMO since the
mid 1980s. To understand the source of these elevated NO3 concentrations and the role
of atmospheric N deposition, Baron et al. (1994) used the CENTURY terrestrial
biogeochemistry model to estimate N uptake by plants and soils in alpine forest and
subalpine tundra within the Loch Vale watershed in ROMO. Based on the model output,
N uptake in the subalpine tundra was considerably lower than in the alpine forest.
Increased export of NO3 into stream water occurred at low N deposition rates (3 to
4 kg N/ha/yr) because much of this atmospheric N was deposited in the tundra or on
exposed rock surfaces, providing little N uptake. This work by Baron et al. (1994) was
the first N deposition CL research within the Northwestern Forested Mountains ecoregion
and aside from two CL estimates for lichens (Geiser et al.. 2010). the CL estimate
produced from this research is the lowest among the CL estimates reviewed by Pardo et
al. (2011c') for this ecoregion (Table 16-21).
Terrestrial biogeochemistry models have also played a role in the development of some
CL estimates for the ROMO region published since the Pardo et al. (2011c) assessment.
Sverdrup et al. (2012) and McDonnell et al. (2014a) used the ForSAFE-VEG model to
evaluate potential long-term effects of climate change and atmospheric N deposition on
alpine/subalpine ecosystems. Critical loads in both studies were defined as the rate of N
deposition at which plant diversity was protected from a change of 5 to 10 Mondrian
units (Mu; i.e., 5-10% change in plant species cover). Sverdrup et al. (2012) focused on a
"generalized" alpine/subalpine site, with vegetation composition constructed from plant
community data from ROMO and other national parks in the northern and central Rocky
Mountains region. Sverdrup et al. (2012) simulated plant responses to a future climate
(IPCC A2) and four levels of N and S deposition (preindustrial background, Clean Air
Act [CAA] controls, no CAA controls, and no CAA controls + high deposition). Soil
base saturation and soil solution base cation (Bc)-to-aluminum (Al) ratios (Bc:Al) were
predicted to decrease to less than 1% and less than 10% after Year 2100, respectively,
with the scenarios of no CAA emission controls and no CAA controls + high deposition
(indicating soil acidification). Future plant species coverages were predicted to change in
successively greater amounts in response to the altered climate, CAA emission controls,
no CAA emission controls, and no CAA emission controls + high deposition. Calculated
CLs (based on a change of 5 Mu) for N deposition were 1 kg N/ha/yr. Critical loads
related to S deposition were not discussed. All future N deposition scenarios (except
preindustrial background N) resulted in CL exceedance.
The approach of McDonnell et al. (2014a) was similar to that of Sverdrup et al. (2012)
because both applied the ForSAFE-VEG model to a subalpine ecosystem to determine
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CLs for long-term vegetation composition shifts caused by climate change and
atmospheric N deposition. The former study differed from the latter by using the Loch
Vale watershed within ROMO as the study area and also back-projecting the model to
1900. McDonnell et al. (2014a) estimated that the N deposition CL to avert a 10% (5 Mu)
change in future (2010-2100) plant biodiversity was 1.9 to 3.5 kg N/ha/yr, which had
already been exceeded. Future air temperature increases were predicted to further change
plant community composition and exacerbate changes caused by N deposition alone. In
the simulations between 1900 and 2010, the authors found that subalpine fir (Abies
lasiocarpa) sapling coverage had increased by more than 25%, graminoid response was
mixed, and forbs generally had decreased in abundance (Table 16-23). By 2010, Geum
rossii was reduced by more than 50% of its simulated historical cover. In scenarios using
ambient N deposition and 0.5 times ambient N deposition, pronounced increases in the
abundance of the moss Aulacomnium palustre would occur in 2065 and 2080,
respectively, while the dominant graminoid (Carex rupestris) decreased. They also found
that higher total N deposition scenarios would suppress moss cover near the end of the
simulation period. Projected future tree responses were similar between the lowest two
and highest three total N deposition scenarios. Future forb and Carex rupestris cover
remained relatively constant under higher total N deposition rate scenarios.
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Table 16-23 Hindcast absolute and percentage changes in species abundance
between 1900 and 2010 in response to historical reconstructions of
nitrogen deposition (Sullivan et al.. 2005) and historical climate
change (IPCC. 2007b).
Relative Abundance %
Growth Form
Species
1900
2010
Absolute
Change
% Change
Tree
Abies lasiocarpa
6.3
8.4
2.1
33.3
Shrubs
Salix Candida
15.4
25.5
10.1
65.6

Rubus parviflorus
0.0
4.4
4.4
—
Graminoids
Carex rupestris
6.8
16.4
9.6
141.2

Carex elynoides
25.1
12.8
-12.3
-49.0

Festuca ovina
1.7
3.2
1.5
88.2

Calamagrostis purpurascens
4.3
3.0
-1.4
-32.6

Poa abbreviata
3.2
2.3
-0.9
-28.1
Forbs
Geum rossii
19.7
8.9
-10.8
-54.8

Aquilegia caerulea
4.6
4.8
0.2
4.3

Antennaria rosea
1.5
2.0
0.5
33.3

Arenaria fendleri
4.7
1.8
-2.9
-61.7

Minuartia obtusiloba
3.7
1.6
-2.0
-54.1

Zigadenus elegans
0.0
2.2
2.2
—
Moss
Aulacomnium palustre
3.0
2.6
-0.5
-16.7
Source: McDonnell et al. (2014a).
16.5.3.2.	Aquatic Critical Loads
1	Nitrogen limitation of algal productivity is widespread in remote and undisturbed
2	freshwater lakes, such as the high-elevation lakes located throughout the Rocky
3	Mountains (Baron et al.. 201 lb). Based on surveys conducted in the 1980s that quantified
4	the ratio of dissolved inorganic N to total phosphorus, alpine lakes in the Rocky
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1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
Mountains region are predominantly N limited (45% of lakes) or colimited by both N and
P [22% of lakes; (Pardo et al.. 201 IcYI. This may underestimate the previous extent of N
limitation in these systems because lake sediment records indicate that N deposition
began impacting biogeochemistry in some alpine lakes during the 1950s (Das et al..
2005; Wolfe et al.. 2003; Wolfe et al.. 2001). Pardo et al. (2011c) summarized CLs for
ROMO surface waters (Table 16-24) with empirical and modeled CL estimates ranging
from 1.5 to 2.5 kg N/ha/yr. Within ROMO, eastern slope surface waters commonly have
higher dissolved inorganic N (DIN) than western slope surface waters (Wolfe et al..
2001; Baron et al.. 2000).
While the majority of ROMO aquatic research since 2000 focused on eutrophication,
Vertucci and Corn (1996) and Wolfe et al. (2003) described the potential for aquatic
acidification in the Colorado Rocky Mountains. Neither Vertucci and Corn (1996) nor
Wolfe et al. (2003) found evidence for acidification, citing the influence of catchment
sources of base cation neutralization. However, Wolfe et al. (2003) suggested that
persistent N deposition threatened to cause chronic surface water acidification,
particularly given that episodic declines of pH and acid neutralizing capacity (ANC) have
been well documented in east slope headwaters of the Front Range (Williams and
Tonnessen. 2000; Williams et al.. 1996b). Vertucci and Corn (1996) found that although
episodic acidification occurred during snowmelt, at no point was ANC < 0 in the northern
Colorado and southern Wyoming lakes and streams that were sampled. Importantly, the
observed acidification episodes did not coincide with the presence of amphibian embryos,
meaning that amphibians in the region were not threatened by acidification (including the
boreal toad, a candidate species for Endangered Species Act protection).
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Table 16-24 Critical loads for nitrogen for eutrophication for surface water
(high-elevation lakes) in the Rocky Mountains.
CL kg



Deposition/


N/ha/yr
Species
Response
Method
Addition
Site
Reference
1.5
Diatom
Eutrophication
Used exponential equations
Wet deposition
Southern
Bowman

community

correlating with NOx
2.94 kg N/ha/yr
Rockies/
et al.

composition

emissions from CO and
Dry deposition
Loch Vale
(2006)



11 western U.S. states to
0.94 kg N/ha/yr
ROMO




reconstruct historic N





deposition. 1950-1964 wet






N deposition correlated with






alteration of ROMO diatom






assemblages.



1.5
F. crotonesis
Eutrophication
Paleolimnological—analyzed
Not specified
Northern
Saros et

A. formosa

diatom fossil records of four

Rockies/
al. (2003)



lakes in the Beartooth

Beartooth




Mountains. Small Fragilaria

Mountains,




species declined while

WY




Fragilaria crotonensis,






Asterionella formosa, and






multiple Cyclotella spp.






increased



2.0
Not
Eutrophication
Used CENTURY model to
Modeled total
Southern
Baron et

applicable

discern if high lake and
deposition of 0.2
Rockies/
al. (1994)



stream N measurements are
to 16.0 kg
Loch Vale




attributed to alpine tundra
N/ha/yr
ROMO




and subalpine forest N






saturation



2.5
Chlorophyll a
Eutrophication,
Compared oligotrophic lake
Wet DIN
Rocky
Berqstrom


N and P
chemical and chlorophyll a
deposition 2.5 to
Mountains
and


colimitation
data in 42 European and
3.5 kg N/ha/yr for

Jansson



North American regions to
regions 26 to 29

(2006)



inorganic N deposition data
(Rocky






Mountains in






area of ROMO)


3.0
Not
Increases in
Compared observations of
Modeled total
Rocky
Baron et

applicable
lake nitrate
nitrate concentrations in
deposition of 1.5
Mountains
al. (2011b)


concentration
285 lakes to observations to
to 7.5 kg N/ha/yr





N deposition estimates






derived from NADP and






PRISM



CL = critical load; CO = carbon monoxide; DIN = dissolved inorganic nitrogen; ha = hectare; kg = kilogram; N = nitrogen;
NADP = National Atmospheric Deposition Program; N03" = nitrate; NOx = NO + N02; P = phosphorus; PRISM = a model for
spatial climate data; ROMO = Rocky Mountain National Park; yr = year.
Source: Baron et al. (2011 by Pardo et al. (20110). Williams and Tonnessen (2000).
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16.5.3.2.1. Empirical Studies
1	Pardo etal. (2011c) compiled a list of N deposition studies in high-elevation lakes that
2	have observed changes in the growth, biomass, or composition of phytoplankton
3	(Table 16-25). Within these studies, there was a relatively narrow range of lake water
4	nitrate concentrations at which these phytoplankton responses were observed
5	(Table 16-25). In paleolimnological studies of alpine lakes in this region that experienced
6	increased rates of deposition, there were pronounced shifts in the composition and
7	productivity of phytoplankton communities that occurred around 1950 (Table 16-26).
Table 16-25 Lake water nitrate concentrations in nitrogen deposition studies
observing phytoplankton responses.
N Cone,
(mg NOs'-N/L)
Species
Response
Deposition/Addition
Site
Reference
0.9
A. formosa, F.
crotonensis
Stimulation of
A. formosa, F.
crotonensis
Wet deposition
<2 kg N/ha/yr; N addition of
18 pmol N as NaNOs;
N + P enrichment, 18 pmol
N + 5 pmol P as NahtePCM
Beartooth
Mountains, WY
Saros et al.
(2005)
0.9
Chlorophyll a
Increasing
biomass
Wet DIN deposition 1.3 to
11 kg N/ha/yr; total N
deposition <0.1 kg N/ha/yr
to >15 kg N/ha/yr
Sweden
Berqstrom et al.
(2005)
0.9
Chlorophyll a
Increasing
biomass
Wet DIN deposition 2.5 to
3.5 kg N/ha/yr for regions
26 to 29 (Rocky Mountains
in an area of ROMO)
European and N.
American lakes,
including Rocky
Mountains
Berqstrom and
Jansson (2006)
1.0
Chlorophyll a
Shift in species
composition
Ambient deposition not
specified; single pulse N
treatment of 1,000 pg/L
NO3--N as KNOs
Snowy Range,
WY
Nvdick et al.
(2004)
1.1
Chlorophyll a
Diatom spp.
growth
stimulation
Ambient deposition of
3.5 kg N/ha/yr; N treatment
of ambient N03"-N plus
1 mg N03"-N/L
Colorado Front
Range (Loch
Vale)
Lafrancois et al.
(2003a)
Cone. = concentration; DIN = dissolved inorganic nitrogen; ha = hectare; kg = kilogram; KN03 = potassium nitrate; L = liter;
|jg = microgram; mg = milligram; N = nitrogen; NaH2P04 = monosodium phosphate; NaN03 = sodium nitrate; N03"-N = nitrate N;
P = phosphorus; ROMO = Rocky Mountain National Park; yr = year.
Source: Pardo et al. (20110).
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7
8
9
10
11
12
13
14
15
16
Table 16-26 Paleolimnological biological responses in Rocky Mountain lakes
exposed to anthropogenic nitrogen deposition.
Species
Response
Deposition/
Addition
Site
Year(s) of
Shift
Reference
A. formosa
Increased microbial activity,
greater primary production,
change in species
composition
Not specified
Colorado Rocky
Mountains
1950
Enders et al.
(2008)
A. formosa, F.
crotonensis
Greater primary production,
stimulation of A formosa
and F. crotonensis, shift in
species assemblages
Inorganic N
deposition
4 kg N/ha/yr
Eastside ROMO
lakes compared to
remote Colorado
Rocky Mountains
wilderness lake
1950
Das et al.
(2005)
A. formosa, F.
crotonensis
Greater primary production,
stimulation of A formosa
and F. crotonensis, shift in
species assemblages
Wet inorganic N
deposition 2 to
4 kg N/ha/yr
Dry deposition:
0.9 kg N/ha/yr
Colorado Rocky
Mountains
1950
Wolfe et al.
(2001). Wolfe
et al. (2003)
A. formosa, F.
crotonensis
Greater primary production,
stimulation of A formosa
and F. crotonensis, shift in
species assemblages after
1995
Not specified
Beartooth
Mountain Range,
WY (Yellowstone
NP region)
1995
Saros et al.
(2003)
ha = hectare; kg = kilogram; N = nitrogen; NP = national park; ROMO = Rocky Mountain National Park; yr = year.
Source: Pardo et al. (20110).
Since the Pardo et al. (2011c) CLs synthesis, Baron etal. (2011b) synthesized aquatic N
cycling research for lakes in several regions of the U.S., including alpine lakes in the
Rocky Mountains. Using existing water chemistry measurements and modeled estimates
of wet and dry N deposition, Baron etal. (2011b) observed that lake nitrate
concentrations increased where N deposition exceeded 2.0 kg N/ha/yr. This threshold
was similar to previous CLs of 1.5-4 kg N/ha/yr for lakes in this region (Table 16-24).
Baron etal. (2011b) also sought to develop a CL for N driven episodic aquatic
acidification, but found little data apart from that used to generate the (Williams and
Tonnessen. 2000) CL estimate of 4.0 kg N/ha/yr. Using phytoplankton biomass N to P
limitation shifts as the basis for CL calculations, Williams et al. (2017b) determined an
empirical CL of 4.1 kg/TN/ha/yr for remote high-elevation lakes across the western U.S.
The CLs were calculated as the total (wet + dry) N deposition rate at which point below
the CL, N limitation is more likely than P limitation and above the CL, P limitation is
more likely than N limitation. DIN:TP and DIN response categories yielded an average
critical load of 3.8 kg/TN/ha/yr for the lakes.
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4
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6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
Modeling Studies
In addition to the empirical estimates of aquatic CLs, Sullivan et al. (2005) and Hartman
et al. (2007) modeled CL estimates for aquatic acidification in ROMO. Sullivan et al.
(2005) used the MAGIC model to evaluate the sensitivity of two water bodies in the Loch
Vale watershed within ROMO—the Loch and Andrews Creek—to N and S deposition
based on ANC. Hartman et al. (2007) also modeled deposition scenarios for changes in
ANC at Andrews Creek, but used the DayCent-Chem model. Beginning with 1996
deposition rates of 2.2 kg S/ha/yr and 4.2 kg N/ha/yr, Sullivan et al. (2005) estimated CLs
for N or S in order to reach ANC of 0, 20, or 50 (ieq/L by 2046 in the Loch and Andrews
Creek, which had 1996 ANC values of 53 and 34, respectively. Acidification of the
watershed to ANC to 0 or 20 by 2046 in either water body would require doubling of
either N or S deposition (Sullivan et al.. 2005). However, increasing ANC to 50 in
Andrews Creek was not achievable by 2046 in any deposition scenario considered by
(Sullivan et al.. 2005). Hartman et al. (2007) took a somewhat different approach,
modeling 47-year scenarios beginning in 2000 where N deposition increased at rates that
were low (+60%), moderate (+120%), and high (+240%) relative to the contemporary
deposition rate. In these scenarios, episodic acidification (ANC = 0 (j,eq/L) of Andrews
Creek occurred at 6.3-7.1 kg N/ha/yr. Estimates for persistent (chronic) changes in ANC
of Andrews Creek were similar to those of Sullivan et al. (2005). within 1 kg N/ha/yr for
ANC = 20 |icq/L and within 3 kg N/ha/yr for ANC = 0 |icq/L (Table 16-27). Notably, the
Loch does not represent the most acid-sensitive lakes in the Front Range, with
approximately l/5th of the alpine lakes in the Colorado Front Range having a similar or
lower ANC (Eilers et al.. 1989). Williams et al. (2017b) modeled CLS for remote
high-elevation lakes across the Western U.S. based on nutrient limitation shift thresholds.
Modeled critical loads ranged from 2.8 to 5.2 kg/TN/ha/yr and correctly predicted
exceedances in 69% of lakes using NOs -N.
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Table 16-27 Critical loads of nitrogen or sulfur for surface water acidification
Rocky Mountain National Park and other high-elevation lakes in the
Rocky Mountains.
CL kg/ha/yr
of N or S
Species Response
Method
Deposition/
Addition
Site
Reference
4.0 (N)
Not
applicable
Episodic
freshwater
acidification
(ANC < 0 pmol/L)
1995 synoptic survey of NADP DIN wetfall Central
9 high-elevation lakes:
headwater catchments
experienced episodic
acidification due to
inorganic N in wetfall
(ANC < 0 pmol/L in
surface waters during
snowmelt).
at 23 sites.
>2,500 masl: 2.5
to 3.5 kg N/ha/yr
<2,500 masl:
<2.5 kg N/ha/yr
Rockies/
Colorado
Front
Range
Williams
and
Tonnessen
(2000)
5.8 (N)	Not	Freshwater
applicable acidification
(ANC = 50 peq/L)
MAGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
The Loch
(ROMO)
Sullivan et
al. (2005)
6.3 to 7.1 (N)
Not
Episodic
DayCent-Chem model
Deposition:
Andrews
Hartman et

applicable
freshwater
for 2000 to 2047 of four
2.7 kg S/ha/yr,
Creek
al. (2007)


acidification
N deposition (current,
3.5 kg N/ha/yr
(ROMO)



(ANC < 0 peq/L)
+60% increase, +120%






increase, +240%






increase)



7.1 (N)
Not
Freshwater
DayCent-Chem model
Deposition:
Andrews
Hartman et

applicable
acidification
for 2000 to 2047 of four
2.7 kg S/ha/yr,
Creek
al. (2007)


(ANC = 20 peq/L)
N deposition (current,
3.5 kg N/ha/yr
(ROMO)




+60% increase, +120%






increase, +240%






increase)



7.8 (N)
Not
Freshwater
MAGIC model scenario
Deposition:
Andrews
Sullivan et

applicable
acidification
for 1996 to 2046
2.2 kg S/ha/yr,
Creek
al. (2005)


(ANC = 20 peq/L)

4.2 kg N/ha/yr
(ROMO)

12.2 (N)
Not
Freshwater
MAGIC model scenario
Deposition:
Andrews
Sullivan et

applicable
acidification
for 1996 to 2046
2.2 kg S/ha/yr,
Creek
al. (2005)


(ANC = 0 peq/L)

4.2 kg N/ha/yr
(ROMO)

14.7 (N)
Not
Freshwater
MAGIC model scenario
Deposition:
The Loch
Sullivan et

applicable
acidification
for 1996 to 2046
2.2 kg S/ha/yr,
(ROMO)
al. (2005)


(ANC = 20 peq/L)

4.2 kg N/ha/yr


14.6 to 15.3
Not
Chronic
DayCent-Chem model
Deposition:
Andrews
Hartman et
(N)
applicable
freshwater
for 2000 to 2047 of four
2.7 kg S/ha/yr,
Creek
al. (2007)


acidification
N deposition (current,
3.5 kg N/ha/yr
(ROMO)



(ANC < 0 peq/L)
+60% increase, +120%






increase, +240%






increase)



20.6 (N) Not	Freshwater	MAGIC model scenario Deposition:	The Loch Sullivan et
applicable acidification for 1996 to 2046	2.2 kg S/ha/yr, (ROMO) al. (2005)
(ANC = 0 peq/L)	4.2 kg N/ha/yr
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3
4
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6
7
8
9
10
11
12
13
Table 16-27 (Continued): Critical loads of nitrogen or sulfur for surface water
acidification Rocky Mountain National Park and other
high-elevation lakes in the Rocky Mountains.
CL kg/ha/yr
of N or S
Species
Response
Method
Deposition/
Addition
Site
Reference
2.8 (S)
Not
applicable
Freshwater
acidification
(ANC = 50 peq/L)
AGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
The Loch
(ROMO)
Sullivan et
al. (2005)
4.6 (S)
Not
applicable
Freshwater
acidification
(ANC = 20 peq/L)
MAGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
Andrews
Creek
(ROMO)
Sullivan et
al. (2005)
7.8 (S)
Not
applicable
Freshwater
acidification
(ANC = 20 peq/L)
MAGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
The Loch
(ROMO)
Sullivan et
al. (2005)
8.1 (S)
Not
applicable
Freshwater
acidification
(ANC = 0 peq/L)
MAGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
Andrews
Creek
(ROMO)
Sullivan et
al. (2005)
11.1 (S)
Not
applicable
Freshwater
acidification
(ANC = 0 peq/L)
MAGIC model scenario
for 1996 to 2046
Deposition:
2.2 kg S/ha/yr,
4.2 kg N/ha/yr
The Loch
(ROMO)
Sullivan et
al. (2005)
ANC = acid neutralizing capacity; CL = critical load; DIN = dissolved inorganic nitrogen; ha = hectare; kg = kilogram; L = liter;
|jeq = microequivalent; |jmol = micromole; MAGIC = a biogeochemical process model; masl = meters above sea level;
N = nitrogen; NADP = National Atmospheric Deposition Program; ROMO = Rocky Mountain National Park; S = sulfur; yr = year.
16.5.3.3. Integration
Together, there is a considerable amount of research available on the effects of
atmospheric N deposition on ecological processes and characteristics within ROMO.
Assembled together, the CL estimates developed in and around ROMO (Figure 16-40)
show a continuum of effects as atmospheric N loads increase from natural background
rates (0.2 kg N/ha/yr) to levels several times greater than the current estimates of
3-9 kg N/ha/yr (Figure 16-35). Broadly, the chemistry and biota of aquatic systems
appear to be the most sensitive to N deposition, followed by alpine plant communities
(Figure 16-39). It is notable that many of the published CLs (Table 16-21. Table 16-22.
and Table 16-24) are near or below the rates of atmospheric N deposition currently
observed in ROMO. Thus, there is considerable evidence that current rates of
atmospheric N deposition are influencing the composition and function of aquatic and
terrestrial ecosystems within ROMO [Figure 16-41; (RMNP Initiative. 2014; Porter and
Johnson. 2007)1
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Rocky Mountain alpine lakes: shift in diatom community dominance.
. Baron etal., 2006 (ROMO) (1.5)
. Baron et a!., 2011 (285 RM lakes) (1.0)
Saros et al., 2005 (other RM) (1.5)
. Baron et al., So. (RM Loch Vale ROMO) (1.5)
Diatom assemblages already dominated by species indicative of moderate N enrichment, N limited oligotrophic
lakes switch to P-limitation after receiving only modest inputs of reactive N shifting the controls on diatom
species changes along the length of the nitrate gradient
Arnett et al., 2012 46 high elevation Rocky Mountains lakes (1-3.2 wet deposition)
Chlorophyll a concentration increase
• Elser et al., 2009a Colorado eastern lakes (2-7)
Eutrophication and P co-limitation lakes
• Bergstrom & Jansson, 2006 (RM lake) (2.5)
Diatom growth in A. formosa
Nanus et al., 2012 (location?) (>3)
Alpine vegetation
Bowman et al., 2011 (ROMO) (3.0)
Increased foliar chemistry, mineralization, nitrification, and nitrate leaching
Reuth & Baron, 2002; Reuth et al., 2003, ROMO (<4)
Alpine vegetation change (Corex rupestris indicator)
Bowman et al, 2006 Niwot Ridge (4.0)
Episodic acidification of surface waters
Baron et al., 2011 Niwot Ridge (4.0)
Williams and Tonnesson, 2000 Central Rockies, Colorado Front
Range
RR-N:RR-P ranging from 0.33 (P limitation) to 1.16 (N-P-co-
limitation) with the exception of 1 lake outlier of 1.42 (Green
Lake)
. Elser etal., 2009b (>6)
Nitrate leaching below rooting zone in alpine
Bowman et al., 2011, ROMO
I Community changes in alpine vegetation
| . Bowman et al, 2006 South Rockies Niwot Ridge (10.0)
Nitrate increases in soil solution in alpine
Bowman et al., 2011 ROMO (14.0)
O	5	10	15	20
Nitrogen deposition (kg/ha-yr)
ha = hectare; kg = kilogram; N = nitrogen; P = phosphorus; RM = Rocky Mountain; ROMO = Rocky Mountain National Park;
RR = relative response to changes in N or P; yr = year.
Figure 16-40 The continuum of ecological sensitivity to nitrogen deposition.
June 2018	16-134	DRAFT: Do Not Cite or Quote
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Chronic
acidification
Episodic
acidification
o
fi
!fc
LU
E
Ł
Vi
>s
W
o
o
LU
Change in alpine
plant species
Changes in soil
and foliar
chemistry
Change in
phytoplankton
composition and
abundance
Increased lako
and stream nitrate
J Potontial Futuro Effects
Current Effects

p
m
Ł3

N Load (kg-ha'yr1)
ha = hectare; kg = kilogram; N = nitrogen; yr = year.
Notes: current rates of N deposition impact terrestrial and aquatic ecosystems and relatively small increases in deposition may
cause further ecological change.
Source: Porter and Johnson (2007).
Figure 16-41
Critical load thresholds for current and possible future
biogeochemical and biological effects of nitrogen deposition.
16.5.4. Highlights of Additional Research Literature and Federal Reports since
January 2008
A number of studies documenting the effects of N or S deposition on terrestrial and
aquatic systems in ROMO or the ROMO region have been published since 2008. We
discuss these studies briefly here, but refer readers to the appropriate portion of the main
body of the ISA for a more detailed review of this work.
16.5.4.1. Terrestrial
A number of the studies reviewed in Appendix 4 and Appendix 6 of the current NOx-SOx
ISA as part of the body of terrestrial N deposition research published since 2008 were
conducted in and around ROMO. Farrer et al. (2013) and Bowman et al. (2012) both
conducted N addition studies in alpine tundra ecosystems to understand the response of
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plants to N deposition. Farrer et al. (2013) did not find an increase in net primary
productivity, but found both positive and negative effects on the aboveground biomass of
individual plant species. Bowman et al. (2012) did not find any significant overall effect
on aboveground plant biomass, plant tissue N concentration, or plant diversity; but the
abundance of an important sedge (Carex) species almost quadrupled. This contrasts with
Farrer et al. (2015) and Yuan et al. (2016). who found that N additions decreased plant
species diversity. In mixed grass meadows near ROMO, an experiment that recreated
historical, low N deposition conditions (soil N availability reduced by 63%, winter
precipitation reduced 50%) reduced biomass of invasive cool-season grass Bromus
tectorum (Concilio et al.. 2016).
There are no published CLs for changes in soil microbial communities in the ROMO
area, but several studies in alpine tundra ecosystems have quantified microbial responses
to N additions ranging from 8 to 288 kg N/ha/yr. Dean et al. (2014) observed decreased
ericoid fungi abundance and decreased richness and diversity of root-associated fungi,
while Farrer etal. (2015) and Boot et al. (2016) observed decreased microbial biomass,
fungal biomass, and bacterial biomass. Farrer et al. (2013). Nemergut et al. (2008). and
Yuan etal. (2016) documented shifts in microbial community composition.
Lieb et al. (2011) found that a decade of N additions to alpine ecosystems lowered soil
acid buffering capacity, decreased concentrations of base cation Mg2+, and increased
concentrations of Mn2+ and Al3+. After 15 years of N additions, Boot et al. (2016) found
that N addition reduced soil pH, soil percentage of C, and soil C:N in the O-horizon.
16.5.4.2. Aquatic
As with terrestrial ecosystems, there have been a number of new studies quantifying the
effects of N deposition on aquatic ecosystems in the Rocky Mountains since 2008
(Table 16-28). McCrackin and Elser (2012) measured denitrification in sediments of
alpine lakes in the Colorado Rocky Mountains that received elevated (5-8 kg N/ha/yr) or
low (<2 kg N/ha/yr) N deposition inputs. In high-deposition lakes, the NOs -N
concentration was significantly higher than in low-deposition lakes, but there was also
evidence that denitrifying bacteria could remove a meaningful portion of N inputs to the
lakes. While current levels of N deposition have not saturated the microbial capacity for
denitrification, the abundance of denitrifying bacteria has not increased in response to
additional N.
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Table 16-28 Summary of freshwater eutrophication studies in the Rocky
Mountains since 2008.
Species
Response
Method
Deposition/
Addition
Site
Reference
Not	Diatom assemblages
applicable already dominated by
species indicative of
moderate N enrichment
Diatom calibration
from surface
sediment samples
Wet deposition: 1 to
3.2 kg N/ha/yr
46
Arnett et al.
high-elevation (2012)
lakes in the
Rocky
Mountains
Not	Average lake DIN:TP of	Mesocosms	Ambient deposition Colorado Gardner et
applicable 16.3 + 2.76, suggesting	amended with	not specified	Front Range al. (2008)
widespread N saturation,	NO3", PO43", or	Addition'
but unclear if this is a new	PO43" + NO3"	930 mg/L NO3"
or historical condition
Chlorophyll a Greater N deposition
increased the N:P ratio in
lakes. Phytoplankton
growth P limited in high N
deposition lakes
(~7 kg N/ha/yr), but N
limited in low-N deposition
lakes (~2 kg N/ha/yr)
N and P enrichment
bioassay
experiments in
high- and
low-deposition
lakes
Deposition: 2 to
7 kg N/ha/yr
14 eastern
Colorado
alpine lakes
Elser et al.
(2009a)
Chlorophyll a
Chi a conc. increase.
Responses indicated a
range from P limitation to
N-P colimitation with one
N limited outlier lake.
Enrichment
bioassay
Wet deposition
>6 kg N/ha/yr
14 eastern
Colorado
alpine lakes
Elser et al.
(2009b)
A. formosa Maximum diatom growth Growth kinetics Total deposition Rocky	Nanus et al.
rate at 0.5 |jM of NO3" experiments	>3 kg N/ha/yr	Mountains (2012)
Chi a = chlorophyll a; DIN = dissolved inorganic nitrogen; ha = hectare; kg = kilogram; L = liter; |jM = micromolar; mg = milligram;
N = nitrogen; N03" = nitrate; P = phosphorus; P043" = phosphate; TP = total phosphorus; yr = year.
Mastetal. (2011) examined trends in precipitation chemistry and other factors that
influence long-term changes in chemistry of high-elevation lakes in three Colorado
wilderness areas. Mastetal. (2011) observed that sulfate concentrations in precipitation
decreased at rates of-0.15 to -0.55 (j,eq/L/year during the Years 1985 through 2008 at
10 monitoring stations in Colorado. In high-elevation lakes where SO42 was primarily
derived from atmospheric sources, the lake SO42 concentrations decreased by -0.12 to
-0.27 (ieq/L/year. In lakes where watershed sources were the dominant source of SO42 .
the S042 concentrations in lake water increased during this time period. Mast et al.
(2011) attributed this trend to the effects of climate warming on pyrite weathering.
Similarly, Heath and Baron (2014) observed increases in summer (June-August) fluxes
and concentrations of cations, SiC>2, SO42 . inorganic N, and ANC in stream water from
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1984-2010 within the Loch Vale watershed in ROMO and attributed these trends to
climate change effects on nitrification, mineral weathering, and the release of elements
entrained in snow and ice. In comparison, wet N deposition rates were relatively constant
and wet SOr deposition decreased.
Mast et al. (2014) measured long-term changes in stream NOs concentration over three
decades at the Loch Vale watershed in RMNP. The concentrations of NO3 in stream
water increased during the early 1990s, peaked in the mid 2000s, and then declined by
more than 40%. Nanus et al. (2012) developed a surface water NO3 threshold for growth
of the diatom A. formosa for high-elevation lakes in the Rocky Mountains. Arnett et al.
(2012) tried to develop a CL for diatom community composition change along a gradient
of 46 high-elevation lakes receiving 1 to 3.2 kg N/ha/yr in wet deposition, but observed
that diatom communities were dominated by a species associated with moderate N
enrichment even in lakes where NO3 concentrations were at the detection limit
(<1 |ig/L).
New information on relative NO3 inputs from glacial versus snowpack meltwaters
indicate water of glacial origin has higher NO3 , which may influence interpretation of
biological data from high-altitude lakes and streams in some regions of the U.S.,
including the Rocky Mountains (Slemmons et al.. 2015; Slemmons et al.. 2013; Saros et
al.. 2010; Baron et al.. 2009). In the central Rockies and Glacier National Park, fossil
diatom richness in snowpack-fed lakes was found to be higher relative to lakes fed by
both glacial and snowpack meltwaters (Saros et al.. 2010). Sedimentary diatom analysis
of alpine lakes in the Rocky Mountains indicated that increases in A. formosa occurred at
least a century ago in glacially fed lakes whereas shifts in the planktonic diatom
communities occurred after 1970 in snow-fed lakes (Slemmons et al.. 2017).
16.5.5. Rocky Mountain National Park Initiative
The large body of research on the effects of atmospheric deposition on terrestrial and
aquatic ecosystems in ROMO and the relatively low CLs has led the National Park
Service (NPS) to partner with the Colorado Department of Public Health and
Environment and the U.S. EPA to develop a plan in 2005 to decrease atmospheric N
pollution within ROMO (Porter and Johnson. 2007). Formally, this collaboration is
organized as the Rocky Mountain National Park Initiative (RMNPI). The goal of the
collaboration is to understand the impacts of N deposition in ROMO and the sources of N
emissions, and resolve the N deposition issue in ROMO (RMNP Initiative. 2014). The
three agencies developed the Nitrogen Deposition Reduction Plan, which intended to
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decrease N wet deposition in ROMO from a 2006 baseline of 3.1 kg N/ha/yr to a 2032
target of 1.5 kg N/ha/yr.
The Nitrogen Deposition Reduction Plan aims for a linear decrease in deposition between
2006 and 2032, with interim milestones of 0.3 kg N/ha reductions in annual deposition
every 5 years. In 2012, the 5-year rolling average of wet deposition was 2.9 kg N/ha/yr, a
number that was 0.2 kg N/ha/yr above the interim milestone target, but also reflective of
a 4-year trend toward decreased wet N deposition (RMNP Initiative. 2014). However, in
the most recent status report (Morris. 2016). the 5-year rolling average has increased in
each of the last 2 years and is now 3.3 kg N/ha/yr (Figure 16-42). well above the intended
target. There has been no overall trend in wet N deposition or nitrate deposition since
2007, but the ammonium concentration has increased at four of the five monitoring sites
since 2009. Within the nine-county Denver Metropolitan and Northern Colorado Front
Range region, the two largest estimated sources of ammonia emissions were livestock
(75% of emissions) and fertilizer use [14% (RMNP Initiative. 2015)1. However,
considerable uncertainty remains in ammonia emissions and transport estimates, and the
members of RMNPI have dedicated resources to research ammonia in the region. The
agency partners within the RMNPI are working with the agricultural industry in Colorado
to decrease ammonia emissions by adopting a set of best management practices for
agricultural N (Figure 16-43).
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3.30 kg N/ha/yr*
-2032 Gfidepath
•5-yr Rolling Average Wet N Deposition
2nd Interim Milestone
(2.4 Kg N/ha/yr)
3rd Interim Milestone
{2,1 kg N/ha/yr)
1st Interim Milestone
(2.7 kg N/ha/yr)
4th Interim Milestone
(1.8 kg N/ha/yr)
Resource Management Goal
(1.5 kg N/ha/yr)
Natural Conditions
(0.2 kg N/ha/yr)

8
o
(M
Year
551
a
ha = hectare; kg = kilogram; N = nitrogen; yr = year.
Source: (Morris et al.. 2014).
Figure 16-42 Rocky Mountain National Park Initiative glidepath and current wet
nitrogen deposition at Loch Vale in Rocky Mountain National
Park.
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RMXP INITIATIVE ACCOMPLISHMENT TIMELINE
•	Established resource
management goal (critical
load)
•	RoMANS field campaign
Evaluate Loch Vale
monitoring data
Critical load interim goal
to be reviewed with
respect to XDRP and CP
Ag = agriculture; APCD = Air Pollution Control District; AQ = air quality; AQCC = Air Quality Control Commission; BMP = best
management practice; CDPHE = Colorado Department of Public Health and Environment; CLA = Colorado Livestock Association;
CP = Contingency Plan; EPA = U.S. Environmental Protection Agency; NADP = National Atmospheric Deposition Program;
NPS = National Park Service; NDRP = Nitrogen Deposition Reduction Plan; RMNP = Rocky Mountain National Park;
RoMANS = Rocky Mountain Atmospheric Nitrogen and Sulfur.
Source: CDPHE (2012). https://www.colorado.gov/pacific/sites/default/files/AP PO ROMO-lnitiative-Accomplishment-Timeline.pdf.
Figure 16-43 Rocky Mountain National Park Initiative accomplishment timeline.
16.5.6. Interactions between Nitrogen Deposition, Precipitation, and
Large-Scale Ecological Disturbances
Precipitation patterns have changed over the past decades in ROMO, with consequences
for biodiversity responses to N deposition. Spring snowfall has diminished over the last
two decades (Clow et al.. 2016). while winter snowfall has increased over the last century
(Concilio et al.. 2016). As a consequence, snowmelt occurs 7-18 days earlier than in
1993 (Clow et al.. 2016). which extends the growing season. In a review of Niwot Ridge
research. Bowman et al. (2014) suggests that N additions tend to have the opposite effect
that high snowpack has on soil properties and processes (i.e., if N deposition accelerates
soil N cycling, higher snowpack slows soil N cycling). Long-term trends in decreasing
snowpack caused by changes in temperature and dust deposition (Clow et al.. 2016). will
intensify the effects of N deposition upon biodiversity.
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Most forests within ROMO are composed of lodgepole pine or a mixture of Engelmann
spruce and subalpine fir. Historically, these forests have been vulnerable to two natural
mortality agents: bark beetles and wildfire (Ehlc and Baker. 2003). Bark beetles can kill
more than 90% of mature trees in heavily infested areas (Rhoades et al.. 2013). while
wildfires in these types of forests tend to burn at high intensities and kill trees across
large landscapes (Sibold et al.. 2006; Buechling and Baker. 2004; Veblen et al.. 1994;
Romme and Knight. 1981). The large number of trees killed in these disturbances
radically alters many ecological processes, including N cycling.
After tree mortality events, the decrease in biotic demand for N results in higher levels of
inorganic N in the soil and increased concentrations of N in the leaves of surviving or
recolonizing plants (Diikstra and Adams. 2015; Mikkelson et al.. 2013; Rhoades et al..
2013). These changes in N cycling are similar to those caused by N deposition (Fcnn et
al.. 1998) and those in experiments in and near ROMO simulating N deposition (Rueth et
al.. 2003). Although these disturbances increase N availability, beetle epidemics or
wildfires in Rocky Mountain lodgepole pine and spruce-fir forests have not historically
caused large increases in soil nitrate leaching or high concentrations of N in stream water
(Dunncttc et al.. 2014; Morris et al.. 2013; Rhoades et al.. 2013). processes associated
with excess N availability and the state of "N saturation" that occurs in ecosystems
receiving large amounts of N deposition (Fcnn et al.. 2003a). However, there is concern
that N deposition along the Colorado Front Range will enhance N availability to the point
where disturbances will lead to large N leaching losses (Rhoades et al.. 2013). This
means that there is potential for N deposition to interact with bark beetles and wildfire to
increase stream and lake water N content and degrade ecosystem function.
Nitrogen deposition can increase insect attack on plants (Throop and Lerdau. 2004) and
there is some evidence that bark beetle activity and the bark beetle-caused tree mortality
can be increased by N deposition (Jones et al.. 2004). Unlike in other parts of the western
U.S. where N deposition has allowed fires to occur more frequently by increasing fuels
through greater plant growth (Fenn et al.. 2003a). there is unlikely to be a large effect of
N deposition on fire occurrence in ROMO. Fires in ROMO occur mostly during extreme
weather conditions and are not as directly influenced by the amount of fuel available
(Sibold et al.. 2006). Notably, warm climate conditions increase fire frequency in
lodgepole pine and spruce-fir forests in the Rocky Mountains (Westerling et al.. 2006)
and support bark beetle outbreaks in these forests (Bentz et al.. 2010; Hebertson and
Jenkins. 2008).
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16.6.
Southern California
16.6.1. Background
This case study focuses on the impact of N and S deposition on aquatic and terrestrial
ecosystems in southern California. Two areas are considered that have abundant
information on ecological effects: the Sierra Nevada and southern California arid land
ecosystems. Here we focus on the national parks located in both of these ecosystems, the
Sequoia and Kings Canyon national parks (SEKI), located in the southern Sierra Nevada
Mountains in southern California, and Joshua Tree National Park (JOTR), located in the
desert to the southeast (Figure 16-44). SEKI and JOTR were selected to emphasize
Class I areas in the Sierra Nevada (SEKI) and southern California arid land ecosystems
(JOTR). In addition, research from surrounding ecoregions relevant to these ecosystems
is included because more work has been conducted within the surrounding areas than
within the park boundaries.
The majority of the information provided in this case study represents the Sierra Nevada
Mountains/SEKI. Limited research conducted in JOTR was found in the peer-reviewed
literature. Yosemite National Park (YOSE) to the north shares many commonalities with
SEKI regarding air pollution exposure, sensitivities, and impacts. Thus, research
conducted in YOSE is relevant and included here. Background information is provided
below, including a general description of the region, protected status designation
(e.g., Class I, Wilderness Area), and regional land use. This case study includes
atmospheric deposition to the parks and surrounding areas, dose-response relationships,
and critical loads.
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ALPINE
CHAPARRAL
COASTAL SAOF. SCRUB
DESERT
¦ GRASSLAND
MIXED CONIFER
¦	MIXED liARDWOOD
¦	OAK WOODLAND
RIPARIAN
SAGEBRUSH STEPPE
m WATER
WETLAND
Mta
¦hi ii' uriLti i«n
Cm tm ('inri<— hniim
Ut -	W
Notes: Land cover presented in this figure represents potential natural vegetation before urbanization and modern agricultural
development.
Source: Reclamation et al. (1996V from Fenn et al. (2010V Adapted to show case study locations.
Figure 16-44 Map of the distribution of vegetation types and land cover in
California.
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16.6.1.1.
Description of Case Study Region
Sequoia and Kings Canyon national parks are located in the Mediterranean California
ecoregion (Level I Omernick classification) and are part of the Sierra Nevada Mountains.
These parks are known for exceptionally large trees, high mountain peaks and ridges, and
deep canyons and include the highest point in the contiguous U.S. (CONUS), Mount
Whitney at 14,494 ft (4,418 m) above sea level. There are more than 200 marble caverns,
many with endemic cave invertebrates. Sequoia National Park was originally created in
1890 to protect the giant sequoia trees (Sequoiadendron giganteum) from logging. It was
the second national park and the first formed to protect a particular living organism. In
1940, Congress created the contiguous Kings Canyon National Park to include the
glacially formed Kings Canyon. These parks have increased in size to encompass
1,353 mi2 (3,504 km2); 97% is designated and managed as wilderness.
Ecosystems in SEKI include chaparral vegetation at low elevation, extending up to alpine
areas in the high mountains. Mycorrhizae, lichens, and herbaceous species communities
are susceptible to declines in species richness/biodiversity due to the effects of N
deposition. Likewise, high-elevation lakes and streams as well as low-order streams are
susceptible to eutrophication and acidification driven by N deposition. Five rivers begin
in SEKI, and past and present glaciers contributed to more than 3,000 lakes (many at high
elevation) and 2,000 miles of streams, generally of low Strahler order (Boiano ct al..
2005). Landscape variation contributes to diverse habitats that contain various
assemblages of terrestrial, aquatic, and subterranean biota. As of 2005, of the more than
1,500 taxa of vascular plants in SEKI, 138 were deemed special status; 24 were listed by
the California Native Plant Society as rare, threatened, or endangered. Fifty-one taxa of
the parks' 312 terrestrial and aquatic vertebrate species had special status designations.
Five extant native vertebrate species were federally listed in 2005: Sierra Nevada bighorn
sheep (Ovis canadensis sierrae), bald eagle (Haliaeetus leucocephalus), Little Kern
golden trout (Oncorhynchus mykiss whitei), mountain yellow-legged frog (Rana
muscosa), and Yosemite toad (Anaxyrus canorus). As of 2005, the state listed the
peregrine falcon (Falco peregrinus), great gray owl (Strix nebulosa), and willow
flycatcher (Empidonax traillii) as endangered and the red fox (Vulpes vulpes), wolverine
(Gulo gulo), and Swainson's hawk (Buteo swainsoni) as threatened. Forty extant native
vertebrate species were listed as sensitive, including: white-tailed jackrabbit (Lepus
townsendii), fisher (Martespennanti), ring-tailed cat (Bassariscus astutus), Townsend's
big-eared bat (Corynorhinus townsendii), western mastiff bats (Eumops perotis), merlin
(Falco columbarius), northern harrier (Circus cyaneus), California spotted owl (Strix
occidentalis), California legless lizard (Anniellapulchra), western pond turtle (Actinemys
marmorata), Mount Lyell salamander (Hydromantes platycephalus), San Joaquin roach
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fish (Lavinia symmetricus), and Kern River rainbow trout [Oncorhynchus mykiss gilberti;
(Boiano et al.. 2005)1.
Joshua Tree National Park is in the North American Desert ecoregion (Level I ecoregion
by Omernick classification). The park was initially dedicated as a U.S. National
Monument in 1936 and later declared a national park in 1994 when the U.S. Congress
passed the California Desert Protection Act (Public Law 103-433). The park covers
1,235 mi2 (3,200 km2), more than half designated as wilderness. It straddles the San
Bernardino County/Riverside County border, including parts of two deserts: the higher,
wetter Mojave Desert and the lower, drier Colorado Desert. The Little San Bernardino
Mountains cross the southwestern edge of the park. JOTR is covered primarily by
semiarid and arid vegetation types, including Joshua tree (Yucca brevifolia) woodlands.
While rainfall is less than 10 cm/year, JOTR's groundwater supplies 200 surface water
sources (e.g., springs, oases, ephemeral streams, or washes).
Joshua Tree National Park was preserved to protect the natural resources found at the
junction of three California ecosystems. The Colorado Desert is a western extension of
the Sonoran Desert. It occupies the southern and eastern portions of the park, and is
characterized by ocotillo (Fouquieria splendens) shrubs and cholla (Opuntia spp.) cacti.
The Mojave Desert crosses the northern portions of the park and contains Joshua tree
woodlands. The third ecosystem type in JOTR is a community of California juniper
(Juniperus californica) and pinyon pine (Pinus spp.) located in the westernmost portion
of the park, above 1,220 m elevation, in the Little San Bernardino Mountains.
There are several threatened and endangered species found within JOTR. Plant species
include the triple-ribbed milkvetch (Astragalus tricarinatus) and Coachella Valley
milkvetch (A. lentiginosus var. coachellae), which are federally endangered; as well as
the Parish's daisy (Erigeronparishii), which is federally threatened. The desert tortoise
(Gopherus agassizii), California's state reptile, is on both California's and the federal
threatened lists. Twenty-six other species are listed as "special concern." JOTR was
originally inhabited by the Serrano, the Chemehuevi (sometimes called the Southern
Paiutes), and the Cahuilla people, and the park contains over 500 archaeological sites.
(https://www.nps.gOv/i otr/index.htm').
16.6.1.2. Class I Areas
The 1964 Wilderness Act and the NPS Organic Act are both used as tools to protect
SEKI and JOTR from air pollution impacts. SEKI and JOTR (as well as nearby YOSE)
are also Prevention of Significant Deterioration (PSD) Class I areas. They receive the
highest level of protection under the Clean Air Act (CAA) against air pollution
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degradation. The CAA (42 USC 7470) authorized designation of Class I areas to protect
air quality in national parks (over 6,000 ac [2,428 hectares]) and national wilderness
areas (over 5,000 ac [2,023 hectares]) in an effort to preserve pristine atmospheric
conditions and air quality-related values (AQRVs). Class I areas are subject to the PSD
regulations under the CAA. PSD preconstruction permits are required for new and
modified existing air pollution sources. Air regulatory agencies are required to notify
federal land managers (FLMs) of any PSD permit applications for facilities within
100 km of a Class I area.27 The FLMs are authorized to review and comment on PSD
Class I permit applications with the permitting agency.
16.6.1.3. Regional Land Use and Land Cover
Land cover in SEKI is largely forested and mountainous terrain in and near the Sierra
Nevada Mountains. The westernmost portions of the park are covered by mixed conifer
and chaparral vegetation. Further to the west is a zone of intensive agriculture
(Table 16-29).
Table 16-29 Land coverages of Sequoia, Kings Canyons, and Joshua Tree
national parks.
Land Cover Category
Sequoia National Park
(km2)
Joshua Tree National Park
(km2)
Kings Canyon
National Park
(km2)
Developed
5
7
4
Barren land and exposed
rock
416
207
690
Deciduous forest
12
0
2
Evergreen forest
736
5
519
Mixed forest
8
0
1
Shrub/scrub
390
2,936
516
Grassland/herbaceous
68
54
100
Wetland
2
0
2
km = kilometer.
27 http://webcam.srs.fs.fed.us/psd.
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Land cover in JOTR is largely desert and semiarid land (Table 16-29. Figure 16-44).
Nearby human population centers (Figure 16-45) include the Coachella Valley
(-350,000 people, including the cities of Palm Springs and Indio), as well as the towns of
Twentynine Palms (25,600 people) and Joshua Tree (7,400).
Hi
V
\
w •
r
\
i
I
r \
¦ L "I.	r)
J Vosennita

.Sierra National
Forest
Sequoia
National Park
Kind's Canyon
National Park
NEVADA
CALIFORNIA
Los Padres
National Forest
"55s

Santa Monica
Mountains
Southern California
Case Study Region
Case Study Areas
Native American
Reservations
USA City Populations
•	1,000,000 plus
•	500,000 - 999,999
250,000 - 499,999
•	100,000 - 249,999
20 40 60
CP
)	y\
Mojave National 4^;
Preserve
v\
Figure 16-45 Southern California case study region showing locations of
human population centers.
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16.6.2. Deposition
Characteristics of N and S deposition are shown in Figure 16-46 through Figure 16-48 for
JOTR and Figure 16-49 through Figure 16-51 for SEKI. Data shown in the figures were
obtained from the hybrid modeling/data fusion product, TDEP
(http://nadp.sws .uiuc.edu/committees/tdcp/tdcpmaps/). and described earlier in
Appendix 2. However, the time series of wet deposition is taken directly from data on the
NADP/NTN website. This was done to track changes in deposition since the passage of
the Clean Air Act Amendment because the CMAQ simulations involved in TDEP extend
back only to 2000. See Appendix 2.4. Appendix 25_ and Appendix 2JS for more
information on deposition in the U.S. Other maps showing the contributions of individual
species to dry and/or wet deposition are given in Appendix 2.7.
Figure 16-46 indicates that the general pattern of deposition in JOTR of N and S is
broadly similar, tending to be higher near the NADP site located in the NW corner of the
study area and decreasing inland. As expected, N deposition is highest in and surrounding
Los Angeles. The park is often subject to transport of emissions from the Los Angeles
Basin. However, as can be seen from Figure 16-46B. this influence decreases
substantially towards the southeast. In addition, the area surrounding JOTR shows a high
degree of regional heterogeneity. Deposition of S is consistent with the rest of the
western U.S., but is considerably lower than in the eastern U.S.
As shown in Figure 16-47A. deposition of N is estimated to be mostly in oxidized forms
in JOTR. Although there are areas principally to the south or to the west where N is
deposited mostly in reduced forms. In Figure 16-48. wet deposition of NO3 , NH4 .
SO42 . and H+, apart from being lower than at many other sites, has not shown consistent
trends over the past 25 years. Comparison of Figure 16-46 and Figure 16-47 shows that
dry deposition dominates over wet deposition of N and S in the JOTR study area.
In SEKI, Figure 16-49 shows a 3-year average total deposition of N and S for
2011-2013; Figure 16-50 shows the partitioning between oxidized and total N;
Figure 16-51 shows the 25-year time series for wet deposition of NO3 , NH4 . SO42 . and
H+ obtained at the NADP/NTN monitoring sites near SEKI.
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~ LO* Ar»gal*v CA
O Mon'tof CA ( 6?
) Monitor location*
Jo*ho» Tr« Study *'«
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•> S*» V* «>
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Lo* Ang#l#v CA

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# Monitor Location*

Jothu* Tree Study *¦'»'




¦ «¦

0
N OqKMDWn (Ifl H'M)
'VV^VVV*
* * % \v so
CONUS = contiguous U.S.; N = nitrogen; S = sulfur; ha = hectare; kg = kilogram.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 2.7.
Figure 16-46
Patterns and temporal trends of nitrogen and sulfur deposition in Joshua Tree National Park and
surrounding region in California. A and B show the 3-year average total deposition of nitrogen
and sulfur for 2011 -2013
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CONUS = contiguous U.S.; N = Nitrogen, S = Sulfur.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 2,7,
Figure 16-47 Patterns and temporal trends of nitrogen and sulfur deposition in Joshua Tree National Park and
surrounding region in California. A shows the partitioning between oxidized and reduced
nitrogen; B and C show the 3-year average total percentage of wet deposition of nitrogen and
sulfur for 2011-2013.
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100

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Annual Wet Deposition and 3-Year Moving Average at
Site CA67: 1990-2014
•	NIV
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2001
2003
2005
2007
Year
2009
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. _ _
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2013
2015
H+ = hydrogen ion; ha = hectare; mol = mole; NH4+ = ammonium; N03 = nitrate; S042 = sulfate; yr = year.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 2.7.
Figure 16-48
Patterns and temporal trends of nitrogen and sulfur deposition in Joshua Tree National Park and
surrounding region in California. The 25-year time series for wet deposition of nitrate,
ammonium, sulfate, and hydrogen ion obtained from the National Atmospheric Deposition
Program/National Trends Network monitoring site CA 67.
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CONUS = contiguous U.S.; H+ = hydrogen ion; ha = hectare; mol = moie; NH/ = ammonium; NO32" = nitrate; S042 = sulfate; yr = year.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 11.
Figure 16-49 Patterns and temporal trends of nitrogen and sulfur deposition of Sequoia and Kings Canyons
national parks and surrounding region in California. A and B show the 3-year average total
deposition of nitrogen and sulfur for 2011 -2013.
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CONUS = contiguous U.S.; H+ = hydrogen ion; ha = hectare; mol = mole; NH4+ = ammonium; N032" = nitrate; S042 = sulfate; yr = year.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 2,7,
Figure 16-50 Patterns and temporal trends of nitrogen and sulfur deposition of Sequoia and Kings Canyons
national parks and surrounding region in California. A shows the partitioning between oxidized
and reduced nitrogen, indicated as the fraction of total nitrogen which is oxidized; B and C show
the 3-year average total percentage of wet deposition of nitrogen and sulfur for 2011-2013.
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Annual Wet Deposition and 3-Year Moving Average at A
ivo
too
no
2
^ MO
8 ivo
1*00
a v)
Site CA75: 1990 - 2014
•*»C
•	NO,
•	«>,*
• H* i Ji
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• %
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1 • ir, • tv^'v
iw im /ooo M4 ma Mil
Year
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Annual Wtt Deposition and 3-Year Moving Average at B
TOO
1M
>. 1M
\ 140
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•5 10
S to
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JO
0
1H
Sit# CA99: 1990 2014
•	HO.
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#	• *Y / # • k	\#/*
V..... • . ,\vw.
! •./ . . V..	*v	 '•

• • . • v«" . • *u
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MOO	M04
Yetr

MU
CONUS = contiguous U.S.; H+ = hydrogen ion; ha = hectare; moi = moie; NH4+ = ammonium; N032" = nitrate; S042 = sulfate; yr = year.
Surrounding areas in California and Nevada and inserts showing the whole CONUS are shown to place the depositional environment for both portions of the study area in context.
Other maps showing the contributions of individual species to dry and/or wet deposition are given in Appendix 2.7.
Figure 16-51 Patterns and temporal trends of nitrogen and sulfur deposition of Sequoia and Kings Canyons
national parks and surrounding region in California. A and B show the 25-year time series for wet
deposition of nitrate, ammonium, sulfate, and hydrogen obtained from the National Atmospheric
Deposition Program/National Trends Network monitoring sites CA99 and CA75.
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Deposition of N is much higher in SEKI than JOTR. There is also much greater spatial
variability in N deposition in SEKI than JOTR. On the other hand, S deposition is similar
in magnitude between the two sites; S is also characterized by relatively low spatial
variability at the two sites. In contrast to JOTR, deposition of N is estimated to be mostly
in reduced forms in SEKI (see Figure 16-50). This is not surprising, given the proximity
of SEKI to the San Joaquin Valley. There is considerable variability in the percentage of
N deposition as either in oxidized or reduced forms. Wet deposition of NO3 , NH4+,
SOr . and H+ is larger over SEKI than over JOTR, but still lower than at some NADP
sites in the eastern U.S. Similar to JOTR, dry deposition dominates over wet deposition
of both N and S in SEKI.
A recent modeling study of Class I areas, including SEKI and JOTR, used a GEOS-Chem
adjoint model to identify the geographic sources of reactive N deposition, as well as the
emission sector sources of reactive N deposition within the parks (Lee et al.. 2016). 90%
of emissions of Nr that affect SEKI originate within 400 km of the park, with livestock
ammonia as the major emission source (>50%) of N deposition within SEKI (see
Figure 16-52). At JOTR, 90% of N emissions originate within 600 km of the park, and
mobile sources of NOx are the major emission source (>60%) of N deposition to JOTR
(see Figure 16-52.
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Geographic footprint of Nr deposition
0.005 0.01 0.02 0.03 0.04 0.06 0.08 0.1
[kg N/ha/yr]
Emission sectors
NOx(surface inventory)
| NOx(electric generating units)
NOx (non-electric generating
industrial stacks)
J NOx (aircraft)
NH3(livestock)
H NH3 (fertilizer)
^ NH3(natural)
NOx(lightning)
^ NOx (soil)
Figure 18-52 Annual-averaged monthly footprint of reactive N deposition in
Joshua Tree National Park (3.2 kg N/ha/yr) and Sequoia National
Park (5.7 kg N/ha/yr). Also shown for each park is a pie chart of
fractional contribution from emission sectors, as estimated by
GEOS-Chem adjoint model.
16.6.3. Critical Loads and Other Dose-Response Relationships
16.6.3.1. Terrestrial
1	The communities of terrestrial species with well-documented effects caused by N
2	deposition include lichens, mycorrhizae, grasses, and trees. While well studied
3	ecosystems include mixed conifer forest, alpine/subalpine, and chaparral/semiarid
4	shrubland.
5	Data on the effects of N deposition on mixed conifer forest ecosystems in the
6	Mediterranean California ecoregion (Omernick Level I), including the Sierra Nevada
7	region, are well summarized by Fenn et al. (201 lb). A more recent overview of empirical
8	and modeled CLs for mixed conifer forests was provided by Fenn et al. (2015). Nitrate
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leaching from soil in combination with high evapotranspiration rates lead to high nitrate
concentrations in surface waters in this area.
In general, the alpine and subalpine plant communities, such as those found in SEKI and
YOSE, are sensitive to eutrophication, an observation based largely on research
conducted in alpine ecosystems in the Rocky Mountains (Bowman et al.. 2006). Alpine
and subalpine communities are high-elevation plant communities that have developed
under conditions of low nutrient supply, in part because soil-forming processes tend to be
poorly developed at high elevations. This contributes to N sensitivity. Consequently,
changes in alpine plant productivity and species composition have been reported in
response to increased N inputs (Bowman et al.. 2006; Vitousek et al.. 1997).
In chaparral and other semiarid shrublands in southern California, nitrogen enrichment
can contribute to an increase in net primary productivity; however, water availability is
often limiting to growth (Vourlitis. 2012). Arid ecosystems represent an extreme in terms
of water availability and water is the principal limiting resource. Therefore, long-term
(many years) experiments and the potential for interactions among water, fire, and
nutrient supply in governing plant responses to N additions should be considered. Based
on N addition experiments, Vourlitis and Pasquini (2009) suggested that dry-season
addition of N significantly changes the community composition of coastal sage scrub
(CSS) vegetation, but not chaparral. The impacts of such changes on the plant community
may magnify over time. It has been proposed that increases in both fire frequency and N
deposition over the last several decades in southern California may have promoted the
conversion of CSS land to non-native grasslands (Talluto and Suding. 2008). These same
effects may also occur in other arid ecosystems including JOTR. Both empirical and
modeling studies suggest that loss of biodiversity in response to N input and climate
change can be greater than the effects of either stressor alone (Porter et al.. 2013).
Empirical Studies
A number of empirical studies (N gradient and N experimental additions) have been
conducted to identify tipping points and to define areas of CL exceedance. Key studies
conducted in the southern California case study region are listed in Table 16-30.
Fenn et al. (2010) reported empirical CL exceedance maps for seven major vegetation
community types in California (Figure 16-53). Thirty-five percent of the land area
covered by these vegetation types (100,000 km2) was estimated to be in exceedance of
the nutrient-N CLs (3-8 kg N/ha/yr) that were estimated for mixed conifer forests,
chaparral, and oak woodlands. Nearly half of the area covered by CSS (54%) and
grasslands (44%) was judged by the authors to be in exceedance of the CLs for protecting
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the ecosystem against changes associated with invasive grasses. Substantial chaparral
(53%) and oak woodland (41%) cover was estimated to be in exceedance for impacts on
epiphytic lichens. About 30% of the desert area investigated was in exceedance, based on
the presence of invasive grasses and increased fire risk. An estimated 30% of the mixed
conifer forest, based on lichens, was in exceedance of the CL. Forested and chaparral
communities were less sensitive. Only 3-15% of the forested and chaparral areas were
estimated to be in exceedance of the CL for NOs leaching. By combining the exceedance
areas of the seven vegetation types, Fenn et al. (2010) estimated that 25% of the land area
of the vegetation types included in the study were in exceedance for protecting against
nutrient enrichment.
Pardo etal. (2011c) conducted a synthesis study that evaluated published data to identify
the empirical CLs for N in Level I ecoregions across the U.S. The authors estimated CL
values as low as 3 kg N/ha/yr to protect sensitive resources in the southern California
case study area. Table 16-30 lists the CLs identified by Pardo etal. (2011c) for North
American Desert and Mediterranean California ecoregions (Omernick Level I) and new
studies published after Pardo et al. (2011c). If deposition exceeds the CL, the risk of
harmful effects increases.
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< umfwrMir < ritk^l I i uhI:
IliKb-fnd Ihmkvkl I irmlvncf
CL = critical load; ha = hectare; kg = kilogram; N = nitrogen; yr = year.
When CLs based on two different N responders were used within a given vegetation type (e.g., lichen effects and nitrate leaching),
only the more sensitive responder was used (lowest CL). The higher value of two CLs, when two values were available, was used in
the case of coastal sage scrub, mixed conifer forest (lichen community effects), desert scrub and pinyon-juniper, and the lichen
effects CL was used for chaparral.
Source: Fenn et al. (2010). Adapted to show case study locations.
Figure 16-53 Composite critical load exceedance maps for all seven vegetation
types included in the study of Fenn et al. (2010) showing the
combined exceedance areas and the level of exceedance (kg
N/ha/yr).
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Table 16-30 Summary of recent empirical dose-response and critical load studies focused on the southern
California case study area and published since Pardo et al. (2011c).
Study Terrestrial	Aquatic	Location	Focus	CL/Exceedance
Mediterranean California Ecoregion Level I (Omernick)
3.1-39 kg N/ha/yr
The lowest critical load is 3.1
based on lichen tissue chemistry
above the clean site threshold.
Nitrate leaching and fine root
biomass critical load is 17, soil
acidification is 26 kg N/ha/yr,
and susceptibility to beetle
infestation is 39 kg N/ha/yr
3.1-14 kg N/ha/yr
3.1 kg N/ha/yr is a modeled
value for lichens, while
10 kg N/ha/yr is the value for
nitrate leaching
Pardo et al.
(2011a)
•
Mediterranean California
Ecoregion Level I (Omernick);
coastal sage scrub ecosystem
Exotic invasive grass cover,
native forb richness, arbuscular
mycorrhizal richness
7.8-10 kg N/ha/yr
Pardo et al.
(2011a)
•
Mediterranean California
Ecoregion Level I (Omernick);
serpentine grassland ecosystem
Annual grass invasion replacing
native herbs
6 kg N/ha/yr
Clark et al.
(2013)
•
Mediterranean California and
North American Desert
Loss of herbaceous plant
species
Species loss 1-30%
Ellis et al.
(2013)
•
SEKI and YOSE
Protection of lichens
Estimated CL 2.5 to
7.1 kg N/ha/yr
Pardo et al. •	Mediterranean California	Lichen chemistry and community
(2011a)	Ecoregion Level I (Omernick); changes, nitrate leaching, soil
mixed conifer forest ecosystem acidification, reduced fine root
biomass
Pardo et al. •	Mediterranean California	Nitrate leaching and changes in
(2011a)	Ecoregion Level I (Omernick); the lichen community
chapparal ecosystem
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Table 16-30 (Continued): Summary of recent empirical dose response and critical load studies focused on the
southern California case study area and published since {Pardo, 2011,
2804014@@author-year}
Study
Terrestrial Aquatic
Location
Focus
CL/Exceedance
Cox et al.
(2014)
•
Mediterranean California
Ecoregion Level I (Omernick);
coastal sage scrub ecosystem
Conversion from coastal sage
scrub to annual grassland
Estimated CL 11 kg N/ha/yrto
conserve coastal sage shrub
vegetation
Bvtnerowicz
et al. (2015)
•
Mediterranean California
Ecoregion Level I (Omernick);
chapparal ecosystem
Lichens in chaparral, San
Bernardino Mountains
Exceedances of
CL = 5.5 kg N/ha/yr for lichens
occurred mostly in western and
northern portions of the study
area
Bvtnerowicz
et al. (2015)
•
Mediterranean California
Ecoregion Level I (Omernick);
mixed conifer forest ecosystem
Mixed conifer forest
Exceedances of
CL = 3.1 kg N/ha/yr for mixed
conifer forest covered much of
San Bernardino Mountains
Allen et al.
(2016)
•
Mediterranean California
Ecoregion Level I (Omernick);
coastal sage scrub ecosystem
Rapid decline in mycorrhizal
biodiversity
CL = 11 kg N/ha/yr
North American Desert Ecoregion Level 1 (Omernick)
Pardo et al.
(2011a)
•
North American Desert
Ecoregion Level I (Omernick)
Protect herbaceous plants and
shrubs
CL 3 to 8.4 kg N/ha/yr
Simkin et
al. (2016)
•
North American Desert
Ecoregion Level I (Omernick)
Grass and forb decreasing
species richness
CL open canopy: 8.3-9.9
(mean = 9.2, n = 240)
CL closed canopy: 13.5-17.0
(mean = 16.5, n = 32)
Clark et al.
(2013)
•
Mediterranean California and
North American Desert
Loss of herbaceous plant
species
Species loss 1-30%
CL = critical load; ha = hectare; kg = kilogram; N = nitrogen; SEKI
= Sequoia and Kings Canyons National Parks; YOSE = Yosemite National Park; yr = year.
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Data compiled by Pardo etal. (2011c) suggested that ambient N deposition is higher than
the lower limit of the expected CL to protect against nutrient enrichment effects in some
of the national parks in the Sierra Nevada, mainly SEKI, and to a lesser extent, YOSE.
Potential CL exceedances were reported for the protection of mycorrhizal fungi, lichens,
herbaceous plants, and forest vegetation, and also to restrict NOs leaching in drainage
waters. Fenn et al. (2010) estimated low CL values (3-8 kg N/ha/yr) for mixed conifer
forests, chaparral, and oak woodlands. These low critical load estimates were driven by
the presence of highly N sensitive species (lichens) and N poor vegetation (annual
grasslands and desert scrub plant communities). Fenn et al. (2010) concluded that N
deposition at or above these critical load values might cause vegetation community
changes because the increased availability of N tends to favor invasion by non-native
grasses into arid land ecosystems.
Clark et al. (2013) estimated the loss of herbaceous plant species caused by atmospheric
N deposition based on a national CL database. Using the lower and more conservative
estimates of CLs, the ambient N deposition was judged to be in exceedance of nutrient N
CLs over extensive portions of the Mediterranean California and North American Desert
ecoregions. Estimated plant species losses ranged from less than 1 to 30%, but variability
was high and uncertainty in these estimates was substantial.
Modeling Studies
Dynamic and steady-state models have been used to estimate critical or target loads in the
southern California case study region. Key studies, highlighted in Table 16-31. have
focused on CLs of nutrient N and acidity.
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Table 16-31 Terrestrial critical and target load and exceedance modeling studies
in southern California.
Study
Location
Model
Focus
Mixed Conifer
Hurteau et al. (2009)
Teakettle
Experimental Forest
and YOSE
STELLA®7
Modeled how changing precipitation and N deposition
levels affect shrub and herb biomass production. Herb
cover growth rate greater at 12 kg N/ha/yrthan
24 kg N/ha/yr. Precipitation was more important driver for
shrub cover.
Fenn et al. (2015)
Sierra Nevada and
San Bernardino Mts.
DayCent
Modeled CLto protect against NO3" leaching; CL
ranged from 17 to 30 kg N/ha/yr.
Semiarid
Li et al. (2006)
SEKI
DayCent
Simulated N deposition of 7.7 kg N/ha/yr; increased
loss of N to NO3" export and gaseous NO.
Rao et al. (2010)
JOTR
DayCent
Effects on fire risk in creosote bush and pinyon-juniper
vegetation types. CL = 32 kg N/ha/yr (creosote bush)
and 39 kg N/ha/yr (pinyon-juniper). Wet areas having
low soil clay (6-14%) may have CL as low as
1.5 kg N/ha/yr.
Rao et al. (2010)
JOTR
DayCent
Modeled CL was determined as the N deposition load
at which fire risk began to increase exponentially above
the background fire risk. The CL for creosote bush
scrub was 2.1 kg N/ha/yr and for pinyon-juniper
woodland was 3.6 kg N/ha/yr.
CL = critical load; ha = hectare; JOTR = Joshua Tree National Park; kg = kilogram; Mts. = mountains; N = nitrogen; NO = nitric
oxide; N03" = nitrate; SEKI = Sequoia and Kings Canyons National Parks; YOSE = Yosemite National Park; yr = year.
16.6.3.1.2.1. Mixed Conifer
1	Hurteau et al. (2009) developed and tested a model to determine how changes in
2	precipitation and N deposition might affect shrub and herb biomass in mixed conifer
3	forests in Teakettle Experimental Forest (located between SEKI and YOSE) and YOSE.
4	They estimated the prescribed fire intervals needed to counteract high fuel loads.
5	Increased understory (especially grass) biomass enhances fuel connectivity and fosters
6	larger and more severe fires. Under a model scenario that specified higher interannual
7	variability in precipitation and increased N deposition, model results suggested that
8	implementing fire treatments at an interval approximately equivalent to the historical
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range (15-30 years) would likely be needed to maintain understory vegetation fuel loads
at levels comparable to the control.
Based on 28 streams draining California mixed conifer forests studied by Fenn and Poth
(1999). Fenn et al. (2015) estimated that peak NOs concentrations in runoff were usually
less than 14.3 |icq/L: therefore, this level of NO3 leaching was identified as the likely
critical threshold value to identify watersheds that exhibit signs of N saturation. Fenn et
al. (2015) specified the empirical CL for protecting against NO3 leaching in California
mixed conifer forests equal to the N deposition level at which this identified critical
surface water NO3 concentration is exceeded (Fenn et al.. 2015; Fenn et al.. 2008). The
CL was derived using linear regression of stream water NO3 concentrations during the
winter high flow period and annual throughfall N deposition at 11 locations in the
southern Sierra Nevada and San Bernardino mountains. The calculated CL to protect
against NO3 leaching at the sites (6-71 kg N/ha/yr deposition) was 17 kg N/ha/yr (95%
CI 15-19 kg N/ha/yr).
16.6.3.1.2.2. Semiarid Shrublands
Li et al. (2006) used the DayCent model to quantify water, C, and N fluxes for a
chaparral ecosystem in SEKI that received about 7.7 kg N/ha/yr of atmospheric N
deposition. Simulated NO3 export and gaseous N emissions (mostly as nitric oxide
[NO]) generally agreed with observations. The model projections suggested that
increased N deposition would likely increase the flux of N from terrestrial ecosystems as
N03 in streams and gaseous NO. The authors concluded that the representations within
the DayCent model of N mineralization, runoff, and NO3 export in chaparral or similar
semiarid vegetation needed improvement.
Rao et al. (2010) used DayCent to estimate CLs of N deposition in JOTR, with focus on
the effects of N input on fire risk in creosote bush (Larrea tridentata) and pinyon-juniper
vegetation communities. Fire risk was expressed as the probability that annual biomass
production would exceed the risk threshold of 1,000 kg/ha of biomass available for
burning. Critical load was calculated as the N deposition at the point along the deposition
gradient where modeled fire risk began to increase exponentially. Mean estimated CLs
across multiple soil types, receiving precipitation of less than 21 cm/year, were 3.2 and
3.9 kg N/ha/yr for creosote bush and pinyon-juniper plant communities, respectively.
Critical loads decreased (more nutrient-sensitive) with decreasing soil clay content and
increasing precipitation. The wettest areas that had low clay content (6 to 14%) had
estimated low CLs, as low as 1.5 kg N/ha/yr (Rao et al.. 2010). These values fall at the
lower end of the CL range identified by Pardo et al. (2011c) and Pardo et al. (2011a) for
herbaceous vegetation in North American Deserts. Fire risks in the two vegetation types
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were highest under moderately high N deposition (9.3 and 8.7 kg N/ha/yr, respectively);
above those deposition levels, modeled fire risk was influenced more by precipitation
amount than by N supply.
16.6.3.1.2.3. Arid/Semiarid
DayCent modeling results suggested a CL less than 8.2 kg N/ha/yr for low-elevation
desert containing invasive Mediterranean grass (Schismus barbatus) and less than
5.7 kg N/ha/yr at higher elevation sites containing non-native red brome (Bromus rubens;
Rao et al.. 2010). Invasive non-native grass production in JOTR was simulated under
varying levels of N input and precipitation. Results suggested changing fire frequency.
Simulated fire risk was higher when N deposition was above 3 kg N/ha/yr, but then
stabilized at N deposition above 5.7 kg N/ha/yr in pinyon-juniper and at 8.2 kg N/ha/yr in
creosote bush scrub plant communities (Rao et al.. 2010). Model results of this study also
suggested that fire risk is controlled more by precipitation than by grass productivity
(Allen and Geiser. 2011).
16.6.3.2. Aquatic
This section highlights recent monitoring, dose-response, and critical load studies of N
and S deposition to aquatic systems in SEKI, JOTR, and similar southern California
ecosystems.
16.6.3.2.1. Acidification
In the Sierra Nevada, both S and N deposition can contribute mobile acid anions (SO42 ,
NO3 ) to watershed soil solution, streams, and lakes. Many aquatic and terrestrial
ecosystems in the Sierra Nevada are generally sensitive to acidification. Lakes in these
mountains are highly sensitive to impacts from acidic deposition because of the granitic
bedrock, thin acidic soils, large amounts of precipitation, coniferous trees, and dilute
surface waters (Mclack and Stoddard. 1991; Melacketal.. 1985; McColl. 1981). Because
the levels of acidic deposition at the higher elevations have been relatively low, however,
acidification effects to date have been minimal.
Many lakes in SEKI have received N deposition high enough to cause limited chronic
N03 leaching, which can contribute to both acidification and eutrophication. Widespread
chronic lake or stream acidification has not occurred to any degree, although Sullivan
(2000) concluded that some episodic acidification has likely occurred. Acid anion
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concentrations (mainly SOr and NOs ) in most high-elevation lakes were low in surveys
conducted during summer and fall, but higher during spring when snowmelt commonly
causes pulses of high NOs concentrations in surface waters (Melack et al.. 1989).
During the fall of 1985, about one-third of the lakes sampled by the Western Lake Survey
(WLS) in and near SEKI had acid neutralizing capacity (ANC) < 50 |icq/L. and two of
the lakes in the park had ANC < 20 (j,eq/L (Landers etal.. 1987). The possibility of
recovery from lake acidification during the period 1985-1999 was evaluated by another
chemical survey conducted by the U.S. Geological Survey (USGS) during the fall of
1999 (Clow and Sueker. 2000). Clow et al. (2003) reported that the resampled lakes were
generally low in ionic strength and had pH between about 6.0 and 7.0; the median ANC
was 59 (j,eq/L in SEKI and 32 |icq/L in YOSE located to the north. The observed lake
chemistry further confirmed the high sensitivity to acidification.
Key acidification characterization and monitoring studies conducted in the Sierra Nevada
are listed in Table 16-32. None of these studies are recent (post-2008) other than the
study of Clow et al. (2010) in YOSE. The current condition of these waters is probably
similar to what was found during previous decades with no additional recovery.
The weight of evidence suggests that many high-elevation lakes in SEKI have in the past
received N deposition high enough to cause some chronic NO3 leaching. This can
contribute to both acidification and eutrophication. Widespread chronic lake or stream
acidification has not occurred to any degree, although Sullivan (2000) concluded that
some episodic acidification has likely occurred. Acid anion concentrations (mainly SO42
and NO3 ) in most high-elevation lakes were low during summer and fall, but higher
during spring when there are commonly pulses of high NO3 concentrations in surface
waters during snowmelt (Melack et al.. 1989). which is seldom sampled due to logistical
and safety constraints. Acid-base chemistry is not expected to have changed appreciably
in recent years.
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Table 16-32 Example surface water acidification studies in Sequoia and Kings
Canyons National Parks and other Sequoia and Kings Canyons
National Parks—relevant areas in the southern California case study
region.
Study
Location
Time
Period
Focus
Clow et al. (2003) Clow
and Sueker (2000)
National parks in western
U.S., including SEKI
1999
Resurvey of WLS lakes
Clow et al. (2003)
Sierra Nevada region
1985 and
1999
Decrease in ANC in lakes in SEKI partly due
to climatic differences at time of sampling
Clow et al. (2010)
YOSE
2003
Multiple linear regression modeling to
predict NO3" concentrations and ANC in
lakes across the park
ANC = acid neutralizing capacity; N03 = nitrate; SEKI = Sequoia and Kings Canyons National Parks; WLS = Western Lake
Survey; YOSE = Yosemite National Park.
The hydrology of alpine and subalpine ecosystems in the Sierra Nevada is controlled by
snowmelt because more than 90% of the annual precipitation in this region falls as snow.
The relatively small loads of acidic deposition can contribute to moderately high
concentrations of SOr and NOs in lakes and streams during the early phase of
snowmelt (Stoddard. 1995). Potential biological effects of acidic deposition on surface
waters in the Sierra Nevada are most likely caused by acidification attributable to high
N03 concentrations, which tend to be episodic rather than chronic (Wigington ct al..
1990).
Climatic fluctuations that control the amount and timing of precipitation, melting of the
snowpack, growth of plants, depth to groundwater, and concentration by evaporation of
constituents in solution all influence soil and surface water chemistry. These factors
regulate interactions between air pollutants and the aquatic and terrestrial biological
receptors. Because conditions vary over time, many years of data might be required to
establish the existence of trends in episodic surface water chemistry (Sullivan. 2000).
Effects of acidification on vertebrate animals in the Sierra Nevada have not been
documented. However, declines and elimination of frogs and toads due to uncertain
causes have been documented throughout the western U.S., including in the national
parks in California. Possible impacts of air pollution on aquatic amphibians is noteworthy
because amphibians are especially sensitive to environmental change or degradation
(Bradford and Gordon. 1992; Blaustein and Wake. 1990).
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Arid and semiarid ecosystems in southern California are generally not known to be
sensitive to acidification from either S or N deposition. Precipitation to these ecosystems
is limited. There are relatively few surface waters, and it appears that they have
appreciable acid buffering capacity. However, semiarid chapparall soils under very high
N deposition in the San Gabriel Mountains have decreased considerably in pH between
the 1970s and 1990s (Fenn et al.. 2011a; Pardo et al.. 201 lc).
16.6.3.2.1.1.	Empirical Studies
Empirical studies have shed light on dose-response relationships in lakes and streams of
southern California. Several new studies are addressed here. Heard et al. (2014)
investigated CLs to protect against lake acidification in the Sierra Nevada. An
experimental study by Kratz et al. (1994) investigated responses of macroinvertebrates.
Earlier studies of effects of N deposition in this region (mainly before 2008) addressed
aspects of water chemistry and the response of algae to nutrient addition. More recent
publications include the reviews of Baron et al. (2011a). Pardo etal. (2011c). and Fenn et
al. (2015).
Heard et al. (2014) investigated the likelihood that paleoreconstructed changes in ANC in
lakes of the Sierra Nevada had been caused by atmospheric deposition of SOr and NOs
during the 20th century. The deposition estimates suggested that the CL at Moat Lake
was exceeded in about 1920 and the chemical recovery started in about 1970. The initial
inferred decline in the ANC of Moat Lake occurred between 1920 and 1930. This time
period corresponded with estimated acidic deposition (SO42 + NO3 ) equal to about
74 eq/ha/yr. This was taken (Heard et al.. 2014) to be the critical load (CL) to protect this
lake against acidification. Diatom-reconstructed ANC values changed from near
100 |icq/L before 1920 to near 60 |icq/L during the 1970s. Recovery of ANC after 1970
was attributed by Heard et al. (2014) to decreased deposition of S. In addition, during the
late 20th century, warmer air temperatures may have contributed to higher ANC via
increased weathering rates.
16.6.3.2.1.2.	Modeling Studies
Some limited work has been conducted on modeling of aquatic critical loads using
steady-state approaches for acidity. Data from over 12,500 streams and lakes were used
by the Critical Loads of Atmospheric Deposition (CLAD) science committee of the
NADP (http://nadp. sws .uiuc.edu/committees/clad/) to model steady-state CLs for acidity
of surface waters based on multiple approaches for estimating base cation weathering.
Shaw et al. (2014) used the Steady-State Water Chemistry (SSWC) model to estimate the
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steady-state CL of acidity for 208 Sierra Nevada lakes. Study lakes were generally dilute
(mean specific conductance 8 (j,S/cm) and had relatively low ANC (mean 57 |icq/L).
Most were located in Forest Service wilderness areas, some in proximity to Sierra
Nevada national parks. Using a critical ANC limit of 10 (j,eq/L to protect aquatic biota
from effects caused by water acidification, the model predicted a CL of 149 eq/ha
(14.9 meq/m2) of acidity for Sierra Nevada watersheds situated on granitic bedrock. More
than one-third of the study lakes received acidic deposition that was in exceedance of the
estimated CL. The median study lake showed a CL exceedance of about 80 eq/ha
(8 meq/m2/year). Based on these CL calculations, Shaw et al. (2014) concluded that
high-elevation lakes in the Sierra Nevada have not fully recovered from effects of
acidifying deposition despite large decreases in S and N air pollution and reduced S
deposition over the last several decades.
Aquatic Nutrient Enrichment
There have been several studies of water chemistry dose-response relationships that
pertain to nutrient enrichment in the Sierra Nevada case study area. These have included
studies of N retention and release, critical loads, and N saturation. A meta-analysis of
lakes from 42 regions of Europe and North America concluded that atmospheric N
deposition has caused higher concentrations of NOs in lake water and increased
phytoplankton biomass (Bcrgstrom and Jansson. 2006). Bergstrom et al. (2005) found N
limitation in lakes that received N deposition below approximately 2.5 kg N/ha/yr,
limitation of N and P at N deposition between -2.5 and 5.0 kg N/ha/yr, and P limitation
in lakes that had N deposition higher than about 5.0 kg N/ha/yr. These findings may be
relevant to remote lakes in the Sierra Nevada, including those in SEKI.
Sickman et al. (2002) reported yields and retention of dissolved inorganic N (DIN) at
28 high-elevation lakes, including those within SEKI. Net DIN retention was
1.2 kg N/ha/yr, an estimated 55% of annual DIN loading. Sickman et al. (2001)
calculated that the annual yield of N from the Emerald Lake watershed varied by a factor
of 8 (0.4 to 3.2 kg N/ha/yr), which explained 89 and 74% of the variation in DIN and
organic N, respectively. The N content of runoff was higher during years that
experienced high runoff and was lower during dry years. Increases in NOs concentration
were larger during years with deep, late-melting snowpacks. Therefore, it is expected that
climate warming, with associated earlier snowmelt, might increase N retention in
high-elevation watersheds (Sickman et al.. 2001).
Nutrient enrichment of Sierra Nevadan lakes is not solely a function of N inputs.
Atmospheric P deposition can also be important. Sickman et al. (2003) proposed that
atmospheric P deposition and enhanced cycling of P due to climate change were the most
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likely sources of the observed increase in P loading to the Sierra Nevada lakes. It is not
known why atmospheric deposition of P to these lakes has increased over time.
Possibilities include increased use of organo-phosphate pesticides (Kcglcv et al.. 2000)
and wind-blown transport of soils and dust that are high in P from the agricultural San
Joaquin Valley to the mountains (Sickman et al.. 2003; Lesack and Melack. 1996;
Bergametti etal.. 1992).
Based on results of bioassay experiments and application of coarse nutrient limitation
indices, Sickman et al. (2003) concluded that phytoplankton growth in Emerald Lake in
SEKI during the early 1980s was limited by P supply. During subsequent years, in
response to increased atmospheric P input, the nutrient balance shifted into a pattern of
colimitation by N + P, and finally N limitation. Sickman et al. (2003) observed that the
total P concentration in the lake doubled from 1983 to 1999. Particulate C concentration
in the lake in 1999 was four times higher than the average during the period 1983-1998.
Together with the observed increase in total P in lake water, this suggested that
eutrophication had occurred during that time (Sickman et al.. 2003). It was proposed that
increased P loading to Sierra Nevada lakes decreased NOs concentration in lake water
and promoted N limitation. The increased P loading may have been due to increased
emissions and deposition of P and/or accelerated internal P cycling that might be caused
by changes in climatic conditions and the timing of snowmelt and runoff (Sickman et al..
2003; Johnson. 1998: Dettingerand Cavan. 1995).
Baron etal. (2011a) synthesized CLs of N deposition for protecting against nutrient
enrichment of high-elevation lakes. Relationships between NO3 concentrations and N
deposition suggested a CL near 2 kg N/ha/yr to prevent NO3 leaching to lakes in the
Sierra Nevada. Fennetal. (2011a) also summarized data reflecting the relationship
between atmospheric N deposition and stream NO3 concentration, in this case for
chaparral vegetation in the San Dimas Experimental Forest in the San Gabriel Mountains
(Meixner et al.. 2006; Riggan et al.. 1994; Riggan et al.. 1985). Devil Canyon watershed
in the western San Bernardino Mountains (Meixner and Fenn. 2004; Fenn and Poth.
1999). and Chamise Creek in SEKI (Fenn et al.. 2003a; Fenn et al.. 2003c). Stream water
N03 concentrations in streams draining chaparral vegetation were among the highest
reported in North America, with concentrations in some streams higher than 200 |icq/L
(Fennetal.. 2011a; Fenn et al.. 2003a; Fenn et al.. 2003c; Fenn and Poth. 1999; Riggan et
al.. 1994; Riggan et al.. 1985). Evaluation of N saturation in chaparral must also consider
the prevalence of fire in this ecosystem type because the stand-replacing fire interval is
only about 40 to 60 years (Minnichand Bahre. 1995). Much of the atmospherically
deposited N that has accumulated in the litter and aboveground vegetation over the
intervening years is released suddenly to the atmosphere during fire. In contrast, much of
the accumulated N in the mineral soil generally remains on-site after a fire (Fenn et al..
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201 la), and fire contributes to increased NO3 concentrations in stream water (Riggan et
al.. 1994; Anderson and Poth. 1989).
The recent work that has been conducted on aquatic ecosystems has focused mostly on
the Sierra Nevada region. Analogous nutrient enrichment work has not been conducted in
or around JOTR, in part, because there are few surface waters in that region.
16.6.3.2.2.1. Algae
Changes in nutrient supply can affect aquatic biota at all trophic levels, but algae are
perhaps most likely to show eutrophication effects from N deposition. Studies have
shown an increase in lake phytoplankton biomass in response to increases in N deposition
in the Sierra Nevada region (Sickman et al.. 2003). Wyoming (Lafrancois et al.. 2003b').
Sweden (Bcrgstrom et al.. 2005). and across Europe (Bcrgstrom and Jansson. 2006).
Increases in algal biomass have been associated with changes in algal assemblages that
favor certain species over others. A widespread increase in the relative abundance of
Asterionella formosa and Fragilaria crotonensis occurred in oligotrophic lakes across the
western U.S.; these changes have been documented from both lake sediment cores and
limnological surveys (Saros et al.. 2003; Wolfe et al.. 2001; Interlandi et al.. 1999;
Goldman. 1988). In Lake Tahoe to the north of YOSE, there has been an increase since
about 1950 in the ratio of araphidinate pennate diatoms to centric diatoms. This was
largely due to increased abundance of Fragilaria crotonensis. This change in diatom
relative abundance was associated with higher N loading to the lake (Goldman. 1988).
Jassbv et al. (1994) showed that atmospheric deposition supplies most of the N to Lake
Tahoe.
Because high-elevation lakes in the Sierra Nevada tend to be highly oligotrophic, small
changes in nutrient supply can affect algal productivity (Sickman et al.. 2003). Chamise
Creek in SEKI was reported to have high NO3 leaching in response to throughfall N
deposition near 10 kg N/ha/yr, suggesting N saturation (Fenn et al.. 2003a; Fenn et al..
2003c). The U.S. Forest Service has suggested a policy threshold of 2 |icq/L for stream
N03 concentration in N limited ecosystems in the western U.S. This load has been
suggested as a tipping point to trigger management concern for possible over-enrichment
of aquatic ecosystems with N (Fenn et al.. 2011b).
16.6.4. Highlights of Additional Research Literature and Federal Reports since
January 2008
Key research literature published since January 2008 is highlighted in Table 16-33.
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Table 16-33 Key recent research literature focused on the case study region.
Publication
Focus
Allen and Geiser(2011)
N addition experiment in JOTR
Allen et al. (2009)
Soil N along depositional gradient in JOTR
Baron et al. (2011a)
Critical load and dose-response review
Bowman et al. (2011)
Critical loads for high-elevation vegetation
Bvtnerowicz et al. (2015)
Empirical N critical loads in San Bernardino Mts.
Clow et al. (2010)
Lake acid-base chemistry characterization in YOSE
Clark etal. (2013)
Loss of herbaceous plant species in response to N deposition
Cox etal. (2014)
Conversion from coastal sage scrub to grassland
Ellis etal. (2013)
Empirical critical loads
Esque et al. (2010)
N dynamics after fire in Mojave Desert shrubs
Fenn et al. (2008)
Effects of N on epiphytic lichens and stream nitrate leaching
Fenn etal. (2010)
Empirical critical loads and exceedances
Fenn etal. (2011a)
Critical loads review
Fenn etal. (2015)
Critical loads review
Geiser et al. (2010).
Effects of N on lichens
Grulke et al. (2008)
Forest susceptibility to wildfire in San Bernardino Mts.
Heard etal. (2014)
Critical loads for lakes
Hurteau and North (2009)
Effects of N on a sensitive herbaceous plant species
Johnson et al. (2011)
Soil acid-base chemistry in mixed conifer watersheds
Jovan (2008)
Response of lichens to N
Kimball etal. (2014)
Effects of water and N on CSS following fire
McCallev and Sparks (2009)
Soil temperature effects on N loss in Mojave Desert
Pardo et al. (2011c)
Critical load and exceedance for nutrient N enrichment
Pasauini and Vourlitis (2010)
Net primary production in chaparral
Rao et al. (2009)
Effects of N deposition on mineralization
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Table 16-33 (Continued): Key recent research literature focused on the case study
region.
Publication
Focus
Rao et al. (2010)
N and fire effects on semiarid plant communities
Rao et al. (2011)
Critical loads for desert vegetation
Shaw et al. (2014)
Critical loads for lakes in Sierra Nevada
Stark etal. (2011)
Effects of N deposition and climate on Mojave Desert vegetation
Talluto and Sudina (2008).
Conversion of CSS to grassland
Vamstad and Rotenberrv (2010)
Changes in plant species composition after fire
Vourlitis (2012)
Response of semiarid vegetation to N and climate
Vourlitis and Fernandez (2012)
N response to N input in semiarid shrublands
Vourlitis and Pasquini (2008)
C and N dynamics in pre- and post-fire chaparral
Vourlitis and Pasauini (2009)
Dry season addition of N to CSS and chaparral
C = carbon; CSS = coastal sage scrub; JOTR = Joshua Tree National Park; Mts = mountains; N = nitrogen; YOSE = Yosemite
National Park.
16.6.5. Summary
1	This case study focuses on the ecosystems found in two national parks in southern
2	California, SEKI and JOTR. SEKI largely consists of forested and mountainous terrain in
3	and near the Sierra Nevada Mountains. JOTR is covered largely by desert and semiarid
4	land. Both are generally downwind of air pollution (mainly N) sources. The majority of
5	the information provided in this case study is applicable to ecosystems found within
6	SEKI. Limited research conducted in JOTR was found in the peer-reviewed literature. A
7	summary of studies relevant to evaluating CLs is shown by Figure 16-54.
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m
on
c
"o
Q.
~a
c
<
>-
"a
a
| Manne West Coast Forests-Lichen protection (wet deposttio«)-Geiser et a!., 2010
JOTR-Wetted areas with tow day content-Rao et al.f 2010
| Sierra Nevada Mtns.-Prevent N03- leaching-Fenn et al.t 2011a
JOTR-Creosote bush scrub-Fenn et at., 2008,201S
SEKI and YOSE-Lichen protection-Ellrc et a!.» 2013; Pardo et al., 2011a
California-Protection of 7 vegetation types-Fenn et al,, 2010
Sierra Nevada, mainly SŁKI-Protecnon of mycorrb«al fungi, lichens, herbaceous plants and forest vegetation
and to limit N03- leaching-Pando et al, 2011b
I JOTR-Lichen and herbaceous plants-Pardo et al. 2011b
| Marine Wesi Coast Forests-Lichen protection (total deposition!-Geiser
et al., 2010
Western Sierra Nevada-Epiphytic lichen shift-Bowman et at, 2011. Fenn et al., 2008
San Bemadino Mtns.-Mtxed conifer forest-8ynerowicz et al., 201S
JOTR-Pinyon-juniper-Fenn et al., 2008, 2015
| San Bemadino Mtns.-Lie hen protection-Bynerowicz et al.. 201S
San Bemadino Mtns.-Lichen protection in chapparal ecosystems-Fenn et al., 2010
Calrfornia-Lichen protection; Lichen presence at 3 of S3 lichen survey sites et al., 2011a
JOTR-Higher elevation sites containing non-native red brome-Rao et al., 2010
JOTR-Pinyon-juniper protection-Rao et al., 2010
I SEKMncreased the losses of N as NOB- in streams and gaseous NO in chapparal or
similar semi-and vegetation-ti et al., 2006
~
~
~
~
~
JOTR-lnvasive grasses-AJIen et al., 2009; Allen and Geiser, 2011
JOTR-Creosote bush scrub communities-Rao et al., 2010
JOTR-Low elevation desert containing invasive Mediterranean grass
Raoetal., 2010
~ N. American deserts-Fire risk in I vegetation type-Pardo et al.,
2011a, 2011b
~
N. Amencan deserts-Fire nsk in 1 vegetation type-Pardo et
al. 2011a, 2011b
~ Western Sierra Nevada Oligotrophy lichen elimination
Bowman et al., 2011
¦ So. California-Coastal sage scrub-Cox
et al., 2014
~ So. California -Conversion from coastal sage
scrub to annual grassland-Cox et al, 2014
YOSE-Herbaceous species-Hurteau and North, 2009
~
Semi-arid shrubland-Shrub growth and relative abundance of ["
sub-dominant shrubs and herbaceous plants-Pasquini and L_
Vourlitts, 2010
San Bemadino Mtns.-N03- leaching in chapparal
Fenn et al., 2010
~
San Bemadino MIns.-Protection against N03- leaching in
mixed conifer forests-Bynerowia et al.f 201S
Sierra Nevada and San Bemadino Mtns.-Protection against
N03- leaching Fenn et al.. 2008,201S
So. California-Mixed conifer forest-Breiner, 2007
JOTR-Creosote bush-Rao et al., 2010
JOTR-Pin yon-juniper-Rao et al.,
2010
~5	7	g	5	nr
N deposition, Kg N/ha/yr
~TI	IT
TT
14 >14
ha = hectare; JOTR = Joshua Tree National Park; kg = kilogram; Mtns. = mountains; N = nitrogen; N. = north; NO = nitric oxide;
N03" = nitrate; SEKI = Sequoia and Kings Canyons National Parks; So = southern; YOSE = Yosemite National Park; yr = year.
Figure 16-54
Continuum of critical loads in southern California case study area
and relevant surrounding region.
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The Sierra Nevada Mountains, in which SEKI is located, contain many acid-sensitive
aquatic resources. Limited monitoring studies indicated that the current condition of these
waters is probably similar to that found during previous decades with no additional
recovery. Aquatic acidification effects to date on SEKI's condition have been minimal at
higher elevations due to low acidic deposition, and widespread lake or stream
acidification has not occurred to any degree, although some episodic acidification has
likely occurred. There are expectations for N retention in high-elevation watersheds as
snowmelt changes temporally due to climate warming. Small changes in nutrient supply
to high-elevation, oligotrophic lakes in the Sierra Nevada can affect algal productivity
and increase abundance of Asterionellaformosa and Fragilaria crotonensis (Saros ct al..
2003). Amphibian decline and the possible association with air pollution is noteworthy.
However, one of the most pronounced effects of N deposition in the Sierra Nevada has
been shown to be an alteration of the lichen community which has CLs ranging from 3.1
to 5.2 kg N/ha/yr (Bowman et al.. 2011; Fenn et al.. 2008).
JOTR's arid and semiarid ecosystems are not known to be sensitive to acidification
because there is limited precipitation, few surface waters, and high acid buffering
capacity. N deposition may increase productivity of invasive grasses after fire, and
deposition and climate change may promote an altered fire cycle that does not allow
sufficient time for Joshua tree woodlands to re-establish between fires. It has been
suggested that N deposition may have already exceeded JOTR's critical load of
8.4 kg N/ha/yr for terrestrial plant communities (Pardo et al.. 2011c).
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