-------
co
f
I
i
CC.
30-
20-
10-
o-
-10-
-20
-30
o
o
o
o
o
• Bulk Precipitation
o Integrated Forest Study
•o
20 40 60
Atmospheric N Input (kg/ha/year)
80
Figure 10-11., Ecosystem nitrogen retention as a function of atmospheric nitrogen input.
Source Johnson (1992)
Indeed, it appears as if the soil is being "mined" for the nitrogen necessary to supply
vegetation increment systems with very low atmospheric nitrogen inputs This is readily
apparent when nitrogen output is plotted as a function of input minus vegetation increment
(Figure 10-14) Input minus increment can be thought of as nitrogen that is available for
either (1) soil heterotroph uptake or (2) nitrate leaching A negative value for input-
increment implies that either the soil is being "mined" for nitrogen to supply tree needs or
that there is an unmeasured nitrogen input contributing to tree nitrogen needs In either case,
the data suggest that, contrary to views expressed in the literature (see review above), trees
are actually more effective competitors for nitrogen than soil heterotrophs under
10-61
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80-
60—
I
|
40 —
20-
O
• Bulk Precipitation
o Integrated Forest Study
o
go o
I ' I ' I ' I
10 20 30 40
Atmospheric Input (kg/ha/year)
I '
50
60
Figure 10-12. Tree nitrogen increment as a function of atmospheric nitrogen input.
Source Johnson (1992)
nitrogen-deficient conditions Also, the nearly 1 1 relationship between nitrogen output and
input-increment after the latter exceeds zero (r = 0 84) indicates that nitrogen deposited in
excess of vegetation needs is not taken up by heterotrophs, but rather is subject to
nitrification and nitrate leaching, perhaps because heterotrophs in these systems are limited
by organic substrates or other nutrients
There are several possible explanations for the rather stnlong differences in soil
nitrogen retention and loss patterns between fertilizer and nutrient cycling/air pollution
studies. First, heterotrophic demand for nitrogen in fertilized sites is likely to be greater
than in sites subjected to chronically elevated atmospheric nitrogen inputs Fertilizer
10-62
-------
20-
(0
-20
o
V)
1
3
<3 -40-
-60-
o
o
o
o
o
• o
• Bulk Precipitation
O Integrated Forest Study
O
10
60
N Deposition (kg/ha)
Figure 10-13. Calculated soil nitrogen retention (input-increment-leachiug) as a function
of atmospheric nitrogen input.
Source Johnson (1992)
nitrogen is typically applied to nitrogen-deficient ecosystems, where nitrogen demand by soil
heterotrophs is likely to be high, whereas heterotrophic demand for nitrogen may have been
substantially satisfied in sites with chronically high atmospheric nitrogen inputs
Heterotrophic activity in fertilized sites is also likely to be stimulated by mobilization of soil
organic carbon, which typically occurs after fertilization (especially with urea, Ogner, 1972,
Foster et al , 1985a) Second, as noted above, the slow, steady inputs of nitrogen via air
pollution, like slow, steady inputs of fertilizer nitrogen, probably favor nitrification Third,
nonbiological retention of nitrogen is likely to be greater with fertilization than atmospheric
deposition Ammonium and NH3 fixation in 2 1 clays is likely to be substantially increased
10-63
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80
Net Loss from Soil
60 —
JS
i
Integrated Forest Study
Bulk Precipitation
40 —
20 —
-60
-20 0 20 40
Input Minus Vegetation Increment (kg/ha/year)
60
Figure 10-14. Nitrogen leaching as a function of atmospheric nitrogen input minus tree
nitrogen increment. Points above the 1:1 line imply net soil loss, and
points below the line imply net soil retention.
Source Johnson (1992)
under conditions of high concentrations of one or both following fertilization It has also
been shown that NH3 can react chemically with soil organic matter to form very stable,
nonlabile compounds (Foster et al, 1985b) Conditions following urea fertilization are
especially conducive to these reactions in that pH is increased and NH3 concentrations are
high. These conditions would not normally occur in sites subject to chronically high
atmospheric nitrogen inputs In summary, ecosystems retain a greater amount of
atmospherically deposited nitrogen than of fertilizer nitrogen, however, no observable
relationship exists between atmospherically deposited nitrogen and either tree increment or
calculated soil retention It appears that mtnfiers may not be as poor competitors for
nitrogen as was previously suspected, particularly in cases where nitrogen inputs are
10-64
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increased in small, frequent doses, such as with air pollution Heterotrophs appear to be the
most effective short-term competitors for nitrogen in nitrogen-poor sites, but trees appear to
be the most effective competitors for nitrogen over the longer term, as indicated by the
apparent mining of nitrogen from soils where atmospheric nitrogen inputs are low and tree
nitrogen requirements are high
10.5.4 Effects of Pollutant Nitrogen Inputs on Soils
10.5.4.1 Soil Biota
The most obvious and immediate effects of pollutant nitrogen inputs on soils are
those on the microbial community An increase in the activity of heterotrophs and mtnfiers
associated with an increase in decomposition and nitrification might be expected in response
to nitrogen inputs Studies of microbial responses to nitrogen fertilization have produced
mixed results, however Kelly and Henderson (1978) found increased bacterial activity, but
reduced invertebrate populations, 1 year after fairly high levels of urea fertilization (550 and
1,100 kg nitrogen/ha) This change was important because invertebrates play a major role in
the initial breakdown of litter However, the authors found little effect of fertilization on the
decomposition of white oak leaf litter Kowalenko et al (1978) found that fertilization with
NH4NO3 and potassium chloride caused a reduction in soil microbial activity (as measured
by CO2 evolution) for at least 3 years This may have been due to toxic or shock effects due
to very large increases in both nitrogen and other ions over a very short tune. Weetman and
Hill (1973) reviewed the effects of fertilization on soil flora and fauna and concluded that
fertilization had a lasting, stimulating effect despite short-term toxic effects of fertilizer
components (especially ammonium) Again, we must consider the effects of single, large
inputs of nitrogen, typical of fertilization studies, as opposed to the slow, steady inputs of
nitrogen at lower concentration typical of pollutant inputs Aside from the limited
information on the effects on mtnfiers, virtually nothing is known regarding the effects of
slow, steady inputs of nitrogen on soil microbial communities
10.5.4.2 Soil Chemistry
The foremost concern about long-term, capacity-controlled effects of excessive
nitrogen deposition and NO3" leaching is soil acidification and the mobilization of A13+ into
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soil solution and surface waters As a prelude to assessing the effects of excessive nitrogen
deposition on soil acidification and A13+ mobilization, a brief review of the components of
soil acidity and cation exchange processes is presented
Soil acidity can be measured in a number of ways, but for the purposes of this
discussion, we will refer to base saturation as the primary measure or indicator of soil
acidity. Base saturation refers to the degree to which soil cation exchange sites, negatively
charged sites to which positively charged ions are adsorbed, are occupied with base cations
(calcium ions [Ca2+], magnesium ions [Mg2+], and potassium ions [K+]) as opposed to
"5 aim I
Al and hydrogen ions (H ) Base saturation is a measure of soil acidification, with lower
values bemg more acid Figure 10-15 shows a soil with 50% base saturation on the left and
a soil with 10% base saturation on the right
Input
Input
Mineral
Weathering
Mineral
Weathering
Leaching
Leachmq
Figure 10-15. Schematic diagram of cation exchange for base cations, aluminum ions,
and hydrogen ions in circumneutral (50% base saturation, left) and acid
(10% base saturation, right) soils.
10-66
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Ulnch (1983) describes the various buffering ranges soils go through as they acidify
first is the base cation buffering range, where incoming acid and base cations are exchanged
primarily for base cations with very little H+ and Al + increase (Figure 10-15, left)
As soils acidify, exchangeable base cations are replaced by exchangeable A13+ and H+, and
soils are said to be in the aluminum buffering range (Figure 10-15, right) Incoming cations
(acid and base) are exchanged primarily for H+ and Al + in soils that are in the aluminum
buffering range (Figure 10-15, right)
With the use of a simulation model, Reuss (1983) showed that the transition from the
base cation to the aluminum buffering range is very abrupt His results showed that soil
3 +
acidification has little effect on the concentration of Al in soil solution over a large range
of base saturation values above 20% However, he noted that fairly minor changes in base
"> I
saturation within the 10 to 20% range can cause quite large increases in soil solution Al
concentration This implies that soils with base saturations of 10 to 20% are extremely
sensitive to change (although this does not necessarily imply that vegetation will respond to
soil change) A series of simple laboratory column studies could tell us much about how far
some of our forest soils are from the aluminum buffering range and how much additional
acid input might be required to put them into this range
Once soils are in the aluminum buffering range, the rate of base cation leaching will
obviously decrease because Al + is now a dominant cation in soil solutions In a soil free of
vegetation, continued inputs from the atmospheric deposition, which contains base cations as
well as H+, will eventually acidify the soil to the point where base cation outputs equal base
cation inputs With forest or other vegetation growing on the soil, however, continued base
cation uptake could reduce the base saturation of the sod to the point where export of base
cations is less than input by deposition (Figure 10-15, night) Thus, vegetation uptake can,
by depleting soil exchangeable base cations, cause the soil to begin accumulating base cations
even when the soil is subject to high leaching rates Of course, this accumulation of base
3 +
cations is accompanied by substantially increased leaching of Al , and the potentially
detrimental effects of the latter must be considered
The same cation exchange principles that will eventually cause a soil to begin
accumulating incoming base cations when soils acidify into the aluminum buffering range can
»y I 9-1- I
also cause an ecosystem to begin accumulating an individual cation (Ca , Mg , or K ) if
10-67
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tree uptake depletes soils of an individual cation (Johnson and Todd, 1987) In this case, the
conservation of the individual cation in question need not be accompanied by significant
*5 i
overall soil acidification and increased leaching of Al , leaching of the other base cations
may be increased instead Johnson et al (1985) noted such a situation with respect to Ca +
in an oak-hickory forest on the Walker Branch watershed in Tennessee In this ecosystem,
Q i Q i 0-1-
tree Ca is very high, soils are very low in exchangeable Ca , and consequently Ca
leaching is low Thus, the ecosystem shows a net Ca2+ gain from atmospheric inputs
O [ _i i
(accompanied by net losses of Mg , K , and sodium ions [Na ])
The greatest uncertainty in assessing and projecting rates of exchangeable base cation
depletion and/or soil acidification is the estimation of primary mineral weathering rates The
weathering of primary soil minerals (e g , hornblende, feldspar, plagioclase) represents an
input to the exchangeable base cation pool (Figure 10-16) Calculations of the potential rate
of soil change from exchangeable pools and input-output budgets (e g , Tomhnson, 1983)
represent the worst-case scenario, that is, they assume that weathering is zero A high rate
of soil leaching offset by a high rate of weathering results in a high rate of turnover, but not
a net depletion of exchangeable cations
Equations and simple models of soil weathering are available for primary to secondary
mineral transformations (e.g , Lindsay, 1979) However, these equations are of little value
for soils with sizeable nonexchangeable base cation reserves contained in ill-defined minerals
(such as amorphous iron [Fe] and aluminum [Al] oxides, Johnson et al, 1985) A further
complication arises when mineral weathering is enhanced by organic acids formed in forest
litter or exuded by tree roots (Boyle and Voigt, 1973) Thus, at present, there are only
empirical approaches to assessing weathering, such as mass balance calculations One mass
balance approach involves measuring fluxes and changes in exchangeable cation pools over
time and calculating weathering, by difference (Matzner, 1983) A simpler mass balance
approach is to estimate the total weathering loss from a soil by the difference in soil element
content at present and that of an equivalent amount of primary minerals (i e , element content
at the time the soil began to form) and divide by the amount of time the soil has been
exposed to weathering (e g , since the last glaciation) (Mazzanno et al , 1983) The latter
gives an average weathering rate over geologic tune, but it does not represent current
weathering rates in the soil The former method gives a better estimate of current
10-68
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Low Input
High Input
Mineral
Weathering
I
Mineral
Weathering
Leaching
Leaching
Figure 10-16. Schematic diagram of cation exchange for base cations, aluminum ions,
and hydrogen ions hi acid soils with low (right) and high (left)
atmospheric deposition rates.
weathering rates in the soil, but it is subject to large uncertainties due to errors in each of the
estimates used to calculate it Nonetheless, the plot-scale mass balance method, although
imprecise, seems the best for obtaining realistic estimates of current soil weathering rates,
especially in systems where leaching has been increased by artificial acid irrigation (Stuanes,
1980)
Because forest soils acidify naturally, it must be true that weathering rates do not keep
pace with base cation denudation rates, even under pristine conditions The relative
contribution of acid deposition to the rate of acidification can be assessed by measuring
element fluxes (Ulrich, 1980, Matzner, 1983, Johnson et al, 1985), and the actual
magnitude of the acidification rate, which equals base cation export minus weathering input,
can be estimated by measuring changes in exchangeable base cations and acidity through tune
10-69
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(taking into account seasonal variations in surface soils, see Haines and Cleveland, 1981)
The effects of excess nitrogen and S deposition on the rate of soil acidification cannot be
evaluated by simply measuring changes in soils through tune, however, because the natural
rate of soil acidification (via natural leaching and vegetation uptake) cannot be accounted for
by simply measuring changes in soils If soils do not change during the measurement period,
it can be stated that neither acid deposition nor natural processes have caused soil
r
acidification. However, if soils have acidified, measurements of fluxes are necessary to
determine the extent to which acid deposition has contributed to the observed rate of
acidification
There are no documented cases in which excessive atmospheric nitrogen deposition has
caused soil acidification, however, the potential exists if additions are high enough for a
sufficiently long time Nitrification is an acid-producing process (Alexander, 1963), and thus
the potential for soil acidification exists In practice, however, the levels of nitrogen input
necessary to produce measurable soil acidification are quite high For instance, Tamm and
Popovic (1974) report a drop in soil pH from approximately 5 to 4 5 after repeated nitrogen
fertilizations totaling 3,900 kg/ha over a period of 10 years Van Miegroet and Cole (1984)
report that 50 years of nitrogen fixation by red alder (Alnus rubra) caused the soil beneath
that stand to be 0 5 pH units lower (pH 4 6) than that in an adjacent Douglas-fir
(Psmdotiisga menziesii) stand (pH 5 2) Total nitrogen input rates were not known, but
typical rates for red alder range from 50 to 200 kg/ha/year (Van Miegroet and Cole, 1984)
Van Breemen et al (1982, 1987) report high acidification pressure on forests of the
Netherlands subject to very high inputs of nitrogen from nearby agricultural activities (often
considerably in excess of 50 kg nitrogen/ha/year, Van Breemen et al, 1982, 1987, Nilsson
and Grennfelt, 1988) The H+ budgets for these sites indicate the clear possibility (if not
probability) that soils have been acidified, but actual changes in soil acidity over tune have
not been measured
Soil acidification is usually thought of as an undesirable effect, but in some cases, the
benefits of alleviating nitrogen deficiency may outweigh the detriments of soil acidification
For instance, Van Miegroet and Cole (1984) found that excessive N2 fixation by red alder
caused large increases in NO3" leaching and a significant amount of soil acidification relative
to adjacent natural Douglas-fir stands, yet Douglas-fir growth is invariably superior on sites
10-70
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formerly occupied by red alder due to the differences in nitrogen status (Tarrant and Miller,
1963, Binldey, 1983, Van Miegroet et al , 1992)
10.5.5 Effects on Natural Waters
A major recent concern over the effects of soil acidification due to atmospheric
3 +
deposition of both nitrogen and S is the mobilization of Al , which can be toxic to some
terrestrial vegetation and might be earned to surface waters where it is toxic to fish As in
the case of soil acidification, a brief review of processes leading to soil solution and surface
water acidification will be presented as a prelude to discussions as to the effects of
atmospheric nitrogen deposition on these processes
r\
Increased concentrations of NO3" or any other mineral acid anion (e g , SO4 " or Cl") in
soil solution lead to increases in the concentrations of all cations in order to maintain charge
balance Figure 10-16 shows the effects of low (left) smd high (right) inputs of cations,
which are also accompanied by low and high inputs of amons, respectively, to the fictitious
soil with 10% base saturation shown on the right of Figure 10-15 As can readily be seen,
i o _i_
the concentrations of H and Al in soil solution are determined not only by base
saturation, but also by total cation (and anion) input rales Extremely acid soils are a
necessary but not sufficient condition for the mobilization of Al +, elevated inputs of cations
and amons, whether by atmospheric deposition, fertilization, or natural processes, must also
occur
The composition of the cations in a solution in equilibrium with soil can be described
fairly accurately by well-known selectivity equations developed more than 50 years ago
(Reuss, 1983) In essence, these equations predict thai the concentration of a given cation in
soil solution is governed by the proportion of this cation on the soil exchange complex and
the total ionic concentration in soil solution
Reuss (1983) points out one very interesting aspect of these equations with respect to
<> _i_
the question of Al mobilization As total ionic concentration increases, the concentration
Q .L O _i_
of Al increases to the 3/2 power of the increase in the concentrations of ratio Ca and
Mg2+ and to the third power of K+, Na+, and H+ In other words, as total cation and
anion concentrations increase, individual cation concentrations increase as follows
A13+ > Ca2+ and Mg2+ > K+, Na+, and H+ Thus, soil solution A13+ concentrations
10-71
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«3 I
increase not only as the soil acidifies (i e , as the proportion of Al on the exchange
complex increases), but also as the total ionic concentration of soil solution increases
(These equations also imply that K+, Na+, and H+ will be the least affected by increased
NO3" leaching.)
There are several studies in which Al + concentrations in both soil solution and stream
waters have been shown to be positively correlated with NO3" concentrations The NO3" -
34-
Al pulses in soil solution have implications for forest nutrition and are invoked in some
hypotheses of forest decline discussed in the next section Researchers on aquatic effects of
*3 -4- _J-
acid deposition have long noted springtime pulses of NO3", Al , and H in acid-affected
surface waters of the northeastern United States (Galloway et al, 1980, Dnscoll et al,
1989b). In less acid systems, NO3" pulses may be associated with base cations rather than
A13+ and H+ Foster et al (1989) noted pulses of NO3" and base cations in soil solutions
and streams at the Turkey Lakes site in Ontario Dnscoll et al (1989a) reviewed the North
American data relevant to the role of nitrogen in the acidification of surface waters and
explored relationships between atmospheric nitrogen deposition, soil carbon to nitrogen ratio,
and stream water nitrate concentrations They found no consistent relationships between
these factors, and suggested that vegetation uptake, as hypothesized by Vitousek and Reiners
(1975) may be one of the most important factors in determining stream water nitrate
concentrations.
10.5.6 Effects of Pollutant Nitrogen Deposition on Vegetation Nutrient
Status
Because nitrogen is the most commonly limiting nutrient for growth in forest
ecosystems in North America (Cole and Rapp, 1981), deposition of nitrogen in any
biologically available form to most forest ecosystems is likely to produce increased
vegetation growth to some extent Kauppi et al (1992) reported that, in stark contrast to
earlier claims of forest decline, the biomass of European forests increased over the 1971 to
1990 period. They attribute this growth increase to increases in nitrogen deposition and base
their conclusions on the magnitude of the increase in nitrogen deposition and all known
responses of European forests to nitrogen fertilizer It is logical to assume that the same
growth increase would occur in many forests in North America (especially western North
10-72
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America) with increased nitrogen deposition, given known nitrogen deficiencies and
responses to nitrogen fertilization (Aber et al, 1989, Gessel et al, 1973) The degree of
response will depend on the amount of nitrogen deposited, the nitrogen demand by
vegetation, and the competition from soil heterotrophic organisms for this nitrogen, as
described above In addition to changes in growth, increased nitrogen deposition can cause
significant changes in tree physiological function, can alter susceptibility to insect and disease
attack, and can even alter plant community structure (see Section 10 5 6 1) This section
briefly reviews plant physiological responses associated with increased nitrogen nutation (see
Section 10 6 for more in-depth coverage), gives a more in-depth review of soil-mediated
effects of nitrogen deposition on vegetation, and updates plant community/successional
changes that are reported to be occurring in high-deposition areas of Europe
10.5.6.1 Physiological Effects of Excess Nitrogen Inputs
Nitrogen addition can have several impacts on trees in addition to improvement of
growth, including susceptibility to other pollutants Nitrogen fertilization has been noted to
increase the resistance of eastern white pine (Pinus strobus L) to SO2 injury (Cotrufo and
Berry, 1970) Nitrogen fertilization usually depresses mycorrhizal development (Weetman
and Hill, 1973, Menge et al, 1977) Because the mycorrhizal association is thought to be an
adaption to nutrient deficient conditions, suppression of mycorrhizae by nitrogen inputs might
be expected
Several hypotheses posed to explain current forest declines in eastern North America
invoke the effects of excess nitrogen deposition on physiological processes These
physiological responses generally invoke altered carbohydrate allocation, causing increased
sensitivity to drought, frost, or insect attack Fnedland et al (1984) posed the hypothesis
that excessive nitrogen deposition induced growth latei into autumn, which caused
susceptibility to frost in red spruce in the northeastern United States Evans (1986) followed
up on this, observing that winter injury apparently occurred to first-year twigs and adding the
alternative hypothesis that excessive nitrogen deposition could have caused reduced bark
formation as well as, or instead of, late growth into the autumn in first-year twigs Waring
(1987) poses a hypothesis in which boreal coniferous species are unable to store nitrate taken
up from soil solutions, necessitating the formation of amino acids in green leaves, causing
10-73
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reduced allocation of carbohydrate to roots and increased susceptibility to drought and
pathogens.
More recent studies on response of red spruce to nitrogen lend no support to the
various hypotheses for nitrogen-induced physiological damage and decline descnbed above
Sheppard et al (1989) found the evidence for pollutant-induced susceptibility to freezing
injury in red spruce to be weak, based on laboratory studies with detached shoots DeHayes
et al. (1989) found that treatment of red spruce seedlings with NEySfC^ increased rather than
decreased cold tolerance Thus, the hypothesis that nitrogen causes direct damage to red
spruce is not supported by laboratory studies Climate is thought to play a major role in the
severe red spruce decline in the northeastern United States, perhaps with some additional
exacerbation due to the direct effects of acid mist on foliage (Johnson et al, 1992) There is
some evidence to suggest that indirect effects of nitrogen saturation, namely nitrate and
Al leaching, may be contributing factors to red spruce decline in the southern Appalachians,
and this literature is reviewed below
10.5.6.2 Soil-Mediated Effects on Vegetation
Nitrogen inputs in excess of tree and heterotrophic nitrogen demand may cause
immobilization of some nutrients (especially P and S) and losses of other cation nutrients due
to increased nitrate leaching, as discussed above In some cases, the benefits of enhanced
nitrogen status will greatly outweigh the detrimental effects of decreased availability of other
nutrients For instance, the benefits of nitrogen fixation during a red alder (Alnus rubra
Boug.) stage to subsequent Douglas-fir (Pseudotsuga menziesn [Mirb ] Franco) forests in the
Pacific Northwest are well documented despite the fact that excessive nitrogen fixation during
the red alder stage causes considerable phosphorus immobilization and soil acidification
(Van Miegroet and Cole, 1984) In other cases, effects of excessive nitrogen deposition may
be clearly deleterious to plant nutation For instance, Roelofs et al (1987) report that K and
Mg deficiencies in declining Dutch forests are caused by excessive foliar leaching due to
high inputs of NH4+.
Ulrich (1983) hypothesized that these nitrate-induced A13+ pulses during warm dry
years caused root damage and were a major contributor to what has been termed "forest
injury" observed in Germany during the mid 1980s This hypothesis is disputed by other
10-74
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German forest scientists who point out that "forest injury" occurred on base-rich as well as
base-poor soils (the base-nch soils are not subject to Al + pulses) (e g , Rehfuess, 1987)
Mulder et al (1987) document NO3" - A13+ pulses in soil solutions from forest sites in the
Netherlands Aluminum toxicity is one of several nitrogen-related hypotheses posed to
explain forest decline in that country (Other hypotheses are discussed in the following
section ) Johnson et al (1992) found pulses of NO3" and total Al in soil solutions during late
autumn from red spruce forests in the Great Smoky Mountains of North Carolina The
pulses were attributed to a combination of high rates of nitrogen mineralization and low
uptake in these over mature forests The soils at these sites were very rich in nitrogen,
(up to 10,000 kg nitrogen/ha) and atmospheric nitrogen deposition was also quite high (26 kg
nitrogen/ha/year), both of which contribute to the high rates of NO^" leaching at these sites
The peak total Al concentrations (70 jttM/L) associated with these NO3" pulses were below
the concentration for monomenc Al at which injury to red spruce seedlings occurs in
laboratory studies (200 jt*M/L, Joslin and Wolfe, 1988), and there was no visible evidence of
red spruce decline at these sites However, the possibility of Al inhibition of Ca and Mg
uptake cannot be excluded Spot checks revealed that 80 to 90% of total Al in these soil
solutions was in monomenc form It is noteworthy that Bondietti et al (1989) found an
inverse correlation between Al and Ca concentrations m tree rings of red spruce in the
southern Appalachians
Shortle and Smith (1988) present a hypothesis for the decline of red spruce in which
Al inhibits Ca uptake, Ca deficiency reduces cambial growth (because the demand for Ca per
unit of cambium surface is constant), reduced cambial growth causes a reduction in
functioning sapwood, and reduced sapwood causes a reduction in leaf area However,
Johnson (1983) finds no support for the Al hypothesis in the seriously declining trees of
Camel's Hump, VT He found that, although the degiee of dieback and decline increases
with elevation, both Al concentration and Al Ca ratios in fine roots decrease with elevation
He further points out that high elevation soils where much of the decline occurs are histosols
(organic soils) where Al toxicity is unlikely due to the mitigating effects of organics on soil
solution Al activity
Thus, the situation with respect to the Al hypothesis and red spruce decline remains
very unclear There is little support for the Al hypothesis in the northeast, where decline is
10-75
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very severe Cook and Johnson (1989) concluded from extensive tree ring and climatic
analyses that red spruce has been out of equilibrium with its climate for the last 150 years,
making it susceptible to injury from a variety of causes, both naturally and anthropogemcally
induced. Given the soil solution Al levels found in southern Appalachian red spruce forests,
the possibility of some Al effect cannot be excluded, yet decline in this region is much more
subtle (being evidenced primarily by somewhat controversial tree nng analyses) and no
unexpected levels of mortality have yet occurred Brnkley et al (1989) report that forests in
the South have responded most strongly to additions of nitrogen and phosphorus, probably
because growth of most stands in this area have been nitrogen- and phosporus-hmited
10.5.6.3 Ecosystem-Level Responses to Nitrogen Deposition
Growth responses to increased nitrogen inputs may not always be regarded as desirable,
especially if they result in changes in species composition For instance, unproved growth
and vitality due to increased nitrogen deposition may not be deemed desirable in wilderness
areas Different genera and species respond differentially to increased nitrogen availability,
for instance, deciduous species (angiosperms) generally have a greater demand for nitrogen
per unit biomass produced than do coniferous species (gymnosperms) (Cole and Rapp, 1981)
TUrnan (1987) found marked changes in species composition as a result of experimental
nitrogen additions to abandoned old fields in Minnesota Thus, there is a real possibility for
changes in ecosystem composition with increased nitrogen loading Changes from heathland
to grassland in Holland have been attributed to high rates of nitrogen deposition (Roelofs
et al, 1987). EUenberg (1987) points to further species changes in Central European
ecosystems as a likely consequence of elevated nitrogen He states that "more than 50% of
the plant species in Central Europe can only compete on stands that are deficient in nitrogen
supply."
There may be significant ecosystem-level effects of nitrogen via host-pathogen
interactions. Increased nitrogen inputs can affect tree resistance to insect and disease either
positively or negatively. Nitrogenous fertilizers are known to reduce the production of
phenols in plant tissues, thereby reducing resistance to infection by pathogenic fungi (Shigo,
1973). Hollis et al (1975) noted that additions of phosphorus and nitrogen to sites deficient
in these elements increased the incidence of fusiform rust in slash pine On the other hand,
10-76
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increased nitrogen input will increase resistance to bark beetle and other insect attacks if it
improves tree nutrient status (Weetman and Hill, 1973). In addition to changes in tree
physiology, increased nitrogen inputs cause changes in stand structure that result in changes
in understory composition and microclimate that could either increase or decrease the
likelihood of insect or disease attack Branstuig and Heil (1985), addressing the recent
changes from heather (Calluna sp ) to grasses in the Netherlands, noted that nitrogen
fertilization (56 kg nitrogen/ha) leads to increased growth of grasses only when Calluna
stands are opened up by beetle attacks By increasing the nitrogen concentration of heather
foliage, high nitrogen input stimulates larval growth and increases body weight of beetles
The effects of increased nitrogen additions on host-pathogen interactions remain largely
speculative Most research to date has been conducted in fertilized forest plantations
Insufficient research has been done on the responses of the either plantations or natural
ecosystems to pathogen attack under conditions of increased atmospheric nitrogen deposition
to make any definitive statements Nonetheless, these interactions are potentially very
important, given the devastation that pathogens can produce, and further attention should be
given to the issue of effects of increased nitrogen deposition, both positive and negative, on
host-pathogen interactions
10.5.7 Critical Loads for Atmospheric Nitrogen Deposition
Recently, there have been efforts to set critical loads for nitrogen deposition for natural
ecosystems (Nilsson and Grennfelt, 1988, Fox et al, 1989, Schulze et al, 1989) (also see
Section 10 4 3) In that the values for these critical loads may take on considerable political
importance, it is appropriate to examine the assumptions that have been made in defining
them
The Workshop held at Skokloster, Sweden in March 1988 (Nilsson and Grennfelt,
1988) adopted the following definition for a critical load "A quantitative estimate of an
exposure to one or more pollutants below which significant harmful effects on specified
sensitive elements of the environment do not occui according to present knowledge " In this
document (Nilsson and Grennfelt, 1988) and the subsequent publication synthesizing much of
it (Shulze et al, 1989), nitrogen critical loads were aimed "to protect soils from long-term
chemical changes with respect to base saturation" (Nilsson and Grennfelt, 1988, Schulze
10-77
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et al.j 1989) The critical loads for nitrogen are estimated from two equations The first
equation is posed as a one that must be satisfied in order to maintain a constant exchangeable
base cation pool in the soil
BC leaching < BC weathering + BC deposition — EC growth, (10-4)
where BC represents base cations Equation 10-4 is perhaps best understood by rearranging
BC leaching + BC growth < BC weathering + BC deposition * (10-5)
Equation 10-5 is simply a statement of mass balance for the soil cation exchange
complex and states that removal rates via leaching (BC leaching) and plant uptake (BC
growth) must be equalled or exceeded by inputs via deposition and weathering (the release of
base cations from unavailable, mineral forms to ionic states available for plant uptake,
leaching, or replenishing cation exchange sites) in order to keep soils from acidifying (keep
base saturation constant) This is followed by another equation describing the roles of NO3"
and SO4 in causing soil leaching
nitrate leaching + sulfate leaching < EC leaching (10-6)
Nilsson and Grennfelt (1988) state that Equation 10-6 assumes that all base cation
leaching is caused by nitrate and sulfate, ignoring the potentially substantial cation leaching
by naturally produced carbonic and organic acids (e g , Johnson et al , 1977) However, the
use of the "less than or equal to" sign in Equation 10-6 does, in fact, allow for leaching by
naturally produced carbonic and organic acids Base cation leaching will be less than nitrate
3+ +
plus sulfate leaching if Al and H are present to significant extent in soil solutions
Combining Equations 10-4 and 10-6, the authors obtain
acceptable nitrate leaching <, BC weathering + BC deposition (10-7)
- BC growth - sulfate leaching
10-78
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In obtaining Equation 10-7, the authors assumed (without stating so) that only the
"equal to" and not the "less than" sign in Equation 10-6 applied, in short, they assumed that
i -j i
all base cation leaching was due to nitrate and sulfate leaching, and that no H and Al
leaching occurred
To estimate nitrate leaching, the authors use the nitrogen balance equation
N input < N growth + N immobilization — N mineralization (10-8)
+ N demtnfication — N fixation + N leaching,
where N represents nitrogen Again, this equation is best understood by rearranging
AT leaching > (N input + N fixation + N mineralization) (10-9)
- (N growth + N immobilization + N demtnfication)
Equation 10-9 can be thought of as a mass balance equation for the soil inorganic
nitrogen pool, with the first three terms being inputs to that pool and the second three terms
being outputs from that pool other than leaching The inputs consist of atmospheric
deposition (N input), fixation (Nfixation), and release from soil organic matter during
decomposition (Nmineralization) The nonleaching outputs include plant uptake (N growth),
heterotrophic uptake (N immobilization), and demtrification (Ndemtnfication) The
remainder must be leaching (N leaching) It is assumed in their analysis that nitrogen
demtnfication and nitrogen fixation are negligible in forest ecosystems and that nitrogen
immobilization minus nitrogen mineralization, which is the net annual nitrogen accumulation
m the soil, equals only 1 to 3 kg nitrogen/ha/year The latter numbers are based on an
estimate of the net nitrogen accumulation in soils of Sweden since the last glaciation
(obtained by dividing nominal soil nitrogen content values by the number of years since
glaciation) Soil nitrogen accumulation rates can be much higher Jenkinson (1970)
documents net annual soil nitrogen accumulations of over 50 kg/ha/year over an 81-year
penod (from 1883 to 1964) after a former agricultural n>ite (Broadbalk) was allowed to revert
to forest at the Rothamsted Experiment Station in England This high rate of soil
nitrogen accumulation was greater than thought possible from atmospheric deposition alone
10-79
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and may have been in part due to the action of free-living nitrogen-fixers in the soil Liming
may have played some role in stimulating these high accumulation rates, a nearby site
(Geescroft) that had not been limed showed nitrogen accumulations of only 23 kg/ha/year
over the same period (Jenkinson, 1970)
Given these equations and estimates of the various parameters within them, the authors
calculate critical loads for various forest ecosystems These values range from a low of
3 to 5 kg nitrogen/ha/year for raised bogs to a high of 5 to 20 kg nitrogen/ha/year for
deciduous forests A critical concentration for nitrate in groundwater (10 mg mtrogen/L) is
then calculated based on an assumption of precipitation surplus (precipitation minus
evapotranspiration) of 100 to 400 mm/year, giving values of 10 to 40 kg nitrogen/ha/year
In contrast to the rather quantitative approach taken at the Skokloster Workshop, a far
more subjective approach is taken in determining critical nitrogen loads for wilderness areas
in the U S Forest Service-sponsored workshop held at Gary Arboretum in Millbrook, NY,
in May 1988. In this case, rather than attempting to come up with specific critical loads, the
workshop participants were asked to establish "green" and "red" lines, the former being
values below which deleterious effects are very unlikely to occur, and the latter being values
above which deleterious effects will very likely occur The "rationale used in selecting
nitrogen values" for terrestrial ecosystem critical loads consists of a bnef overview of the
nitrogen cycle and some educated guesswork, in view of the fact that "data on nitrogen
cycling in wilderness areas is quite scarce at best, and in many areas completely lacking "
Despite the lack of nitrogen cycling data, the authors provide guesses at green- and red-line
values for specific wilderness areas ranging from 3 to 10 kg nitrogen/ha/year for green
values and 10 to 15 kg/ha/year for red values These values were quantitatively similar to
those obtained in the Skokloster workshop, and actually show very little spread between
green and red hnes
10.5.8 An Evaluation of Critical Loads Calculations for Nitrogen
Deposition
There are a number of points that need to be emphasized before the Skokloster critical
load values are used for assessment or policy-making First, the assumption that soils can
accumulate only 1 to 3 kg nitrogen/ha/year is certainly not valid over the short term in most
10-80
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forest ecosystems, as shown amply by a number of forest fertilization studies described in
Section 10 5 3 Having stated that, however, it should also be noted that both heterotroph
and ecosystem-level recovery of atmospherically deposited nitrogen seems to be lower than
that of fertilizer nitrogen, as also noted in Section 10 5 3 The authors of the critical load
document (Nilsson and Grennfelt, 1988) recognize that nitrogen retention in the soil can be
quite high on a temporary basis, but they assume that only net increment in trees is
significant over the longer term (i e , harvest rotation lengths of 50 to 100 years)
Nonetheless, even "temporary" retention of atmospherically deposited nitrogen could be
significant If nitrogen-deficient systems can retain as much as 600 kg nitrogen/ha in the soil
by heterotrophs (see Table 10-13), an atmospheric nitrogen input of 25 kg/ha/year could be
retained for 24 years Recall that Jenkinson (1970) found an average annual nitrogen
accumulation of about 25 kg/ha/year in soils at the Rothamsted Experiment Station in
England over an 74-year period (1888-1962) This accumulation, which was calculated by
differences in measured soil nitrogen content over tune, is of special interest in that it
actually exceeded estimated atmospheric nitrogen deposition over that period It seems clear
that estimates of atmospheric nitrogen inputs to these sites are low, due either to
underestimates of dry deposition or nitrogen fixation
A critical unknown in soil heterotrophic nitrogen retention is the change (if any) in the
relative competitiveness of trees, heterotrophs, and mtnfiers, as noted earlier There is some
evidence to suggest that mtnfiers become more competitive with slow, steady inputs (Johnson
and Todd, 1988) Also, it is clear that tree nitrogen from the irrigation and fertilizer
experiments noted above (Aronsson and Elowson, 1980, Ingestad, 1981, Landsberg, 1986)
can increase substantially with increasing nitrogen deposition rate, bringing into question
calculations of nitrogen sequestering by trees from areas that are not nitrogen saturated
Also inherent in at least the final calculations is the assumption that no natural leaching
processes are currently contributing to soil acidification That is, all base cation leaching is
attributed to sulfate and nitrate This assumption is clearly false, carbonic and organic acids
are present in all soil systems and contribute to leaching and acidifying processes to varying
degrees (Johnson et al , 1977, Richter et al , 1983, Ulrich, 1980) The net result of this
assumption, ironically enough, is to underestimate soil acidification (i e , the acidification by
10-81
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carbonic and organic acids do not enter into the calculations) and, therefore, set critical loads
(as defined in these calculations) too low
The weakest link in this chain of calculations is, as always, base cation weathering
Although the chemical transformations of many weathering reactions are well known
(Lindsay, 1979), quantification of weathering rates under field conditions has remained
elusive The weathering numbers used in calculating these critical loads are crude mass
balance estimates based on amounts of minerals and cation nutrients left in soils 8,000 to
12,000 years after the last glaciation (when fresh minerals were first exposed) These
calculations do not account for changes in weathering rates with tune (rates were likely much
faster initially with fresh minerals than later during the course of weathering), nor do they
account for the possibility of increased weathering rates with increased acidification pressure
or with vegetation rooting (e g , Boyle and Voigt, 1973)
The entire critical loads concept that formed the basis of the Skokloster document is
based on preventing soil acidification Implicit in this goal is the assumption that soils reach
and remain in some kind of steady-state, nonacid condition in nature, an assumption that is
probably fallacious given the presence of extremely acid soils in pristine, unmanaged forests
(e.g., Johnson et al, 1977) Furthermore, it is not at all clear that soil acidification is
always harmful As shown in the red alder/Douglas fir succession example above, the
benefits of nitrogen deposition may well outweigh the detriments of soil acidification
It should be kept in mind that forests of the northern hemisphere have historically been
nitrogen deficient, and that growth increases brought about by fertilization (often at levels far
in excess of critical loads) have been regarded as beneficial, at least in commercial forest
lands. Value judgments inevitably come into play in setting critical loads for pollutant
deposition of nutrients, especially in the case of nitrogen
The green and red lines for nitrogen deposition established for wilderness areas in the
Gary Arboretum workshop (Fox et al, 1989) were almost totally subjective guesses and are,
therefore, open to many criticisms and arguments Given the fact that wilderness areas,
especially those in the western United States, are very likely nitrogen-limited, even the green
lines are not a guarantee of having no effect, as is acknowledged by the authors They state,
however, that "in our j'udgement, the Green Line levels are sufficiently low that perceptible
deleterious effects upon plant health, changes in species composition, or degradation of water
10-82
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quality are unlikely " In view of the very low nitrogen deposition rates in some parts of the
western United States, (1 to 2 kg/ha/year, Table 10-13), it seems likely that increases of up
to 10 kg/ha/year will result in some increases in plant growth and plant health, and, quite
possibly, changes in species composition The judgment that deleterious effects on plant
health and water quality are unlikely to occur at these levels seems to be a reasonable one for
the short term (i e , until biological nitrogen demand is satisfied in these slow-growing
ecosystems), but remain open to serious question over the long term
10.5.9 Conclusions
There is little doubt that, because of its role in plant growth, nitrogen deposition has
had an effect on many, if not most terrestrial ecosystems Because most forest ecosystems in
North America are nitrogen deficient, one of the most noticeable initial changes in response
to increased nitrogen deposition is likely to be a growth increase (Gessel et al , 1973, Aber
et al, 1989) Whether such a growth increase is deemed desirable or undesirable in a
particular ecosystem is entirely a matter of management objectives (timber production or
species preservation), and, ultimately, a value judgment by society
All current information indicates that "mtrogen-safurated" forests are relatively rare and
limited in extent (e g , Cole and Rapp, 1981), especially in managed forests Forest
management practices, especially with respect to harvesting and fire, will have a major effect
on the degree to which forests become nitrogen saturated The critical load values given in
the Skolster document (Nilsson and Grennfelt, 1988) are unlikely to produce nitrogen
saturation in highly productive, intensively managed foiests of the timberbelts in the
southeastern and northwestern United States that are frequently harvested and/or subjected to
control burning Indeed, there is considerable concern that intensive management practices
in these forests are causing nitrogen depletion (Boyle and Ek, 1972, Kimmuis, 1977, Smith
et al, 1986)
Because of the great variation in both natural forest nitrogen uptake rates and
management intensity, it is not reasonable to assign one critical load for all forest
ecosystems Intensively managed, short-rotation forests might beneficially utilize up to
100 kg nitrogen/ha/year, whereas a value as low as 10 kg nitrogen/ha/year may produce
undesired growth increases in very slow-growing virgin forests in wilderness areas
10-83
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In summary, it is clear that both the assumptions and the mechanics for calculations of
critical loads are seriously faulted Specifically, the assumption that soil acidification should
be the primary consideration for setting critical loads is not supported by a substantial body
of literature indicating that nitrogen status itself is most often the determinant of forest
ecosystem productivity Also, the assumption that a single or even a series of critical loads
can be set for forest ecosystems of widely varying ages and site conditions is certainly not
valid. Finally, calculations of critical loads fail to account for natural processes of soil
acidification, and implicitly assume that (1) nitrogen fixation is negligible and (2) soils are
naturally in a steady-state condition
10.6 TERRESTRIAL ECOSYSTEM EFFECTS-VEGETATION
Ecosystems respond to environmental stresses through their constituent organisms (see
Section 10.2) Plant populations, when exposed to any environmental stress, can exhibit four
different reactions (1) no response—the individuals are resistant to the stress, (2) severe
response—mortality of all individuals and local extinction of the extremely sensitive
population, (3) physiological accommodation—the growth and reproductive success of
individuals are unaffected because the stress is physiologically accommodated, and
(4) differential response—members of the population respond differentially, with some
individuals exhibiting better growth and reproductive success due to genetically determined
traits (Taylor and Pitelka, 1992, Garner, 1992) Differential response results in the
progressive elimination over several generations of sensitive individuals and a shift in the
genetic structure of the population toward greater resistance (microevolution) Physiological
accommodation and microevolution, with only the latter affecting biodiversity, are the most
likely responses for exposure to chronic stress (i e , stresses that are of intermediate-to-low
intensity and of prolonged duration) (Taylor and Pitelka, 1992) The primary effect of air
pollution on the more susceptible members of the plant community is that they can no longer
compete effectively for essential nutrients, water, light, and space, and are eliminated The
extent of change that may occur in a community depends on the condition and type of
community, as well as the pollutant exposure (Garner, 1992)
10-84
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Plant responses are foliar or soil mediated. Subsequent to the dry and wet deposition of
nitrogen forms from the atmosphere (Section 10 4), nitrogen-containing compounds can
impact the terrestrial ecosystems when they enter plant leaves and alter metabolic processes
(Chapter 9) or by modifying the nitrogen cycle and associated soil chemical properties
(Section 10 5) Changes in biochemistry that result in reduced vigor and growth and
decrease the plant's ability to compete for light, water, space, and nutrients can be
manifested as changes in plant populations, communities, and ultimately, ecosystems
(Chapter 9, Section 10 2) Interpretation of the effects of wet and dry deposited nitrogen
compounds at the ecosystem level is difficult because of the interconversion of nitrogen
compounds and the complex interactions that exist between biological, physicochemical, and
climatic factors (Sections 10 2 and 10 5, U S Environmental Protection Agency, 1982)
Nevertheless, reactive nitrogen compounds have been hypothesized to impact ecosystems
through modifications of individual plant physiological processes upon entering plants
through the foliage, or through alterations in the nitrogen status of the ecosystem
10.6.1 Foliage-Mediated Vegetation Effects
Reactive nitrogen compounds can have an impact on terrestrial ecosystems through
ambient air exposures by entering plants, usually through the leaves, and disturbing "normal"
physiological processes However, in the United States, concentrations seldom reach
phytotoxic levels (Chapters 7 and 9) Because information on the direct effects of NO and
NO2 alone and in combination with other pollutants have been described in detail in
Sections 9 3 through 9 6, they will not be discussed here
Very little information is available on the direct effects of HNO3 vapor on vegetation,
and essentially no information is available on its effects on ecosystems Norby et al (1989)
reported that HNO3 vapor (0 075 ppmv) induced nitrate reductase activity (NRA) in red
spruce foliage Because the induction of NRA is a step in the process leading to the
formation of organic nitrogen compounds (amino acids), the nitrate from HNO3 could
function as an alternative source of nitrogen for plant growth However, in plants under
stress, the reduction of nitrate to ammo acids consumes energy needed for alternative
metabolic processes, a potentially slight negative impact
10-85
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The effects of NH3, a reduced nitrogen gas, have been summarized by Van der Eerden
(1982). However, NH3 concentrations seldom reach phytotoxic levels in the United States,
consequently it will not be extensively discussed here (U S Environmental Protection
Agency, 1982) In contrast, high NH3 concentrations in Europe have been observed
(Van Dijk and Roelofs, 1988) Van der Eerden (1982) summarized available information on
the response of crop and tree species to NH3 fumigation and concluded that the following
concentrations produced no adverse effects 0 107 ppmv (75 /ig/m ) yearly average,
••3 o
0.858 ppmv (600 jt*g/m ) daily average, and 14 3 ppmv (10,000 /*g/m ) hourly average
Submicron, ammonium sulfate aerosols have been shown to affect foliage of Phaseolus
•3
vulgans L (Gmur et al, 1983) At a concentration of 26 mg/m (37 ppmv), 3 weeks of
exposures produced leaf chlorosis, necrosis, and loss of turgor Gmur et al (1983) reported
that these foliar symptoms were not correlated with changes in shoot or root dry mass, and
suggested that no relationship to plant growth was expected However, the 3-week
experiment was not long enough for significant changes in dry matter to be observed The
level of NH3 producing the leaf effects (37 ppmv) exceeds normal ambient levels for the
United States, but it is representative of reported high concentration episodes in Europe
(Gmur et al, 1983) Cowling and Lockyer (1981) reported beneficial effects of NH3 on the
growth of Lolium perenne L due to sorption of NH3 nitrogen through leaves Van Hove
et al. (1989b) studied the effects of 50 and 100 ^g/m NH3 on Populus euramencana L
over a 6- to 8-week period and found increases in photosynthesis at 100 jwg/m3, but no
changes in stomatal characteristics up to that level of NH3
10.6.2 Son-Mediated Vegetation Effects
Effects of dry nitrogen deposition to terrestrial ecosystems result from the addition of
nitrogen to ecosystem soils at a rate above that expenenced during normal successional
processes. (The effects of nitrogen deposition on soils has been discussed in Section 10 5 )
Growth responses to added nitrogen would be anticipated in many cases because many
natural systems are nitrogen limited (Krause, 1988, National Research Council, 1979, see
also Sections 10 5 and 10 7) However, if atmospheric additions of nitrogen exceed the
"buffering" capacity of an ecosystem, alterations in soil chemistry are expected to take place
(Section 10.5) Inputs of nitrogen to natural ecosystems alleviate deficiencies and allow
10-86
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increased growth of some plants, but in doing so, also can impact interplant competitive
relationships and alter species composition and diversity in sensitive ecosystems
(U S Environmental Protection Agency, 1982, Ellenberg, 1987, Kenk and Fischer, 1988)
Schulze (1989) also has proposed that excessive additions of nitrogen lead to nutrient
deficiencies of other elements (Ca, Mg) Symptoms of Mg deficiency and drought are
frequently associated with large amounts of soil nitrate Aber et al (1989) stated that when
nitrogen becomes readily available, some other resources (e g , P for plants or C for
microorganisms) become limiting
In addition to the potential for increasing plant productivity through fertilization, the
deposition of nitrogen from the atmosphere to ecosystems has been hypothesized to alter
normal nutrient cycles and physiological processes, resulting in increased susceptibility of
forests to other environmental stresses (Sections 10 5 and 10 6, Lindberg et al, 1987,
Nihlgard, 1985, McLaughlin, 1985, Schulze, 1989) Physiological unbalances resulting from
excessive nitrogen additions are also hypothesized to disrupt the winter hardening process
(Nihlgard, 1985, Fnedland et al , 1984, Waring,, 1987), produce nutrient unbalances
(Nihlgard, 1985, Waring, 1987, Schulze, 1989), and alter carbon allocation patterns within
plants (Nihlgard, 1985, McLaughlin, 1985) Changes in nitrogen supply can have an impact
on an ecosystem's nutrient balance and, as discussed in the previous section, alter many plant
and soil processes involved in nitrogen cycling (Aber et al, 1989) Among the processes
affected are (1) plant uptake and allocation, (2) litter production, (3) immobilization (includes
ammomfication [the release of ammonium] and nitrification [the conversion of ammonium to
nitrate during the decay of litter and soil organic matter]), (4) NO3" leaching, and (5) trace
gas emissions (Aber et al, 1989 [Figure 10-17]) Aber et al (1989) have developed an
integrated set of hypotheses that portray the progression of changes in major plant and soil
processes in northern forest ecosystems in response to chronic nitrogen deposition and
conclude that these ecosystems have a limited capacity to accumulate nitrogen Nitrogen
fixation is usually inhibited at high levels of available nitrogen (Waring and Schlesinger,
1985)
An increase in the nitrogen litter content and m litter decomposition rates and an
alteration in nitrogen cycling have been observed in the more highly polluted areas when
compared with moderate- and low-polluted areas of the San Bernardino Mountains of
10-87
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Process altered by
nitrogen deposition
Deposition
Photosynthesis
X
Animal
Proteins
Soil
\/
I 4
Bacterial
Nitrogen
Fixation
* K
Litter
Production
(Death)
DeathX
A
Microbial
Decomposition
\
\
Trace
Gas
Emissions
V
V
Figure 10-17. Nitrogen cycle (dotted lines indicate processes altered by chronic nitrogen
deposition).
Source Garner (1992)
Southern California (Fenn and Dunn, 1989) A pollutant concentration gradient exists with
24-h O3 concentrations at the high sites in the west averaging 0 1 ppm or higher, moderate
sites ranging from 0 06 to 0 08 ppm, and low sites in the east averaging 0 05 ppm or less
(Fenn, 1991) Nitrogen and sulfur compounds also occur ui the pollutant mixture to which
the mountains downwind of the Los Angeles Basin are exposed (See 10 2, Bytnerowicz
et al., 1987a,b, Solomon et al, 1992) A nitrogen deposition gradient from west to east
parallels the decreasing O3 gradient Deposition of nitrogen exceeds that of sulfur (Fenn and
Bytnerowicz, 1992) Annual average HNO3 concentrations in 1986 ranged from 1 2 ppb
near the Southern California coast to 2 7 ppb in the San Gabriel Mountains (Solomon et al,
1992).
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The effects of O3 exposure and injury to ponderosa (Pinus ponderosa Laws ) and
Jeffrey pine (P jeffreyi Grev & Balf) on a mixed conifer forest in the San Bernardino
Mountains, east of Los Angeles, have been studied for many years (Miller, 1973, Miller,
1984, U S Environmental Protection Agency, 1986) The litter layers under trees severely
injured by O3 is deeper than that under trees less severely injured (Fenn and Dunn, 1989)
A comparison study of litter decomposition rates of L-layer litter indicates that litter from the
more polluted areas in the west decomposed at a significantly (p = 0 01) faster rate than
litter from moderate to low pollution levels (Fenn and Dunn, 1989, Fenn, 1991) Nitrogen
content of litter was greatest at the high pollution sites and was positively correlated with the
litter decomposition rate The higher nitrogen and lower Ca content of the litter suggests
that litter in the western plots originated from younger needles than at the less polluted sites,
possibly due to O3-induced needle abscission Fungal diversity was also greater in the litter
from the western San Bernardino Mountains (Fenn and Dunn, 1989)
When the factors associated with enhanced litter decomposition were investigated, it
was found that nitrogen concentrations of soil, foliage, and litter of ponderosa and Jeffrey
pine were greater in the plots where pollution concentrations were high than in moderate- or
low-pollution sites This was also true for sugar pine (Pinus lambertiana Dougl) and for
incense cedar (Calocedrus decurrens [Torr ] Florin ), two O3-tolerant species The rate of
litter decomposition for all three pme species was greater at the high-pollution sites
Therefore, the increased rate of litter decomposition in the high-pollution plots does not
appear to be related to O3 sensitivity or premature needle abscission, but rather due to higher
levels of nitrogen in the soils (Fenn, 1991)
Nitrogen is the mineral nutrient that most frequently limits growth in both agricultural
and natural systems (Chapin et al , 1987) The uptake of nitrogen and its allocation is of
overriding importance in plant metabolism and governs, to a large extent, the utilization of
phosphorus, potassium, and other nutrients, and plant growth As indicated earlier
(Section 10 1), plants usually obtain nitrogen by absorbing ammonium (or ammonia) or
nitrate (or nitrite) through their roots or through fixation by symbiotic organisms Nitrogen
availability via the nitrogen cycle typically controls net primary productivity Normally, the
acquisition of nitrogen is a major carbon expense for plants Plants expend a predominant
fraction of the total energy available to them in the form of carbohydrates in the acquisition
10-89
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of nitrogen through the processes of (1) absorption, bringing nitrogen into the plant from the
environment, (2) translocation, moving inorganic nitrogen within the plant, and
(3) assimilation, conversion of inorganic to organic nitrogen (Chapin et al, 1987)
Absorption of nitrogen from the soil requires constant and extensive root growth to meet the
needs of a rapidly growing plant because the soil pools of mineral nitrogen, ammonium, or
nitrate in the immediate vicinity of the roots are usually so small that they are quickly
depleted (Section 10 5) A crude estimate suggests that the fraction of carbon budget spent
on absorption, translocation, and assimilation ranges from 25 to 45 % for ammonium, 20 to
50% for nitrate, 40 to 45% nitrogen fixation, and 25 to 50% for formation of mycorrhizae
(Chapin et al., 1987)
Nitrogen uptake influences photosynthesis because in the leaves of plants with
C3 photosynthesis (the pathway used by most of the world's plants), approximately 75% of
the total nitrogen is contained in the choloroplasts and is used during photosynthesis The
nitrogen-photosynthesis relationship is, therefore, critical to the growth of trees (Chapin
et al , 1987). As a rule, plants allocate resources most efficiently when growth is equally
limited by all resources When a specific resource such as nitrogen limits growth, plants
adjust by allocating carbohydrates to the organs that acquire the most strongly limiting
resources (Figure 10-18)
Among boreal and subalpine conifers, nitrogen added to the soil may not increase
growth Depending on the plant species, nitrogen use efficiency above a critical level
decreases. All plants do not necessarily benefit from the added nitrogen in the leaves More
nitrogen in the soil is not mirrored directly by increased nitrogen uptake except at low levels
(Section 105) This is particularly true of conifers that are adapted to low-resource
environments and tend to have low potential growth rates The photosynthetic capacity of
conifer foliage is low and not greatly enhanced by increased nitrogen content (Waring, 1985,
Chapin, 1991) High leaf nitrogen content is not always an advantage when other resources,
among which are light and water, are limited
Nitrate reductase is the enzyme that catalyses the reduction of nitrate to nitrite
Its levels of activity are determined by the supply of nitrate (Section 932) The nitrate
reductase enzyme activity in roots and shoots determines the pattern of nitrate assimilation
Increases in root nitrate supply are associated with large increases in the shoot Nitrogen
10-90
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Leaf
Biomass
X
Initial
Allocation
State
Environmental
Stress
Photosynthetic / Root x NutnentN
Rate = \ Biomass Uptake^
A
Carbon
Reduce Carbon Supply
Low Light
S02/03
Leaf
Biomass
New
Allocation
State
Carboi
Reduce Nitrogen/Water Supply
Drought
Low Soil Fertility
Root
Biomass
Carbon
Figure 10-18. Impact of a reduced supply of carbon to the shoot, or water and nitrogen
to the roots, on subsequent allocation of carbon.
Source Winner and Atkinson (1986)
source and environmental conditions such as light, temperature, pH, CO2 and molecular
oxygen (O2) tensions, and water potential, factors that regulate nitrate reductase activity,
exert a regulatory effect on the supply of reduced nitrogen to the plant (Haynes, 1986)
Studies indicate that the single most important nitrogenous component limiting
photosynthetic capacity is nbulose-l,5-biphosphate carboxylase-oxygenase (RUBISCO), the
primary CO2-fixing enzyme in C3 and the ultimate CO2-fixing enzyme in plants with C4 and
CAM photosynthetic pathways (Chapin et al , 1987) In individual leaves, nitrogen
availability during growth controls the RUBISCO level The importance of photosynthesis
limitation by RUBISCO vanes with light and CO2 availability and with the partitioning of
nitrogen among potentially limiting factors Sun plants invest more nitrogen in RUBISCO
than shade plants, in low light, increased RUBISCO is not beneficial When photosynthesis
is measured at light saturation, leaf nitrogen is closely correlated with photosynthetic
capacity But when light is low, photosynthesis increases very little, if at all, with increasing
leaf nitrogen (Chapin et al , 1987) In dense conifer forests, lack of sunlight makes the
metabolic conversions of nitrate inefficient because photosynthesis (i e , the production of
10-91
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large amounts of carbohydrates) and other light-driven reactions become limiting (Zeevaart,
1976). Altered carbohydrate allocation results
Patterns of carbohydrate allocation directly influence growth rate Excess nitrate alters
carbohydrate allocation between shoots and roots (Figure 10-18) It shifts carbohydrate
allocation to the shoots, increases production of foliage, and provides nitrogen in a form
difficult for the plant to metabolize (Waring, 1987) The capacity of gymnosperms in
general, and subalpine and boreal species in particular, to synthesize the enzymes required to
reduce the increased nitrate in foliage or roots appears to be limited Reduced allocation of
carbohydrates to the roots, on the other hand, is associated with the accumulation of ammo
acids in foliage (Waring, 1987) Conifers are plants characteristic of resource-poor
environments and tend to have low potential growth rates When nitrogen is no longer
limiting, deficiencies of other nutrients may occur (Aber et al, 1989, Kenk and Fischer,
1988). Competition under the above circumstances favors deciduous tree species, and other
plants characteristic of resource-rich environments, rather than conifers (Waring, 1987)
Altered shoot root ratios resulting from different patterns of carbon allocation can lead
to increased susceptibility to drought because shoots grow at the expense of roots under high
nitrogen availability (Freer-Smith, 1988, Norby et al , 1989, McLaughlin, 1985, Waring,
1987). Changes in carbon nitrogen ratios of tissues resulting from an excessive supply of
nitrogen can also result in altered host-pathogen, mycorrhizal, and pest-plant interactions
(Chapin et al, 1987, Grennfelt and Hultberg, 1986, Nihlgard, 1985)
Although much has been hypothesized about the impact of excessive inputs of nitrogen
into forest ecosystems, direct experimental information to prove or disprove these hypotheses
is not widely available Margolis and Waring (1986) showed that fertilization of Douglas fir
with nitrogen could lengthen the growing season to the point where frost damage became a
problem However, Klein and Perkins (1987) presented other evidence that showed no
additional winter injury of high elevation conifer forests when fertilized with 40 kg total
nitrogen/ha/year On the other hand, De Temmerman et al (1988) provided data showing
increased fungal outbreaks and frost damage on several pines species exposed to very high
NH3 deposition rates (>350 kg/ha/year) Numbers of species and fruiting bodies of fungi
have also increased concomitantly with nitrogen deposition in Dutch forests (Van Breemen
and Van Dijk, 1988) An increase in total amino acid concentrations in needles known to
10-92
-------
take place in response to dry deposition of NOX (Section 10 4) has also been suggested to
favor outbreaks of insect pests (Waring and Pitman, 1985, White, 1984) Schulze (1989)
presents a clear progression of evidence that indicates that canopy uptake of nitrogen together
with root uptake has caused a nitrogen imbalance in Norway spruce leading to its decline
Van Dijk et al (1990) conducted a greenhouse study to determine the impact of ammonium
in rainwater on three coniferous trees (Douglas fir, Corsican pine, and Scots pine) and found
no sign of deterioration in seedlings receiving nitrogen at the rate of 48 kg/ha/year At the
very high rates of application of 480 kg nitrogen/ha/year, increases in shoot root ratio and
reductions in fine root and mycorrhrzal biomass were observed However, this level of
nitrogen addition (i e , simulated deposition) is approximately one order of magnitude greater
than most rates of deposition in North America or Europe Kenk and Fischer (1988)
summarized fertilization experiments on German forests and found little evidence for growth-
lumting effects, but since 1960, some indication of increased growth that could be the result
of atmospheric nitrogen deposition was indicated for Norway spruce. Further, they point out
that atmospheric deposition has eliminated or diminished the former widespread nitrogen
deficiencies Miller and Miller (1988) concluded that fertilizer trials are not appropriate for
extrapolation as indicators of forest response to nitrogen deposition (i e , the timing of
applications is typically quite different), but nevertheless they also suggested that results of
such trials ought to be reconcilable with the "natural" phenomenon
In addition to these indirect soil-mediated effects on individual plants, Ellenberg (1987)
has suggested that current balances of interspecific competition in some sensitive ecosystems
can be altered by additional sources of nitrogen and result in the displacement of existing
species by plants that can utilize the excess nitrogen more efficiently (see Section 10 5 4)
Because the competitive equilibrium of plants in any community is finely balanced, the
alteration of any one of a number of parameters (e g , increases in nitrogen) can alter
ecosystem structure and function (Skeffington and Wilson, 1988) For example, Roelofs
et al (1987) proposed that NH3/ammonium deposition leads to heathland changes via two
modes (1) acidification of the soil and associated loss of cations such as K+, Ca2+, and
2_i_
Mg , and (2) nitrogen enrichment, which results in "abnormal" plant growth rates and
altered competitive relationships
10-93
-------
Excessive nitrogen inputs to terrestrial ecosystems can cause differential competitive
advantage among plants within a heathland (Heil and Bruggink, 1987, Heil et al , 1988)
(see also Section 10 7 4 4) The authors established that the changing nature of unmanaged
heathlands in the Netherlands, where Calluna vulgans (L) Hall is being replaced by grass
species, is a result of the eutrophic effect of acidic rainfall and large nitrogen inputs arising
from intensive farming practices in the region Both Calluna vulgans (L) Hall and Molmia
caendea (L) Moench are stress-tolerant species (Grime, 1979), but they have different
growth patterns Calluna is an evergreen, but its long growing season can normally
compensate for its slow growth rate, so that it competes successfully with the faster growing
Molmia under normal nutrient-limiting conditions A large increase in the nitrogen supply,
however, improves the competitive advantage of Molmia, increasing its growth rate so that it
becomes the dominant species in the heathland
In support of hypotheses that nitrogen deposition is altering interspecific competition,
Roelofs et al (1987) have observed that nitrophilous grasses (Molmia and Deschampsid) are
displacing slower growing plants (Erica and Calluna) on heathlands in the Netherlands, and
the authors suggest that a clear correlation exists between this change and nitrogen loading
Statistical data for the correlation was not provided These changes in the Netherlands have
taken place under nitrogen loadings of between 20 and 60 kg nitrogen/ha/year Liljelund and
Torstensson (1988) have shown clear signs of vegetation changes in response to nitrogen
deposition rates of 20 kg/ha/year Van Breemen and Van Dijk (1988) summarized data for
heathlands showing a substantial displacement of heathland plants by grasses from 1980 to
1986. They summarize data showing increases in the presence of nitrophilous plants m the
herb layers of forests It was observed also that the fruiting bodies of mycorrhizal fungi
have decreased in number Ellenberg (1988) has also suggested that long before toxic effects
appear on individual plants, ionic inputs (NO3~ and NH4+) have influenced competition
between organisms
10.6.2.1 Foliage and Soil-Mediated Effects—Combined Stress
The environment is seldom optimal in either natural or agricultural communities It is
not unusual, therefore, for plants growing in natural habitats to encounter multiple stresses
Plant responses to multiple stresses depend on resource (carbon and nitrogen) interactions at
10-94
-------
levels ranging from the cell to the ecosystem (Chapin et al., 1987) At the present tune, data
dealing with the response of trees or other vegetation to the combined stresses of
O3 exposure above ground and nitrate deposition through the soil are sparse Tjoelker and
Luxnioore (1991), however, have assessed the effects of soil nitrogen availability and chronic
O3 stress on carbon and nutrient economy in 1-year-old seedlings of loblolly pine (Pinus
taeda L) and yellow poplar (Linodendron tulipifera L) Elevated O3 concentrations altered
biomass partitioning to needles of the current year Ozone concentrations of 0 108 ppm
reduced the biomass of current-year needles in loblolly pine seedlings grown at the highest
(172 jwg/g) nitrogen supply by 20%, but not those grown with a low (59 jitg/g) supply of
nitrogen The interaction between O3 and nitrogen suggests that plants grown with a high
nitrogen supply are more sensitive to chronic O3 stress in terms of biomass reduction
(Tjoelker and Luxnioore, 1991) Similar results in the growth of domestic radish (Raphanus
sativa L , cv Cherry Bell) were obtained by Pell et al (1990) Brewer et al (1961) and
Harkov and Brennan (1980) observed increased foliar injury when plants were grown with an
adequate nitrogen supply
10.6.3 Nitrogen Saturation, Critical Loads, and Current Deposition
Ecosystem nitrogen saturation and the definition of the critical levels of total
nitrogen deposition at which changes or negative impacts begin to appear in ecosystems have
been the subject of several recent conferences in Europe (Nilsson and Grennfelt, 1988,
Brown et al, 1988, Skeffington and Wilson, 1988) Miller and Miller (1988) proposed three
definitions for nitrogen-saturated ecosystems (1) no response to additional nitrogen,
(2) growth reductions in response to added nitrogen, and (3) added nitrogen leads to
increased losses of nitrate in stream water, and concluded that the third was the most
reasonable (see also Section 10 3) Brown et al (1988) reported that a recent workshop
concluded that nitrogen saturation could be best defined as occurring when nitrogen outputs
from ecosystems exceeded inputs This conclusion was based on a model of plant/soil
nitrogen saturation put forth by Agren and Bosatta (1988) Aber et al (1989) similarly
define nitrogen saturation as the availability of ammonium and nitrate in excess of total
combined plant and microbial nutritional demands The concept of nitrogen saturation leads
to the possibility of defining a critical nitrogen load (deposition rate) at which no change or
10-95
-------
deleterious impacts will occur to an ecosystem (Nilsson and Grennfelt, 1988) It is important
to recognize that the magnitude of such a "critical load" will be site- and species-specific,
being highly dependent on initial soil chemistries and biological growth potentials (i e ,
nitrogen demands)
10.6.3.1 Critical Nitrogen Loads That Have Been Proposed
Skefflngton and Wilson (1988) summarized and discussed the following possible criteria
as potentially useful for defining appropriate critical nitrogen loads on ecosystems
• prevent nitrate levels in drinking or surface waters from rising above
standard levels,
• ensure proton production is less than weathering rate,
• maintenance of a fixed NH3-base cation balance,
• maintain nitrogen inputs below nitrogen outputs (the nitrogen saturation approach),
and
* minimize accelerations in the rates of ecological succession (vegetation changes due
to altered interspecific competition)
De Vries (1988) has also defined criteria for a combined critical load for nitrogen and
S for Dutch forest ecosystems based on the following nitrogen contents of foliage, nitrate
concentrations in groundwater, NH4/K ratios, Ca/Al ratios, and Al concentrations in soil
solution Based on these criteria, De Vries concluded that current rates of nitrogen and
S deposition in the Netherlands exceed acceptable levels
Schulze et al. (1989) have also proposed critical loads for nitrogen deposition based on
an ecosystem total amon and cation balance This approach makes the assumption that
processes determining ecosystem stability are related to soil acidification and nitrate leaching
(see also Section 10 5.6) They concluded that in order to limit the mobilization of
aluminum and other heavy metals resulting from acidification and nitrate leaching (a negative
result), critical nitrogen deposition rates could not exceed 3 to 14 kg nitrogen/ha/year for
silicate soils or 3 to 48 kg nitrogen/ha/year for calcareous-based soils Other cntical loads
have been proposed at rates of nitrogen deposition ranging from as little as 1 kg to levels
near 100 kg nitrogen/ha/year, depending on the impacts considered acceptable and the
criteria used to define the cntical load
Critical loads less than 20 kg/ha/year have been proposed based on criteria to minimize
species changes (Van Breeman and Van Dijk, 1988, Liljelund and Torstensson, 1988)
10-96
-------
Vegetational changes from heathland to grassland occurred in the Netherlands when nitrogen
deposition was greater than 20 kg/ha/year Changes in the beech and oak woodlands in two
areas of southern Sweden were observed when nitrogen deposition ranged from 20 to
30 kg/ha/year (Liljelund and Torstensson, 1988) Changes in the species composition of
softwater pools were noted when NH4+ deposition was in the 10- to 20-kg nitrogen/ha/year
range Nitrogen deposition would have to decrease to less than 6 kg/ha/year to return both
terrestrial and aquatic vegetation to the flora that was abundant decades ago (Van Breeman
and Van Dijk, 1988) Liljelund and Torstensson (1988) point out that establishing critical
loads depends on the criteria used One critical load would be required to prevent species
change, whereas another would be required to prevent community change Using the cntena
that ecosystem nitrogen inputs should not exceed outputs, critical loads have been proposed
as low as 1 to 5 kg nitrogen/ha/year for poorly productive sites with low productivity or in
the range from 5 to 30 kg nitrogen/ha/year for sites having medium quality soils and for
common forested systems (Boxman et al, 1988, Rosen, 1988, Skeffington and Wilson, 1988,
World Health Organization, 1987)
In their summary of a recent conference on critical nitrogen loading, after discussing
various options for setting a critical nitrogen load, Skeffington and Wilson (1988) concluded
that "we do not understand ecosystems well enough to set a critical load for nitrogen
deposition in a completely objective fashion " Brown et al (1988) further concluded that
there was probably no universal critical load definition lhat could be applied to all
ecosystems, and a combination of scientific, political, and economic considerations would be
required for the application of the critical load concept
The following terrestrial ecosystems have been suggested as being at risk from the
deposition of nitrogen-based compounds
• heathlands with a high proportion of lichen cover,
• low meadow vegetation types used for extensive grazing and haymaking, and
• coniferous forests, especially those at high altitudes (World Health
Organization, 1987, Aber et al, 1989)
These oligotrophic ecosystems are considered at risk from atmospheric nitrogen inputs
because plant species having high potential growth rates, but normally restricted by low
nutrient concentrations, can gain a competitive advantage, and their growth at the expense of
existing species changes the "normal" species composition and displaces some species
10-97
-------
entirely (Ellenberg, 1987, Waring, 1987) Sensitive natural ecosystems, unlike highly
manipulated agricultural systems, may be prone to damage from exposure to dry-deposited
nitrogen compounds because processes of natural selection whereby tolerant individuals
survive may not be keeping pace with the current levels of atmospheric nitrogen deposition
(World Health Organization, 1987, Waring, 1987)
10.6.3.2 Current Rates of Total Nitrogen Deposition
Application of the concept of critical nitrogen loading has not yet been widely adopted
in North America (based on the very limited published data), but a comparison of total
nitrogen deposition data for North America and proposed critical loads just discussed should
provide a reasonable comparison of the status of terrestrial systems with respect to changes
expected from elevated levels of nitrogen deposition Tables 10-14 and 10-15 summarize
information regarding the total deposition of nitrogen to a variety of ecosystems/forest types
in North America Table 10-14 summarizes information regarding the total deposition of
nitrogen to a variety of ecosystems/forest types or regional areas in North America and
Europe.
Nitrogen deposition can be divided into four categories, depending on its origin cloud
water, precipitation, dry particles, and gaseous forms Figure 10-19 summarizes wet
deposited nitrate and ammonium deposition data for various states that were part of the
National Acid Deposition Program (NADP) Table 10-15 specifically addresses the issue of
relationships between ecosystems' nitrogen inputs and outputs Data in these tables indicate
that total deposition of nitrogen in North America, particularly the eastern United States, is
comparable to that found for many areas in Europe North American sites would appear to
have total nitrogen deposition rates less than 25 kg nitrogen/ha/year It is also obvious from
these summary tables that much of our information on nitrogen deposition is limited to
information on nitrate and ammonium deposition in rainfall Lindberg et al (1987)
concluded that the lack of data on multiple forms of nitrogen deposition limits our ability to
accurately determine current levels of nitrogen loading
Olsen (1989) summarized nitrate and ammonium concentration and wet deposition data
for the United States and southern Canada for the period from 1979 through 1986 For
1986, the greatest annual rates of ammonium and nitrate deposition were localized in the
10-98
-------
TABLE 10-14. MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL NITROGEN DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Forms of Nitrogen Deposition (kg/ha)a
Site Location/
Vegetation
United States
California, Chaparral
California, Sierra Nevada
Georgia, Loblolly pine
North Carolina, Loblolly pine
North Carolina, Hardwoods
North Carolina, White pine
North Carolina, Red spruce
New Hampshire, Deciduous
New Hampshire, Deciduous
New York, Red spruce
New York, Mixed deciduous
Tennessee, Mixed deciduous
Tennessee, Oak forest #1
Tennessee, Oak forest #2
Tennessee, Oak forest #1
Tennessee, Oak forest #2
Tennessee, Oak forest
Tennessee, Loblolly pine
Washington, Douglas fir
Washington, Douglas fir
U S Regions
Adirondacks
Midwest
Northeast
Northwest
Southeast
Southeast Appalachians
Wet
Cloud Rain
82
—
37
87
48
37
87 62
70
93
73 61
48
29
32
29
69
60
45
43
29
1 0
63
42
21 7
166
206
42
Dry
Particles Gases
-
—
10 42
22 41
05
09 27
36 86
—
-
02 23
08 25
41 61
44 40
44 40
13
12
18 38
06 14
13 06
" "
47
29
—
—
—
3 1
Totalb
23C
(2)
9
15
53
7
27
(7)
(9)
16
8
13
12
11
8
7
10
9
5
(1)
11
7 1
22
17
21
73
Reference
Rigganetal (1985)
Williams and Melack
(1991a)
Lovett (1992)
Lovett (1992)
Swank and Waide (1988)
Lovett (1992)
Lovett (1992)
Likens et al (1970)
Likens (1985)
Lovett (1992)
Lovett (1992)
Kelly and Meagher
(1986)
Kelly and Meagher
(1986)
Kelly and Meagher
(1986)
Kelly (1988)
Kelly (1988)
Lindbergetal (1986)
Lovett (1992)
Lovett (1992)
Henderson and Hams
(1975)
Dnscolletal (1989a)
Dnscolletal (1989a)
Munger and Eisenreich
(1983)
Munger and Eisenreich
(1983)
Munger and Eisenreich
(1983)
Dnscolletal (1989a)
10-99
-------
TABLE 10-14 (cont'd). MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL NITROGEN DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Site Location/
Vegetation
Forms of Nitrogen Deposition (kg/ha)
Wet
Dry
Cloud
Rain
Particles Gases Total Reference
Canada
Alberta (southern)
British Columbia
Ontario
Ontario (southern)
73
55
37
23
122
14
19 5 Peake and Davidson
(1990)
(5) Feller (1987)
(4) Linseyetal (1987)
37 Roetal (1988)
Federal Republic of Germany
Spruce (Southeast slope)
Spruce (Southwest slope)
Netherlands
Oak-birch
Deciduous/spruce
Scots pine
Douglas fir
Douglas fir
165
243
16 5 Hantschel et al (1990)
243 Hantschel et al (1990)
193
95 7d
24-56°
21-42
17-641-
17-64^
115
Van Breemen and Van
Dyk (1988)
Van Breemen and Van
Dyk (1988)
Van Breemen and Van
Dyk (1988)
Van Breemen and Van
Dijk (1988)
Draayers et al (1989)
103
07
02 11 2
3-19°
Lovett (1992)
Royal Society (1983)
United Kingdom
Spruce
Cotton
forest
grass moor
1
0
9
4
8
8
0
0
13
4
5
0
23 4
124
Fowler
Fowler
etal
etal
(1989a)
(1989a)
a— Symbolizes data not available or, in the case of cloud deposition, not present
Measurements of total deposition data that do not include both a wet and dry estimate probably underestimate
total nitrogen deposition and are enclosed in parentheses
°Total nitrogen deposition was based on bulk deposition and throughfall measurements and does include
components of wet and dry deposition
Includes deposition from gaseous forms
10-100
-------
TABLE 10-15. NITROGEN INPUT/OUTPUT RELATIONSHIPS
FOR SEVERAL ECOSYSTEMS
Site/Vegetation
United States
Florida, Slash pine
Georgia, Loblolly pine
Minnesota, Spruce
North Carolina, Loblolly pine
North Carolina, Oak/hickory
North Carolina, Red spruce
North Carolina, White pine
North Carolina, White pine
New Hampshire, N hardwood
New Hampshire, N hardwood
New York, Deciduous
New York, Red spruce
Oregon, Douglas fir
Tennessee, Loblolly pine
Tennessee, Hardwood
Tennessee, Hardwood
Tennessee, Hardwood
Tennessee, Oak forest
Tennessee, Oak forest
Tennessee, Shortleaf/pine
Tennessee, Yellow/poplar
Washington, Douglas fir
Washington, Douglas fir
Washington, Red alder
Washington, Silver fir
Wisconsin, N hardwoods
Canada
Ontario, Maple
Federal Republic of Germany
Norway spruce
Beech
Netherlands
Oak
Oak/birch
Oak
Mixed deciduous
Inputs
(kg/ha/year)
59b
90b
75b
15 Ob
82°
27 lb
8 8°
74b
65
23 6
80b
15 9b
20
87b
13 2b
130
87
7 0-8 Od
11 5b
87
77
1 7
47b
70 Ob
1 3
56
78
21 8
21 8
450
540
560
630
Effluxa
(kg/ha/year)
0
0
0
0
32
11 0-20 0
02
0
40
174
10
30
1 5
0-20
44
3 1
1 8
125
3 2
1 8
35
06
0
710
27
005
182
149
44
220
780
280
680
Reference
Van Miegroet et al (1992)
Van Miegroet et al (1992)
Van Miegroet et al (1992)
Van Miegroet et al (1992)
Cole and Rapp (1981)
Van Miegroet et al (1992)
Cole and Rapp (1981)
Van Miegroet et al (1992)
Bormannetal (1977)
Likens et al (1977)
Van Miegroet et al (1992)
Van Miegroet et al (1992)
Sollinsetal (1980)
Van Miegroet et al (1992)
Kelly and Meagher (1986)
Henderson and Harris (1975)
Cole and Rapp (1981)
Kelly (1988)
Kelly and Meagher (1986)
Cole and Rapp (1981)
Cole and Rapp (1981)
Cole and Rapp (1981)
Van Miegroet et al (1992)
Van Miegroet and Cole (1984)
Turner and Singer (1976)
Pastor and Bockheim (1984)
Foster and Nicolson (1988)
Cole and Rapp (1981)
Cole and Rapp (1981)
Van Breemen et al (1987)
Van Breemen et al (1987)
Van Breemen et al (1987)
Van Breemen et al (1987)
10-101
-------
TABLE 10-15 (cont'd). NITROGEN INPUT/OUTPUT RELATIONSHIPS
FOR SEVERAL ECOSYSTEMS
Site/Vegetation
Norway
Spruce
Sweden
Coniferous
United Kingdom
Mixed hardwood
USSR
Norway spruce
Inputs
(kg/ha/year)
11 2b
2 1
58
1 1
Effluxa
(kg/ha/year)
0
06-10
126
09
Reference
Van Miegroet et al (1992)
Rosen (1982)
Cole and Rapp
Cole and Rapp
(1981)
(1981)
An estimate based on nitrogen losses from the soil profile or from stream flow out of a watershed
Includes precipitation, cloud (where appropriate), particulate, and gaseous forms of nitrogen deposition
"includes nitrogen inputs from precipitation and particulate forms of deposition
Mean of two oak forests in eastern Tennessee
northeastern United Sates and southern Canada (Olsen, 1989) Peak values were 5 and
25 kg/ha/year for ammonium and nitrate, respectively Similar wet deposition data for 1987
showed peak deposition rates of 3 5 and 16 kg/ha/year for ammonium and nitrate,
respectively (National Atmospheric Deposition Program, 1988) Zemba et al (1988)
summarized wet nitrate deposition data from 77 stations located in eastern North America
and found that peak nitrate deposition (>20 kg/ha/year) occurred between Lakes Michigan
and Ontario. They also found the temporal pattern of nitrate deposition was quite even
throughout the year (Schwartz, 1989) Wet deposition of ML,"1" ui Europe ranges between
3.5 and 17 3 kg NH4+/ha/year (Buijsman and Ensman, 1987, Heilet al, 1987) Boring
et al. (1988) have also published an extensive review of the sources, fates, and impacts of
nitrogen inputs to terrestrial ecosystems
For an oak-hickory forest in eastern Tennessee, dry deposition made up greater than
80% of the total atmosphenc deposition of nitrogen ions (Lindberg et al, 1986) Barne and
Sirois (1986) estimated that dry deposition contributed 21 to 30% of total NO3" deposition in
eastern Canada Lovett and Lindberg (1986) also concluded that dry deposition of nitrate is
the largest form of inorganic nitrogen deposited to oak-hickory forests in eastern Tennessee
Significant nitrogen inputs from the deposition of NO2 have been predicted (Hanson et al ,
10-102
-------
10-103
-------
1989; Hill, 1971; Kelly and Meagher, 1986) Duyzer et al (1987) has also predicted that
dry deposition of NH3 can reach levels as high as 54 kg/ha/year in areas of high ambient
concentration (0 017 ppmv) Typical values of NH3 deposition in central Europe and
Scandinavia range between 20 and 40 kg/ha/year (Grennfelt and Hultberg, 1986)
Based on the current rates of nitrogen deposition (loading) occurring in North America
(Tables 10-14 through 10-16), one might conclude that current rates of nitrogen deposition in
North America are sufficient to induce at least minor changes in some ecosystems (i e , rates
of deposition in North America exceed some of the critical load levels proposed for Europe)
However, because ecosystems have a variable capacity to buffer changes caused by elevated
inputs of nitrogen, and because deposition has been taking place for so many years, it is
difficult to make general conclusions about the type and extent of change resulting from
nitrogen deposition in North America Furthermore, current estimates of total nitrogen
deposition to ecosystems and regions of the United States (Tables 10-14 through 10-16)
usually do not account for gaseous nitrogen losses from ecosystems (e g , N2O and NH3),
therefore, the estimates of net nitrogen deposition may be overestimated (Wetselaar and
Farquhar, 1980, Bowden, 1986, Anderson and Levine, 1987, Schimel et al, 1988) Melillo
et al (1989) indicate that losses of nitrogen from ecosystems in the form of N2O are likely to
average in the range of 2 to 4 kg nitrogen/ha/year Higher levels of atmospheric nitrogen
deposition are also expected to lead to increased rates of N2O emissions
10.7 ECOSYSTEM EFFECTS-WETLANDS AND BOGS
10.7.1 Introduction
The diverse ecosystems that make up the biosphere interact through the cycling of
essential elements and compounds The availability of these essential elements determines
the rates of biological processes within a given ecosystem For example, the availability of
nitrogen in the form of NO3" or NH4+, which cycles through an enormous atmospheric pool
of NŁ, is an important determinant of the productivity of ecosystems Ecosystems interact
and function in different ways with complex feedback mechanisms, they influence the cycles
of essential elements and, to some extent, even the earth's climate
10-104
-------
TABLE 10-16. BULK DEPOSITION OF NITROGEN IN NORTH AMERICAN
WETLANDS (kg nitrogeia/ha/year)3
Site
Chesapeake Bay, riverine tidal
emergent marsh
Massachusetts, salt marsh
Massachusetts, basin bog
Minnesota, spruce bog
Minnesota, spruce bog
Iowa, praine marsh
Florida, everglades
Manitoba, emergent marsh
Ontario, poor fen
NH4+
27
14
25
1 7
3 0
40
30
NR
NR
N03'
43
23
50
17
20
40
96
NR
3 1
Org-N
47
3 9
NR
38
05
NR
NR
NR
NR
Tot-N
117
76
73
55
6 6-12 08
Reference
Jordan etal (1983)
Valiela and Teal (1979)
Hemond (1983)
Verry and Timmons (1982)
Urban and Eisenreich (1988)
Davis etal (1983)
Flora and Rosendahl (1982)
Kadlec (1986)
Bayley et al (1987)
aNH4 = Ammonium ion
NOg = Nitrate ion
Org-N = Organic nitrogen
Tot-N = Total nitrogen
NR = Not reported
Wetlands fulfill an important role in these global cycles as net sources and sinks for
biogenic gases They transfer to the atmosphere globsilly significant quantities of methane
(CH4) (Harnss et al , 1982, 1985) and reduced sulfur gases (Steudler and Peterson, 1984)
Elkins et al (1978) discuss the possibility that coastal marshes may function as net sinks for
N2O Because of the anaerobic nature of their waterlogged soils, decomposition of organic
matter in wetland soils is incomplete Consequently, wetlands function as sinks and long-
term storage reservoirs for organic carbon It has been estimated that wetlands once
sequestered a net of 57 to 83 X 106 metric tons of caibon per year worldwide, although
recent widespread drainage of wetland soils has shifted the carbon balance (Armentano and
Menges, 1986) Although this rate of carbon uptake is small in comparison to other global
carbon fluxes, such as the annual release of carbon from combustion of fossil fuel (5 to
6 x 109 metnc tons/year, Rotty, 1983) or the net uptake of CO2-carbon by the ocean
(1 6 X 109 metnc tons/year, Tans et al , 1990), it is important when the net balance between
large fluxes is considered and it is certainly important over geologic tune scales (Armentano
and Menges, 1986)
10-105
-------
These gases (CH4, N2O, and reduced sulfur compounds) modify atmospheric chemistry
and global climate. The destruction of O3 in the upper atmosphere by its reaction with N2O
is one example Combustion sources are currently raising the atmospheric concentration of
N2O (Hao et al, 1987) The rise in anthropogenic releases of NOX to the atmosphere also
increases the deposition of biologically available forms of nitrogen onto the landscape, with
potential effects on productivity (or other aspects of function) and community structure
Locally, wetlands function as habitats for wildlife, flood control systems, stabilizers and
sinks for sediments, storage reservoirs for water, and biological filters that maintain water
quality Studies of riparian forests, for example, generally indicate that they exert a positive
influence on the water quality of receiving streams by intercepting and removing nutrients
from runoff (Yates and Sheridan, 1983, Bnnson et al , 1984, Peterjohn and Correll, 1984,
Quails, 1984). And as sediment traps, salt marshes like those on the Louisiana coast can
accumulate annually an impressive 0 76 cm of sediment (DeLaune et al, 1983) These
functions are a great monetary value to society (Westman, 1977)
Wetlands also harbor a disproportionate (relative to habitat area) share of flora that are
threatened by extinction Of the 130 plant species from the conterminous United States that
are formally listed as endangered or threatened (Code of Federal Regulations, 1987),
18 species (14%) occur principally in wetland habitats On the national list of plant species
that are identified as endangered (Status LE or PE), threatened (Status LT or PT), or
potentially threatened (Status 1 or 2), 1,776 species are listed for the conterminous United
States (Federal Register, 1985), and 302 (17%) of these occur principally in wetland habitats
The national hst of plant species that occur in wetlands includes 6,728 entries (Reed, 1988),
and because this hst includes plant species found primarily in upland habitats as well as
plants from the entire United States and its territories, we can estimate conservatively that the
endangered or potentially threatened wetland plant species represent an alarming 4 5 %
(302/6,728) of this total
Wetland plants are undoubtedly threatened because of loss of habitat, which in the
United States, has been largely a consequence of agricultural development involving drainage
(Tiner, 1984). Total wetland area, including intertidal and palustnne areas, in the
conterminous United States (Figure 10-20) totaled 437,609 km2 during the mid-1950s and
o
decreased to 400,567 km , or 5 1 % of total land area, by the mid-1970s (Prayer et al ,
10-106
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Figure 10-20. Map of the United States showing location of the major groups of inland
freshwater marshes. Contours delineate physiographic regions.
Source Hofstetter (1983)
1983) The net loss of wetland habitats during these two decades is equivalent to an annual
rate of loss of 1,852 km /year (715 mi /year) However, it can also be concluded that
current rates of atmospheric nitrogen deposition in parls of Europe, elevated by
anthropogenic emissions, alter the competitive relationships among plants and threaten
wetland species adapted to infertile habitats Those data are leviewed here, and on this
basis, we can anticipate similar effects of atmospheric nitrogen deposition in the
United States
10.7.2 Atmospheric Nitrogen Inputs
Atmospheric nitrogen inputs occur as both wet and dry deposition Most studies of
atmospheric nitrogen inputs into wetlands focus only on wet deposition or bulk deposition
Accurate measurements of wet deposition are earned out by analyzing nitrogen in
precipitation immediately following a precipitation event Frequently, however, rainfall is
10-107
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accumulated over some period of tune before it is analyzed, and the resulting measurement
of deposition rate is usually referred to as bulk deposition Bulk deposition rates combine
wet deposition with some component of dry deposition Where dry deposition has been
carefully measured, it has been concluded that (1) the relative importance of wet and dry
deposition vanes geographically, (2) that dry deposition can exceed wet deposition (Boring
et al., 1988), and (3) that bulk precipitation samplers underestimate the combined dry plus
wet deposition rate (Dillon et al, 1988) The available wet surface area of vegetation, onto
which nitrogen gases will diffuse, significantly affects the dry deposition rate (Heil et al,
1987). Levy and Moxim (1987) modeled the fate of NOX emissions to the atmosphere and
concluded that dry deposition accounts for greater than one-half of the total NOX deposition
in North America.
The rate of bulk NO3~ deposition has been shown to be positively correlated with the
concentration of NO2 in the air Press et al (1986) measured atmospheric concentrations of
NO2 and bulk deposition of NO3~ at several sites in northern Britain for 18 mo Nitrogen
dioxide concentrations (2-week averages) ranged from near zero to 25 pg/m and were
correlated significantly (p < 0 001) with concentrations of NO3~, collected in bulk samplers,
that varied from near zero to about 3 mg mtrogen/L
A third, and rarely measured, mechanism of deposition that is locally important is the
interception or capture of fog or cloud droplets by vegetation Lovett et al (1982) estimated
that the cloud deposition of NO3" in an alpine habitat in New Hampshire was 101 5 kg
nitrogen/ha/year, compared to a bulk deposition rate of 23 4 kg nitrogen/ha/year The same
phenomenon was observed by Woodin and Lee (1987), who collected 1 45 tunes as much
water as "throughflow" (collected beneath vegetation) passing through experimental
Sphagnum mats in the field as from adjacent bulk deposition gauges Their data also suggest
that the deposition of solutes by this mechanism is important, and that bulk precipitation
samplers underestimate total deposition
Table 10-16 summarizes several studies that report wet or bulk deposition rates of
nitrogen in North American wetlands From the data presented, it may be concluded that
bulk deposition rates of NH4 , NO3", and organic nitrogen vary geographically and their
relative importance vanes In general, however, inputs of NO3", NH4+, and organic
nitrogen are all of the same order of magnitude, and their combined rate of deposition vanes
10-108
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from 5 5 to 12 1 kg mtrogen/ha/year Other studies, however, indicate that wet NO3"
deposition alone exceeds 15 kg nitrogen/ha/year over most of the midwest and 20 kg
mtrogen/ha/year in portions of the northeast United States (Zemba et al, 1988)
Rates of nitrogen deposition, and NH4+ deposition in particular, in areas of western
Europe are greater than in North America In areas of Britain, bulk deposition rates of
43 and 46 kg/ha/year have been reported (Press and Lee, 1982, Ferguson et al., 1984) The
combination of NO3" and NH4+ deposition downwind of Manchester and Liverpool is
reported to be 32 kg mtrogen/ha/year (Lee et al., 1986) Nitrogen deposition in fens near
Utrecht was 21 kg mtrogen/ha/year of inorganic nitrogen and 3 to 5 kg mtrogen/ha/year of
organic nitrogen in bulk precipitation and 18 kg mtrogen/ha/year of inorganic nitrogen in dry
deposition (Koerselman et al , 1990) Roelofs (1983) reported that wet deposition of
nitrogen in the Netherlands averages 15 kg mtrogen/ha/year and is as great as 20 to 60 kg
mtrogen/ha/year in areas of intensive stockbreeding, 75 to 90% of this being deposited as
NH4+ In Europe, 81 % of total NH3 emissions are from livestock wastes, with the greatest
emission densities concentrated in the Netherlands and Belgium (Buijsman, 1987) Annual
NH3 emissions from animal excreta in the Netherlands are reported to be 230 kt/year
(Van der Molen et al, 1989) or about 60 kg/ha/year country-wide
The chemistry of surface runoff from watersheds is probably of greater significance to
most wetlands than the chemistry of direct deposition, but the nitrogen load of surface runoff
probably mcreases with nitrogen deposition and with the size of the catchment area
Atmospheric deposition accounts for a large fraction of the total nitrogen entering watersheds
(Robertson and Rosswall, 1986) Atmosphenc deposition apparently has become a major
source of NO3" to surface waters in North America, especially in the east and upper midwest
(Smith et al , 1987a), and mcreases in total nitrogen concentration at stream monitoring
stations are strongly associated with high levels of atmospheric nitrate deposition (Smith
et al, 1987b) However, the direct contribution made by atmospheric deposition to the
nitrogen load in surface water because of nitrogen in surface runoff is unknown.
Measurements by Buell and Peters (1988) of stream chemistry in Georgia indicated that 93 %
of the precipitation inputs of NH4+ and NO3" were retained by the watershed A study by
Correll (1981) of mass nutrient balances of a small watershed of the Rhode River estuary on
the Chesapeake Bay showed that total wet nitrogen deposition to 88 ha of tidal marshes and
10-109
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mudflats was 740 kg nitrogen (8 4 kg/ha) in 13 mo, compared to total nitrogen in runoff
from 2,050 ha of watershed of 10,000 kg nitrogen Only about 7% (740 kg/10,740 kg) of
the nitrogen entering the wetland was from direct deposition However, in as much as
nitrogen deposition onto the watershed (8 4 kg/ha X 2,050 ha = 17,220 kg) exceeded total
runoff from the watershed to the wetland (10,000 kg), deposition could have contributed the
majority of nitrogen entering the wetland indirectly through runoff But the contributions of
other nitrogen sources to runoff, such as fixation, fertilizer, and animal waste, were not
given.
10.7.3 The Wetland Nitrogen Cycle
The feature of wetlands that sets them apart from terrestrial ecosystems is the anaerobic
(oxygen-free) nature of their waterlogged soils, which alters the relative importance of
various microbial transformations of inorganic and organic nitrogen compounds Generally,
the absence of O2 retards the decomposition of organic matter (Tate, 1979, DeLaune et al,
1981; Van der Valk and AttiwiU, 1983, Godshalk and Wetzel, 1978, Clark and Gilmour,
1983) Complex aromatic ring structures are more resistant to microbial attack under anoxic
conditions (Tate, 1979), leading to the formation and buildup of peat in wetland
environments Anoxic soils also favor the rapid conversion of NO3" to N2O or N2 This
process is accomplished by bacteria and is referred to as denitnfication or dissinulatory
nitrate reduction, and it results in quantitatively important losses of nitrogen from wetland
ecosystems Finally, the hydrology of wetlands favors diffusive exchanges of nitrogen
compounds to and from sediments and advective transport (earned by water) of nitrogen
compounds between ecosystems This often results in movements of NH4+ from anoxic
sediments to the oxidized surface sediment or water column, where nitrification (the
oxidation of NH4+ to NO3" by bacteria) can occur, and the return movement of NO3" to the
anoxic sediment layers, where denitnfication can occur The nitrogen cycle in wetlands has
been reviewed recently by Reddy and Patrick (1984), Savant and De Datta (1982), and
Bowden (1987) Important steps in the nitrogen cycle are summarized in Section 10 3
Table 10-17 presents the nitrogen budgets of wetlands that exhibit a wide range of
nitrogen inputs. The two bog sites (Table 10-18) are representative of wetlands that contain
10-110
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TABLE 10-17. NITROGEN BUDGETS OF SELECTED WETLANDS
(kg nitrogen/ha/year)3
Location and Wetland Type
INPUTS
Precipitation
Fixation
Surface, ground or tidal water
Total
INTERNAL CYCLE
Plant assimilation
Mineralization
OUTPUTS
Demtnfication
Ammonia volatilization
Surface or subsurface DIN export
Surface or subsurface ON export
Total
UK
Salt
Marshb
MR
3 36
43 41
2254
1949
378
NR
24m
43 Om
MA
Salt
Marshc
79
680
6680
743 9
214 O1
193 O1
1430
035
1020
5520
7974
Dutch
Rech
Fend
43 7h
2 1
73
53 1
274 Ok
244 O1
1 4
NR
2 1
45 8n
493
Dutch
Disc
Fend
42 Oh
127
209
756
90 Ok
79 O1
1 1
NR
67
80 4n
882
French
Heath6
8 1
13
0
94
820
740
NR
NR
30
30
MA
Bogf
75
3 36
0
109
380
260
10
Trace
20
10
40
MN
BogS
86
05
0
091
660
500
18
NR
0
20
38
aUK = United Kingdom
MA = Massachusetts
MN = Minnesota
NR = Not reported
DIN = Dissolved inorganic nitrogen
ON = Dissolved and particulate organic nitrogen
Abd Aziz and Nedwell (1986a,b) salt marsh dominated by Puccmellia mantana (a grass)
°Vahela and Teal (1979) salt marsh dominated by Spartina alterniflora
Koerselman et al (1990) Dutch eutrophic recharge and mesotrophic discharge fens, respectively
^oze (1988) mesophilous heathland (shrub bog) dominated by Erica cilians (heath) and Ulex minor
Urban and Eisenreich (1988) ombrotrophic Sphagnum bog forested with black spruce (Picea manana) and
with an understory of shrubs and sedges
gHemond (1983) ombrotrophic bog dominated by Sphagnum
Includes bulk plus dry deposition of inorganic and organic nitrogen
Represents the net exchange of nitrate ion (the major component) and small particulate organic nitrogen rather
than an absolute rate
•"Calculated from Morns et al (1984) and Vahela et al (1984)
From Verhoeven et al (1988), assuming a root shoot quotient of 1 0
'From Verhoeven et al (1988)
"^Represents the net exchange of dissolved organic nitrogen (the major component), ammonium ion, and large
particulate organic nitrogen rather than an absolute rate
"includes primarily hay harvested by mowing
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TABLE 10-18. RESULTS OF NITROGEN FERTILIZATION EXPERIMENTS
IN WETLAND ECOSYSTEMS
Salt Marsh Ecosystems
Spartina
Spartina
Spartina
Spartina
Spartina
Spartina
Spartina
Pucanellia
Pucdnellia
Carex
Panicum hemitomon
Paniaim hemitomon
Typha glauca
Spargamum eurycarpum
bog
bog
fen
wet grassland
Rate of
Nitrogen
Application
(kg/ha/year)
200
200
220
650
670
1,040
3,120
320
320
320
30
100
1,350
1,350
300
7
450
450
Length of
Study
(years)
1
1
3
3
2
1
2
2
2
2
1
1
2
2
1
1
1
1
Control
Biomass
(i/m2)*
1,660
816
320
320
250
450
235
64
64
65
1,320
1,320
1,726
637
180
200
350
400
Percent
Increase
16
25
131
269
120
100
413
175
73
146
6
42
36
86
25
10
57
68
Nitrogen-
Form
Applied
NH4+
NH4N03
Sludge
Urea
Sludge
NH4+
NH4+
NH4+
N03"
NH4+
NH4+
NH4+
NH4N03
NH4N03
Urea
Sludge
Mineral-N
Mineral-N
Reference
Patrick and Delaune
(1976)
Gallagher (1975)
Valielaetal (1975)
Valielaetal (1975)
Valiela and Teal (1974)
Haines (1979)
Morns (1988)
Cargill and Jeffenes
(1984)
Cargill and Jeffenes
(1984)
Cargill and Jeffenes
(1984)
Delaune et al (1986)
DeLaune et al (1986)
Neely and Davis (1985a)
Neely and Davis (1985a)
Sanville (1988)
Sanville (1988)
Vermeer (1986)
Vermeer (1986)
*Control biomass is the maximum, nonfertihzed aboveground standing crop
Percent increase indicates the response of control biomass during the year of fertilization at the indicated rate
of application, computed as 100 X (Expenmental-Control)/Control
°NH4 = Ammonium ion
= Ammonium nitrate
— Nitrate ion
Mineral-N= Mineral nitrogen
plant species that are adapted to low levels of nitrogen They are examples of ombrotrophic
bogs, meaning that they receive nutrients exclusively from precipitation They develop
where precipitation exceeds evapotranspiration and where there is some impediment to
drainage of the surplus water (Mitsch and Gosselink, 1986) Bogs are dominated by
Spfiagnum spp. and may be sparsely forested The Sphagnum builds a dense layer of peat,
10-112
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raised above the elevation of the surrounding land so that they receive neither runoff from
uplands nor inputs from groundwater Peat-forming bog ecosystems are widely distributed
throughout the northern hemisphere, but they are most common in formerly glaciated
regions The distribution of peatland area in North America is shown in Figure 10-21 The
bog ecosystems represented in Table 10-17 are located in Minnesota (Urban and Eisenreich,
1988) and Massachusetts (Hemond, 1983)
05-10%
Peatland Area
>10% Peatland
Figure 10-21. Distribution of North American peatlands.
Source Mitsch and Gossehnk (1986)
In bog ecosystems, the most important nitrogen inputs are from wet and dry deposition
(see the row labeled "precipitation" in Table 10-17) The total input of nitrogen m these
examples is about 10 kg nitrogen/ha/year, and atmospheric deposition accounts for most of
this (Urban and Eisenreich, 1988, Hemond, 1983) Also note that the total nitrogen outputs
10-113
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from the system are approximately 4 kg nitrogen/ha/year The outputs are accounted for by
denitrification (1 to 1 8 kg nitrogen/ha/year) and by export in runoff of dissolved inorganic
nitrogen (as NH4+) and dissolved organic nitrogen (DON) No export of particulate organic
nitrogen was reported, nitrogen accumulated in plant tissues is largely recycled within the
bog.
Bog wetlands are representative of one end of a continuum, but there are also other
wetlands where atmospheric nitrogen deposition represents a significant fraction of the total
input of inorganic nitrogen For example, wetfall contributed more than 95 % of the NH4+
and NO3" entering the 1,000-km2 Shark River Slough, the major fresh water drainage of
Everglades National Park (Flora and Rosendahl, 1982) However, the importance of organic
nitrogen in the surface inflow may be considerable, depending on how easily or rapidly it is
mineralized by the microbial community In this ecosystem, rainfall is about 84% of total
water input, and one can generalize that the significance of atmosphenc nitrogen deposition
increases in wetlands as rainfall increases as a fraction of the total water budget
The French heathland or shrub bog (Table 10-17) is another example of a wetland with
low nitrogen inputs and outputs, but with an intermediate rate of internal cycling The
moderate size of the internal nitrogen cycle depends on the accumulation of a large quantity
of organic nitrogen in the soil humus (Roze, 1988) A fraction of this organic pool
mineralizes each year and is assimilated by the plant community Organic and inorganic
nitrogen in the soil is about 91 % of total nitrogen in this heathland ecosystem, with the
remaining 9 % being contained within the plant biomass A moderate rate of nitrogen
mineralization in the soil is balanced by assimilation by the plant community, and nitrogen is
largely conserved within the ecosystem
In the Dutch fens (Table 10-17), the inputs and outputs of nitrogen are intermediate
between those of the bogs and salt marshes Both fens are influenced by their close
proximity to heavily fertilized pastures, by atmosphenc nitrogen deposition, and by annual
mowing and harvest of aboveground vegetation The fen that occupies a site of groundwater
recharge is influenced by water that is diverted from the highly polluted River Vecht during
periods of high evapotranspiration, and the discharge fen is influenced by nutrients in
groundwater (Verhoeven et al, 1988) However, atmosphenc nitrogen deposition in these
fens supplies more nitrogen than all other inputs combined (Table 10-17)
10-114
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The coastal salt marsh ecosystems in Table 10-17 are characteristic of wetlands that are
adapted to large nitrogen inputs Coastal salt marshes have a temperate, worldwide
distribution They exist within the intertidal zone and are alternately flooded and drained
daily by the action of the tides The example from Massachusetts is a salt marsh dominated
by the grass Spartina alterniflora (Valiela and Teal, 1979) The salt marsh example from the
United Kingdom in Essex is dominated by the grass Puccmellia mantima (Abd Aziz and
Nedwell, 1986b)
In salt marsh ecosystems, the most important nitrogen inputs are from those brought
into the marsh in tidal water and, in some cases, groundwater Input of paniculate organic
nitrogen from sedimentation and/or NO3" is apparently Ihe dominant mechanism by which
these ecosystems remove nitrogen from surface water because the diffusion gradients for
NH4+ and DON normally favor diffusion out of the sediment These surface and
groundwater sources of nitrogen are one to two orders of magnitude greater than inputs from
precipitation (Table 10-17) In the Massachusetts salt marsh, groundwater inputs of NO3"
and DON are important and account for 60 and 56 kg mtrogen/ha/year, respectively, of the
total inputs (Valiela and Teal, 1979) In contrast, the Essex, United Kingdom, marsh is not
influenced by groundwater (Abd Aziz and Nedwell, 1986b) Both salt marshes have large
nitrogen inputs from tidal water, and in the Massachusetts marsh, these are largely as NE^+
(54 kg nitrogen/ha/year), DON (337 kg nitrogen/ha/year), and particulate organic nitrogen
(139 kg nitrogen/ha/year) (Valiela and Teal, 1979) There are additional inputs and outputs,
such as deposition of bird faeces and shellfish harvest, but these are insignificant in
comparison to other rates (Valiela and Teal, 1979)
The large inputs of nitrogen in salt marshes are balanced by equally large outputs
(Table 10-17), but there are important transformations that take place within the marsh
Denitrification accounts for 17 9 % of the total nitrogen loss from the Massachusetts marsh
Because the demtnfication rate is greater than the combined inputs of NO3", this implies that
rates of nitrification are large In both marshes, the greatest nitrogen losses occur in tidal
water exchange, and in the Massachusetts marsh, there is a net loss of all forms of dissolved
nitrogen in tidal water The Massachusetts marsh exports large amounts of NI^+ (73 kg
nitrogen/ha/year), NO3" (25 kg nitrogen/ha/year), DON (380 kg nitrogen/ha/year), and
particulate organic nitrogen (17 kg nitrogen/ha/year) (Valiela and Teal, 1979)
10-115
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Nitrogen inputs and outputs in tidal water were given as net exchanges of different
nitrogen components ui the Essex, United Kingdom, study (Abd Aziz and Nedwell, 1986b),
rather than as absolute rates This is the reason for the discrepancy in the rates of
tidal-water imports and exports of nitrogen in the Essex and Massachusetts marshes
(Table 10-17) However, valid comparisons can be made of the net exchanges There is a
large net export of DON (43 kg nitrogen/ha/year) from the Essex marsh (Abd Aziz and
Nedwell, 1986b), and this is consistent with the net DON loss in tidal water of 45 kg
nitrogen/ha/year from the Massachusetts marsh (Valiela and Teal, 1979) The marshes differ
in the net tidal-water exchanges of other forms of nitrogen
The rate of internal nitrogen cycling (assimilation and mineralization) within ecosystems
is directly proportional to the rate of primary production (e g , Verhoeven and Arts, 1987),
although high rates of productivity can be supported by high external nutrient inputs when
conditions are unfavorable for high mineralization rates (Verhoeven et al, 1988)
Mineralization rates differ greatly between the wetland types represented in Table 10-17
Nitrogen assimilation by the plant communities vanes from 38 to 66 kg nitrogen/ha/year in
the bog ecosystems, compared to 225 to 274 kg nitrogen/ha/year in the salt marsh and fen
ecosystems, respectively The nitrogen cycle in the bog and heathland ecosystems is largely
closed (Figure 10-22) In contrast, the nitrogen cycle in salt marshes and fens is open, and
there is a great exchange of nitrogen with adjacent systems (Figure 10-21) In all these
ecosystems, the rate of nitrogen mineralization almost balances plant assimilation in the
manner of a closed cycle (Table 10-17). However, it is unlikely that the salt marsh could
function as a closed system and maintain its productivity or community structure Likewise,
it is unlikely that the bog ecosystem could maintain its community structure if the nitrogen
inputs were greatly increased by some means In general, as the input rate of nitrogen
increases, there are concomitant increases in the output rate and magnitude of the internal
cycle (Table 10-17) In ecosystems with closed nutrient cycles and small rates of internal
cycling, like bogs, if nitrogen loadings increase significantly, then we can predict that
productivity will increase, but as will be discussed later, the increased productivity will be
accompanied by changes in species composition to those adapted to an elevated nutrient
regime (Figure 10-22)
10-116
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Low N inputs
Low N outputs
Oligotrophic habitats
(e g , ombrotrophic bogs)
Eutrophic habitats
(e g , salt marshes)
Low N inputs
Assimilation
High N inputs
Assimilation
irahzabon
Low N outputs
Low
Productivity
Internal Cycling
Moderate
Species Diversity
Moderate
Productivity
Internal Cycling
High
Mineralization Species Diversity
High N outputs
Low
Species Diversity
High
Species Diversity
Internal Cycling
Mineralization
Figure 10-22. Conceptual relationships among trends in nitrogen cycling, productivity,
and species diversity along a gradient from oligotrophic (nutrient-poor) to
eutrophic (nutrient-rich) habitats.
10.7.4 Effects of Nitrogen Loading on Wetland Plant Communities
10.7.4.1 Effects on Primary Production
Numerous field experiments involving nitrogen fertilization have documented that
primary production in wetland ecosystems is commonly limited by the availability of
nitrogen Results of this type of experiment are presented in Table 10-18 In all of the
fertilization experiments included in the table, only sewage sludge, urea, or mineral nitrogen
m the form of NH4+ or NO3" were applied Except in the case of sewage sludge
applications, where the numerous elements contained in sludge preclude attributing the results
10-117
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to any specific element, the stimulation of growth that was observed can be attributed solely
to application of nitrogen Rates of application ranged from 7 to 3,120 kg nitrogen/ha/year
(Table 10-18), and in most studies, these have been 1 to 2 orders of magnitude greater than
rates of atmospheric deposition (Table 10-16) These applications stimulated increases in
standing biomass by 6 to 413% (Table 10-18)
Several studies have investigated the effects of different nitrogen sources Cargill and
Jefferies (1984) found that applications of NH4+ increased production of Puccinellia
phryganodes (a grass) in a subarctic salt marsh by 175 %, whereas equivalent applications of
NO3" increased production by only 73 % Applications of NO3" were perhaps less effective
than NH4+ because of denitnfication of NO3" by bacteria in the anaerobic marsh sediments
This demonstrates the importance of competition between plants and microbes for specific
inorganic nitrogen compounds, with plants being the best competitors for NH4+
The greatest stimulation of growth is often achieved when nitrogen applications are
combined with applications of other nutrients In the study of Cargill and Jeffenes (1984),
applications of inorganic phosphate (P^ combined with NH4+ stimulated production to a
greater extent than NH4+ alone Sanville (1988) observed that combinations of nitrogen, in
the form of urea, and Pt stimulated production in a Sphagnum bog to a greater extent than
nitrogen applications alone, and that singular additions of Pt had no significant effect on
growth. These results demonstrate that other nutrients, Px in these examples, become
secondarily limiting after nitrogen applications reach a threshold
In one study of a wet heathland in the central Netherlands, total aboveground biomass
failed to respond on experimental sites fertilized for 3 years at a rate of 200 kg
mtrogen/ha/year, but sites fertilized with 40 kg phosphorus/ha/year did show a significant
increase in biomass (Aerts and Berendse, 1988) Thus, wetlands are not universally limited
by nitrogen However, as discussed above (see Section 10 5 2), the Netherlands is an area
of extreme high nitrogen deposition, and the threshold for nitrogen limitation is perhaps
exceeded by anthropogenic inputs in this area
Fertilization experiments of salt marshes in Massachusetts by Valiela and Teal (1974)
and in Louisiana by Patrick and Delaune (1976) involving singular applications of either
nitrogen or P, demonstrated that primary production was stimulated by nitrogen and not by
phosphorus. Vermeer (1986) obtained the same result in freshwater fen and wet grassland
10-118
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communities in the Netherlands However, fertilization with nitrogen increased the biomass
and dominance of grasses at the expense of other species in fen and wet grassland
communities Some Eqwsetum spp (horsetail) had a smaller biomass contribution upon
fertilization This tendency toward a change in species composition or dominance has also
been observed in other fertilization experiments Jefferies and Perkins (1977) found
species-specific changes in stem density at a Norfolk, England, salt marsh after fertilizing
monthly with 610 kg NO3"-nitrogen/ha/year or 680 kg NH4+-nitrogen/ha/year over a period
of 3 to 4 years
A final conclusion of the data in Table 10-19 is that the stimulation of primary
production by nitrogen applications is not a linear function of the rate of nitrogen application
This can be seen by comparing the results of fertilization studies of Spartma (Table 10-18)
The greatest increase in standing biomass, both in terms of absolute amount and in terms of
the percent increase, was obtained in studies where the control biomass was low This
implies that the in situ nitrogen supply in some wetlands already is near a threshold where
other factors become limiting Ultimately, available light energy, water, and temperature are
the limiting factors
The data included in Table 10-18 pertain to growth of aboveground biomass only
In several of these studies, measurements of belowground biomass were also made (Valiela
and Teal, 1974; Haines, 1979, Valiela et al, 1976, Gallagher, 1975) Results were variable,
with some studies showing a small decrease in living belowground biomass (Valiela et al,
1976), and others showing small increases in belowground macroorgamc matter (Gallagher,
1975) or no change (Valiela and Teal, 1974) The normal technique of coring sediments to
measure belowground production is subject to great error (Singh et al, 1984) However, the
evidence from controlled-growth experiments (Morns, 1982, Steen, 1984) clearly shows that
the response of leaf growth to increased nitrogen supply is much greater than the response of
roots
It should be emphasized that all of the fertilization studies summarized in Table 10-18
are short-term results in which nitrogen was applied for 3 years or less We cannot assume
that long-term nitrogen applications will yield the same results Studies of several wetland
ecosystems that have been fertilized for long periods by increased atmospheric inputs indicate
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TABLE 10-19. RATES OF NITROGEN DEPOSITION IN SEVERAL
AREAS OF NORTH AMERICA
Deposition Rate
(kg/ha/year)
Area
NH4+b Total
Source
Alaska0
(Poker Flat)
Sierra Nevada, CAd
(Emerald Lake)
Ontario, Canada®
(Experimental Lakes Area)
British Columbia, Canada6
Upper Midwestf
Southeastern United Statesg
(Walker Branch, TN)
New Hampshire6
CatskiUs0
Adurondacks
0 10 0 06 0 16 Galloway et al (1982)
1 11 1 19
2 30 Williams and Melack (1991a)
1 75 1 96 3 71 Linsey et al (1987)
3 64 1 82 5 46
4 20 2 94 7 14
7 56 2 52 10 08
6 50 2 80 9 30
8 12 4 09 12 24
8 26 2 66 10 92
Feller (1987)
DnscoUetal (1989a)
Lindbergetal (1986)
Likens (1985)
Stoddard and Murdoch (1991)
DnscoUetal (1989a)
aNO3" = Nitrate ion
NH4 = Ammonium ion
°Dry deposition estimated as 35% of total deposition
Diy deposition sampled as part of snowpack, no correction for dry deposition made
°Bulk precipitation measurements, no correction for dry deposition made
Values corrected for dry deposition based on ratios in Hicks (1989)
^Includes estimates for dry deposition and gaseous uptake of nitrogen areas, dissolved organic nitrogen can
occur in greater concentrations than the inorganic species (Moore and Nuckols, 1984)
that changes in species composition and succession accompany the increases in nitrogen
loadings and primary production These studies are summarized below
One implication of a long-term increase in leaf growth is that the demand for mineral
elements and water from the soil will increase Howes et al (1986) observed that the rate of
evapotranspiration increased from a salt marsh dominated by Spartma altemiflora in sites
where aboveground biomass was increased by nitrogen fertilization Increased
evapotranspiration can influence the direction of succession of some wetlands by altering the
water balance of the soil The feasibility of this mechanism to alter bog succession was
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demonstrated in a model by Logofet and Alexandrov (1984) Their model suggests that
nitrogen inputs greater than a threshold of 7 kg nitrogen/ha/year can change the direction of
succession from that of an open oligotrophic bog to a mesotrophic bog dominated by trees
Furthermore, in flowing water systems, like salt marshes, an increase in aboveground
production should lead to an increased export from the system of nutrients that are
incorporated in or leached from aboveground biomass Therefore, the long-term ecosystem
and community responses to increased inputs of nitrogen can not be predicted from results of
short-term field experiments like those summarized in Table 10-19
10.7.4.2 The Fate of Added Mineral Nitrogen
Experiments in the field and laboratory have followed the fate of applied nitrogen by
15 15
using N as a tracer This stable isotope, N, comprises 0 37% of naturally occurring
nitrogen It can be quantified together with the more common isotope of nitrogen,
mtrogen-14, with a mass spectrometer and is used experimentally much like radioactive
isotopes, except that N is normally used in greater than trace amounts due to the lower
sensitivity of the instrumentation used to detect it
Experiments in which different mineral forms of N were added to sediments in the
absence of plants demonstrate that mineral nitrogen is rapidly used by the microbial
community Smith and DeLaune (1985) added the equivalent of 100 kg nitrogen/ha in one
application as 15N-labeled NH4+ (15NH4+) to sediments of a shallow saline lake They
found 15 days after the addition, 20% had been converted to organic nitrogen in the
sediment, and the fraction in organic matter remained constant at this level for the remaining
337 days of the experiment The amount of NH4+ in the sediment decreased exponentially
to a nondetectable level by Day 200 Diffusion of NH4+ into the water column and
denitnfication accounted for a loss of 80% of the 1 NH4+ from the sediment
Lindau et al (1988) made smgle additions of either 15N-labeled NO3" or 15NH4+,
equivalent to 100 kg nitrogen/ha, to the floodwater within chambers containing swamp
sediment By Day 27, only 39 6% and 6 2% of the 1')N from NH4+ and NO3", respectively,
remained in the sediment and overlying water column The lemaimng fractions had been
lost from the chambers by denitnfication The loss of 60% of the applied l NH4+ within
27 days demonstrates that NH4+ can be rapidly converted to NO3" by nitrifying bacteria in
10-121
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aerobic parts of the system, and that NO3" diffuses into the anaerobic sediments where
denitrification occurs Nitrification was apparently the rate limiting step because the loss of
N by denitnfication was more rapid when it was applied as NO3"
DeBusk and Reddy (1987) made single additions of NH4+ to the floodwater above
cores of sediments taken from swamps that had been receiving primary wastewater effluent
for 2 and 50 years prior to the experiment The rate of application was equivalent to 15 kg
nitrogen/ha After 21 days, 0 5 to 2 3 % of the added nitrogen was recovered in the
floodwater, largely as NO3~, and 13 6 to 17 8 % was recovered in the sediment, largely as
organic matter The remaining 80% was apparently lost by denitnfication, indicating that
conversion of NH4+ to NO3~ and diffusion of NO3" to anaerobic sites of denitnfication is
rapid. This result is consistent with that of Lindau et al (1988) Furthermore, there was no
difference in the response of the two sediment types, which demonstrates that the
nitrification-denitrification potential of sediments is unchanged in sediment receiving sewage
effluent for 50 years However, the bacteria in the sediments must have a continuous supply
of suitable carbon substrates as well as nitrogen to sustain continuous nitrification-
denitrification reactions
Short-term measurements of slurrys of marl and peat sediments from the Florida
Everglades (Gordon et al, 1986) demonstrated that 10 to 34% of NO3" added at levels of
10 and 100 ^M (1 i*M = 14 p,g nitrogen/L) was rapidly denitrified within 24 h
Demtriflcation rates decreased following this initial burst of activity as the balance of the
added NO3" was converted to NH4+ This experiment suggests that the process of
dissimilatory nitrate reduction to ammonium (reammonification) competes successfully with
the denitrification process However, this experiment was conducted on sediment slurrys that
were incubated under a nitrogen atmosphere, which prevented nitrification reactions from
occurring. Under an oxygen atmosphere, nitrification would have generated a continuous
supply of NC>3" and denitnfication would then have consumed a greater fraction of the NO3"
over time.
The behavior of mineral nitrogen applied to vegetated wetland sediments is quite
different from the results descnbed above and indicates that plants successfully compete with
microbes for mineral nitrogen DeLaune et al (1983) followed the fate of 15NH4+ placed
below the soil surface in a Louisiana salt marsh dominated by Spartina altemiflora The
10-122
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singular application of NH4+ was equivalent to 72 kg nitrogen/ha At the end of the first
growing season, 93 % of the added nitrogen was recovered in aboveground biomass, roots,
and soil An average of 28 % was in aboveground bioniass and 65 % was in soil and
belowground biomass The high rate of recovery of N in vegetation and soil is consistent
with results of Buresh et al (1981) and Patrick and Delaune (1976) In the study of
DeLaune et al (1983), N recovered in soil and belowground biomass declined to 50% by
the end of the second growing season and to 43 % by the end of the third growing season
Nitrogen in aboveground biomass decreased to 12% of original N by the end of the third
growing season The annual declines were postulated to have occurred due to the loss of
nitrogen from the leaves, either by physical transport of aboveground plant material off the
site or by decomposition of leaf material at the sediment surface followed by nitrification-
demtnfication reactions Similar results were obtained in a freshwater marsh dominated by
Pamcum hemitomon (maiden cane) DeLaune et al (1986) added 30 kg/ha of
15 -4-
NH4 -nitrogen to sediments and recovered a mean of 80% in the combined aboveground
(18%) and belowground biomass and soil (62%) at the end of the first growing season
Dean and Biesboer (1985) applied NH4+ to the floodwater in cylinders containing
sediment only and in cylinders containing Typha latifoha (broadleaved cattail) Additions
were made biweekly during a single growing season for a total application equivalent to
82 kg nitrogen/ha/season At the end of the growing season, 3 weeks after the last addition,
75 3 % of added N was recovered in the plant-soil system A total of 53 6 % was contained
in the plants, including both above- and belowground biomass, and 21 7% was contained in
the soil In the sediment-only system, only 34 6% of Ihe added N was recovered, most of
this, 33% of the added N, was in the sediment The remaining 65 4% was thought to have
been lost through nitrification-denitnfication reactions
The experiments discussed above indicate that plaint biomass is the major sink for free
NH4 , and that in the absence of plants, the major fate is nitnfication-denitnfication
It should be emphasized that the nitnfication-denitnfication process can dominate only in
environments, like wetlands, that have separate and distinct aerobic and anoxic zones of
microbial activity where solutes freely diffuse between them
10-123
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10.7.4.3 Effects of Nitrogen Loading on Microbial Processes
Changes in deposition rate and the chemical form of nitrogen in deposition can
potentially influence microbial processes and details of the internal nitrogen cycle of
wetlands. For instance, the decomposition rate is sensitive to the nitrogen concentration of
decomposing tissues and of the surrounding environment Tissues with elevated nitrogen
concentrations normally are observed to decompose at a faster rate than tissues containing
low nitrogen concentrations (Marinucci et al, 1983, Neely and Davis, 1985b) The
difference in decomposition rates can be impressive For example, litter from nitrogen-
fertilized Spartma altemiflora decomposed 50% faster than control litter (Mannucci et al,
1983).
The dynamics of nitrogen within decomposing litter is also sensitive to the litter's
nitrogen status That is, litter of low original nitrogen content often acts as a net nitrogen
sink during the first months of decomposition, whereas nitrogen-rich litter is likely to be a
exporter rather than an accumulator during decomposition (Neely and Davis, 1985b) There
is some controversy about the mechanism of nitrogen immobilization (Bosatta and Staaf,
1982; Aber and MeliUo, 1982, Bosatta and Berendse, 1984), but its importance to the
wetland nitrogen cycle is recognized (Brinson, 1977, Morns and Lajtha, 1986, Damman,
1988)
Microbial nitrogen transformations are also affected by the nitrogen status of the
environment. It is well known that NH4+ inhibits the activity of nitrogen-fixing bactena
(diazotrophs) (Buresh et al, 1980) It is thought that NH4+ represses synthesis by bactena
of the nitrogenase enzyme (the enzyme in bactena that accomplishes the transformation)
There may be direct inhibition by NH4+ of enzyme activity, as suggested by Yoch and
Whiting (1986). Kolb and Martin (1988) observed a decrease in nitrogenase activity as well
as the proportion of diazotrophs among the heterotrophic bactena in soil after application of
NH4NO3. They suggested that the decrease in proportion of diazotrophs represents a
competitive suppression by nondiazotrophs in the presence of combined nitrogen (NH4+ or
NO3~). Dicker and Smith (1980) observed a similar repression of nitrogen fixation in salt
marsh sediments amended with either NH4+ or NO3"
Acidification, which may be caused by deposition of NOX or NH4+, can impact the
nitrogen cycle The decomposition rate is decreased by acidification (Leuven and Wolfs,
10-124
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1988, Hendnckson, 1985), but the degree of inhibition is dependent on the buffering capacity
of the litter (Gallagher et al, 1987) Nitrification is also affected by acidification
Nitrification was inhibited at pH 4 to 5 in cypress swamps (Dierberg and Brezomk, 1982),
and at pH 5 4 to 5 7 in lakes (Rudd et al , 1988) Acidification blocks the nitrogen cycle by
inhibiting nitrification and leads to an accumulation of NH4+ (Roelofs, 1986, Schuurkes
et al , 1986, 1987, Rudd et al , 1988) Also, the ratio of N2O N2 produced by denitrifying
bacteria is apparently pH sensitive, with little N2O being produced under anoxic conditions at
pH 7 and almost 100% N2O being produced at pH 5 (Focht, 1974) This is significant
because a shift to N2O production upon acidification of the environment could have a
deletenous effect on stratospheric O3
Finally, NO3" and NH4+ have been shown to influence the relative and absolute
production of end products of dissimilatory nitrate reduction (Blackmer and Bremner, 1978,
Knowles, 1982, Prakasam and Krup, 1982) King and Nedwell (1985) observed
approximately equal reduction to either NH4+ or N2O (in the presence of acetylene, the gas
added to assay the rate of production of N2O) in sediment slurrys incubated anaerobically
with 250 pM NO3" As the nitrate concentration was increased up to 2 mM (1 mM =
14 mg mtrogen/L), the proportion of the nitrate that was denitrified to N2O increased up to
83 % High nitrate concentrations have also been shown to favor N2O production and inhibit
N2 production, perhaps due to the competitive role that exists between NO3" and N2O
terminal electron acceptors during anaerobic respiration (Cho and Sakdinan, 1978, Blackmer
and Bremner, 1978) Seitzinger et al (1983, 1984) observed higher ratios of N2O.N2
production and higher absolute rates of N2O production from eutrophic sediments than from
unpolluted sediments of Narragansett Bay, RI Smith and DeLaune (1983) reported that N2O
production from salt marsh and brackish marsh soils increased from 0 22 and 0 04 mg
2 o
N2O-mtrogen/m /day, respectively, to 1 5 and 2 9 mg N2O-mtrogen/m /day after amending
the sediments with 1 2 to 1 5 g NH4+-nitrogen/m2 Olhers (Betlach and Tiedje, 1981),
however, failed to observe an inhibition of N2O reduction in the presence of NO3" Little is
known about the significance of this process in general or the potential for NO3" or NH4+ in
deposition to alter natural rates of N2O production Only a small fraction of depositional
nitrogen inputs are likely to be evolved as N2O For example, Pedrazzim and Moore (1983)
recovered only 0 39% of fertilizer nitrogen as N2O from submerged soils amended with 34 g
10-125
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f\ _l f\
NO3"-nitrogen/m and 12 g NH4 -nitrogen/m in the laboratory However, on a global
basis, even small changes in the production of N2O are potentially significant considering the
role of N2O in the destruction of stratospheric O3 (Crutzen, 1970, Hahn and Crutzen, 1982)
10.7.4.4 Effects on Biotic Diversity and Ecosystem Structure
In the introduction, it was pointed out that wetlands harbor about 17 % of the total
number of plant species formally listed as endangered in the United States Although it is
beyond the scope of this review to survey the physiological ecology of these wetland plants,
several species on this list are widely recognized to be adapted to nitrogen-poor or infertile
environments. These include the isoetids (Boston, 1986) and the insectivorous plants (Keddy
and Wisheu, 1989, Moore et al, 1989, Wisheu and Keddy, 1989), like the endangered green
pitcher plant, Sarracems oreophila In eastern Canadian wetlands, nationally rare species are
found principally on infertile sites (Moore et al , 1989, Wisheu and Keddy, 1989)
Therefore, management practices should recognize that alterations in competitive
relationships between species occur when the fertility of the environment changes
These assertions are supported by research on flonstic changes related to nitrogen
deposition in central Europe Nitrogen supply is a critical factor in plant nutation in many
natural ecosystems and in agriculture and grassland management as well Ellenberg (1988)
surveyed the nitrogen requirements of 1,805 plant species from West Germany and
concluded that 50% can compete successfully only in habitats that are deficient in nitrogen
supply. Furthermore, of the threatened plants, 75 to 80% are indicator species for habitats
poor in nitrogen supply (i e , they grow only in nitrogen-poor habitats) When stratified by
ecosystem type, it is also clear that the trend of rare species occurring with greater frequency
in nitrogen-poor habitats is a common phenomenon across many ecosystem types
(Figures 10-23 and 10-24)
There is a history in western Europe of changes in wetland community composition that
are thought to result from deposition of atmosphenc pollutants Sphagnum species are
largely absent from ombrotrophic peat bogs in areas of Britain where they were once
common (Talks, 1964, Ferguson et al, 1984, Lee et al , 1986) Ombrotrophic wetlands
downwind of the Manchester and Liverpool conurbations have been extensively modified by
atmospheric pollution for greater than 200 years, with the virtual elimination of the dominant
10-126
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CO
700+
5 7 9
rich
(b)
50
30
10-
A
1 3
poor
«-.. threatened
n = 474
« « non threatened
n - 1274
-»-•--• -1^»
579
rich
(x - 0 20)
C> = 0 35)
Figure 10-23. Distribution of 2,164 Central European plant species on a nitrogen
indicator value gradient from very poor (1), to sufficient (5), to rich (7),
to surplus (9), due in part to nitrogen deposition, (a, c) Species with
unknown preference are indicated with a "?", and those not influenced by
nitrogen supply are indicated with an "x". (b) Most threatened species
can compete only on nitrogen-deficient stands, (c) The fraction of
threatened species diminishes with increasing nitrogen until sufficiency
(5) is reached and then remains constant. In every type of ecosystem,
threatened species are concentrated in the poor to very poor portion of
the nitrogen gradient.
Source Ellenberg (1988)
peat-forming Sphagnum mosses from more than 60,000 ha of bog (Lee et al, 1986) This
has led to a loss of water retention and widespread erosion Nitrogen pollutants from
atmospheric deposition have been implicated in this process, although studies of this
particular area should be interpreted cautiously because of its long history of exposure to
multiple pollutants (Lee et al, 1986) The combination of NO3" and NH4+ deposition, about
32 kg nitrogen/ha/year, is more than double the deposition rates in the Berwyn Mountains in
North Wales, which still support healthy Sphagnum communities, and contributes
10-127
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Wetland and Moorland
Often Mechanically
Disturbed Places
50
40"
30-
A 20-
co 10"
CD
'o
CD
Q.
W en
H- 50
O
•>=>
^ AT\-
1 40
^ 30*
o/\ _
tU
10
threatened n=1 17
non threatened n=1 19
m
Y//
V
^
I
1
1
^
A
•
s
-*.
"••
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•'
A
,«v
>
\
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X
N
i
/
\
I
i
f
^
1
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V-
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i
^
»..,
^
""" {
n=14(
n=241
\
\
>L^
N
1
3 ....
7
N
»
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1 3 5 79.
poor rich
«
<
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I
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i
&
'/^//^
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S»-
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n=10:
n=28(
-•^
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j
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\
\
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/•
-•-
s
n=52
n=34
A
'-•
\
•»x
i
N
V
N
3 5 79,
rich
Healthland and Grassland
Woodland and Bush
Figure 10-24. Distribution of Central European plant species along a gradient of
nitrogen indicator values (see Figure 10-23) across ecosystem types.
In every analyzable type of ecosystem, threatened plant species are
concentrated in the poor (1) to very poor portion of the gradient.
Source: Ellenberg (1988)
significantly to a supraoptimal nitrogen supply (Lee et al, 1986) In the Netherlands, there
has been a great decline during the past three decades in communities dominated by wsetids
in soft water areas and their conversion to later successional stages dominated by grasslands
or by Juncus bulbosus (rush) and Sphagnum spp (Roelofs, 1983, 1986, Roelofs et al, 1984,
Schuurkes et al., 1986)
10-128
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Vermeer and Berendse (1983) correlated biomass with species numbers and soil
chemical characteristics in several fen and grassland communities in the Netherlands
In fens, they found a negative correlation between biomass and NH4+ concentration and a
positive correlation between biomass and pH There was also a positive correlation between
biomass and number of species In wet grasslands, a positive correlation was found between
biomass and NO3", Px, and K+ In all wetland types investigated, they report that species
number was greatest when the standing biomass of the site was in the range of 400 to
500 g/m2 They concluded that domination by a few species is associated with eutrophic
conditions at the high end of the biomass scale as well as with conditions unfavorable for
growth at the low end of the scale Similarly, in wetlands of eastern Ontario and western
Quebec, the greatest diversity of species (3 to 24 per 0 25 m ) occurs at intermediate
9 2
standing crops (60 to 500 g/m ) and the lowest density of species (2 to 5 per 0 25 m ) at
standing crops greater than 1,500 g/m2 (Moore and Kecldy, 1989, Wisheu and Keddy, 1989)
In Great Britain, species density in fens was greatest (about 12 per 0 25 m2) at standing
22 2.
crops less than 1,000 g/m and lowest (3 per 0 25 m ) when standing crop was 4,000 g/m
or greater (Wheeler and Giller, 1982) Exceptions to tins trend are found where annual
mowing and harvest of wetland vegetation minimize the accumulation of surface litter
(Verhoeven et al, 1988), and possibly where intense pressure from grazing animals favors
domination by specific plant species (Jensen, 1985, Berendse, 1985)
10.7.4.5 Mechanisms of Nitrogen Control Over Ecosystem Structure
Nitrogen supplied m excess of a plant's nutritional requirements has a direct toxic effect
on some species The concentrations of six elements in the tissues of five Sphagnum species
have been investigated in relationship to atmospheric deposition in Europe (Ferguson et al,
1984) When Sphagnum species were transplanted from a relatively clean-air site to a
polluted site, the concentrations of nitrogen, sulfur, lead, Fe, and phosphate mcreased
significantly, but the concentration of potassium did not The greatest change observed was
for nitrogen, which mcreased by absolute amounts that varied from 17 7 mg/g of tissue in
Sphagnum recurvum to 5 3 mg/g in Sphagnum capilhfoimm above control levels of about
10 mg/g (1 % of dry weight) Because the nitrogen supply originating from the soil probably
did not differ, as indicated by the similarity m total nitrogen concentration of the peat from
10-129
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the polluted and clean sites, it is possible that nitrogen deposition had a direct effect on
nitrogen uptake in these species The authors concluded that the element supply from
deposition at the polluted site, where nitrogen deposition is 43 kg nitrogen/ha/year, is
supraoptimal for growth of ombrotrophic Sphagnum species They noted the existence of a
"good" Sphagnum cover at one site where a nitrogen deposition rate of 20 kg
nitrogen/ha/year was measured Similarly, Press et al (1986) observed tissue nitrogen
concentrations as high as 2 5 % of dry weight in Sphagnum cuspidatum transplanted to a site
of high nitrogen deposition in northern Britain and found that this level of nitrogen was
associated with decreased growth
Competitive relationships among species change with the nitrogen status of the
environment In weakly buffered ecosystems, a high deposition of NH4+ leads to
acidification and nitrogen enrichment of soil Consequently, plant species characteristic of
poorly buffered environments disappear Among the acid-tolerant species, there will be
competition between slow-growing and fast-growing nitrophilous grasses or grass-like
species This process contributes to the observed change from heathlands into grasslands
Molima caerulea (L) Moench and/or Deschampsia flexuosa (L ) Tnn (grasses) expand at
the expense of Enca tetralvc or Calluna vulgans (L) Hall (shrubs) and other heathland
species (Berendse and Aerts, 1984, Roelofs et al, 1987, Aerts and Berendse, 1988, 1989)
In over 70 heathlands investigated, the shrub bogs dominated by Enca tetralvc or Calluna
had dissolved NH4+ levels in the soil water of 55 and 84 /xM, whereas those dominated by
the grasses Deschampsia and Molima had average NH4+ concentrations of 248 and 429 pM
(Roelofs etal., 1987)
Several controlled-growth studies also have been conducted to identify the mechanisms
of nitrogen control over species composition This is a nontnvial task because there are a
great number of interactions among biochemical and geochemical processes There are
direct and indirect effects of nitrogen deposition, and cause and effect can be difficult to
ascertain. Roelofs (1986), for example, states that acidification, which can result from
fy i
deposition of NOX, SO4 ", or NH4 , can decrease the availability of dissolved CO2 in water,
which leads to the complete elimination of submerged plant species Deposition of NH^
and its subsequent mtiification or absorption by plants generates acidity Biochemical
2
conversions of SO4 and NO3 generate alkalinity These processes are mediated by
10-130
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bacteria, macrophytes, and algae (Kelly et al , 1982, Raven, 1985) Atmospheric deposition
of nitrogen can significantly affect the nitrogen budget of some wetland ecosystems, their
acidity, and their carbon budgets (Roelofs, 1986)
Schuurkes et al (1986) studied effects of acidification and nitrogen supply on growth of
several common wetland plants under controlled laboratory conditions All species utilized
NH4+ and NO3" as a nitrogen source, except Sphagnum flexuosum, which did not assimilate
NO3" When NH4+ and NO3" were offered simultaneously in equal amounts, NO3" uptake
was the dominant form of nutation (63 to 73%) in plants that are characteristic of soft waters
(low Ca + and Mg +), whereas NH4+ strongly dominated the nutation (85 to 90%) in
species from acid waters Differences in the site of uptake, either leaves or roots, among
_i_ 9
species were also found They concluded that high deposition of NH4 and SO4 ", the most
important sources of acidification in the Netherlands, is leading to an expansion of acid-
tolerant mtrophilous plants
The nutation of Sphagnum is apparently species specific Although S flexuosum did
not assimilate NO3" (Schuurkes et al , 1986), the activity of nitrate reductase in
S cuspidatum (Press and Lee, 1982) and in S fuscum (Woodin et al , 1985) clearly shows
that NO3" can be utilized by these species 5 magellamcum was shown to grow best when
given the equivalent of 4 1 kg NO3 "-nitrogen/ha/year plus 19 kg NH4+ -nitrogen/ha/year in
simulated rain, when given 0 25 tunes that amount of NO3" and 1 5 tunes (and 4 tunes) as
much NH4+, growth decreased (Rudolph and Voigt, 1986) Bayley et al (1987) reported
that the dominant Sphagnum spp in a poor fen in Ontario, S fuscum and S magellamcum,
were able to assimilate an NO3" input of 4 71 kg nitrogen/ha/year, including 1 6 kg
nitrogen/ha/year applied in simulated acid rain, and growth increased at least during the first
year after the additional nitrogen was applied Roelofs et al (1984) observed that growth of
S cuspidatum was greatest in a medium containing 500 /xM NH4+, and growth was less at
1,000 or 100 /iM NH4+ Press et al (1986) observed that the best growth of this same
species occurred in nitrogen-free solutions, and that even small additions (10 jwM) of either
NH4+ or NO3" reduced growth It is doubtless that some variations in results of nutritional
studies are influenced by other variables, like pH
The genus Sphagnum is an important group in bogs everywhere, and it is important to
understand its nutritional physiology and ecology However, it should be emphasized that
10-131
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the consequences of nitrogen fertilization in a natural environment, with fluctuating climate
and competition among numerous species, can be quite different from what may be predicted
from studies of a single species in laboratory culture For example, Aerts et al (1989) assert
that competition for light dictates the outcome of competition between species that differ in
growth rate potential and nutrient requirement
In a 2-year greenhouse experiment designed to differentiate between acid and nitrogen
effects, Schuurkes et al (1987) exposed mixtures of different wetland plant species to
2. 4-
simulated rain containing various combinations of SO4 ", NH4 , and NO3" Marked changes
were observed in systems receiving rain with 510 and 1,585 juM NH4+, plants typical of
nutrient-poor soft waters (like the isoetids Littorella uniflora [shoreweed], Luromum natans
[water plantain], and Pilulana globuhferd) were adversely affected at this level of nitrogen
input, whereas other species (Juncus bulbosus, Sphagnum cuspidatum, and the grass Agrostts
canmd) expanded Acidification with none or only a small NH4+ addition had no clear
effects, although biomass of Sphagnum was slightly higher Within sulfuric acid treatments,
only pH 3.5 rain markedly acidified the water Based on these experiments, Schuurkes et al
(1987) recommended that to preserve the remaining oligotrophic wetlands, acid inputs should
not exceed 250 mol/ha/year, and that the annual nitrogen deposition should not be greater
than 1,380 mol/ha/year or 19 4 kg nitrogen/ha/year (NO3" + NH4+), except that the
potential acidifying influence of this nitrogen input, if in the form of NE^"1", exceeds the
allowable acid input This limit is supported by Liljelund and Torstensson (1988), who
concluded from their review that the Limit for many species may be well below 20 kg
nitrogen/ha/year and for oligotrophic (nutrient-poor) bogs, is probably about 10 kg
nitrogen/ha/year. These limits are exceeded currently in the United States, where wet nitrate
deposition alone exceeds 15 kg nitrogen/ha/year over most of the Midwest, New York, and
New England (Zemba et al, 1988) The effects of the nitrate deposition, however, are yet
to be determined
10-132
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10.8 AQUATIC EFFECTS OF NITROGEN OXIDES
10.8.1 Introduction
For a variety of reasons, nitrogen deposition has not historically been considered a
serious threat to the integrity of aquatic systems Most terrestrial systems have been assumed
to retain nitrogen strongly, leading to a small probability that deposited nitrogen will ever
make its way to the surface waters that drain these terrestrial systems Nitrogen within
aquatic ecosystems can arise from a variety of sources, including point-source and
non-point-source pollution and biological fixation of gaseous nitrogen, in addition to the
deposition of NOX In cases where nitrogen is known to be affecting aquatic systems, it has
been assumed that some source other than deposition is responsible The amounts of
nitrogen provided to aquatic systems by these other sources often outweigh by a large margin
the amount of nitrogen potentially provided by atmospheric deposition In the past decade,
however, our understanding of the transformations that nitrogen undergoes within watersheds
has increased greatly, and in areas of the country where nonatmosphenc sources of nitrogen
are small, we can begin to infer cases where nitrogen deposition is having an impact on
aquatic systems
Estimating the effects of NOX emissions and nitrogen deposition on aquatic systems is
made difficult by the large variety of nitrogen compounds found in air, deposition,
watersheds, and surface waters, as well as the myriad of pathways through which nitrogen
can be cycled in terrestrial and aquatic ecosystems These complexities have the effect of
decoupling nitrogen deposition from nitrogen effects, and reduce our ability to attribute
known aquatic effects to known rates of nitrogen deposition The organization of this section
reflects this complexity Because an understanding of the ways that nitrogen is cycled
through watersheds is critical to our understanding of nitrogen effects, the section begins
with a brief description of the nitrogen cycle, and of the transformations of nitrogen that may
occur in watersheds Each of the known possible effects of nitrogen deposition
(acidification, eutrophication, and direct toxicity) is discussed separately Within these
discussions, evidence for the importance of nitrogen in causing observed effects is discussed
separately from evidence that deposition is the source of the nitrogen observed in affected
systems
10-133
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10.8.2 The Nitrogen Cycle
Atmospheric nitrogen can enter aquatic systems either as direct deposition to water
surfaces, or as nitrogen deposition to the terrestrial portions of a watershed (Figure 10-25,
see also Figure 10-1) Nitrogen deposited to the watershed is then routed (e g , through
plant biomass and soil microorganisms) and transformed (e g , into other inorganic or
organic nitrogen species) by watershed processes, and may eventually run off into aquatic
systems in forms that are only indirectly related to the original deposition Much of the
challenge of determining when nitrogen deposition is having an effect on aquatic systems
depends on our ability to track nitrogen on its path through watersheds In most cases, this
tracking cannot be accomplished outside of a carefully controlled research program, and we
are forced to make educated guesses about the likelihood that the nitrogen observed in
aquatic systems was originally of atmospheric origin The strength of these educated guesses
will depend, to a large degree, on our ability to identify which nitrogen transformations are
occurring and which are not By eliminating other possible sources or sinks of nitrogen, we
are in a stronger position to determine in which cases observed nitrogen effects are caused
indirectly by atmospheric deposition Our understanding of the nitrogen cycle in terrestrial
and aquatic ecosystems, therefore, plays a central role in controlling our understanding of
deposition effects The key elements of the nitrogen cycle, particularly those that are thought
to be important in determining whether atmospherically derived nitrogen will have an effect
on aquatic systems, are discussed briefly in this section (see also Section 10 3)
10.8.2.1 Nitrogen Inputs
Watersheds are generally several orders of magnitude larger than the surface waters that
drain them, and so the majority of the atmospheric deposition that may potentially enter
aquatic systems falls first on some portion of the watershed Nitrogen may be deposited to
the watershed, or directly to water surfaces, in a vanety of forms, including NO3", NH4+,
and organic nitrogen in wet and dry deposition In addition, plants may absorb gaseous
nitrogen as NOX (Rowland et al, 1985) or HNO3 vapor (Vose et al, 1989), and nitrogen
thus absorbed may subsequently enter the watershed nitrogen budget as litter fall, or through
the death of plant biomass (Parker, 1983, Olson et al , 1985) These nitrogen constituents
10-134
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Deposition
Wet
NC
I>3 NH4
Dry
r>
JOX NHX
NO
_ .
assimilation
i 1
NHj
nitrification
a
/ \
- ass
3
^
Plant
»•— ^^—»
Biomass
"Vs.
ssimilation'X
nineralization
imilation /
Microbi
Biomaj-
al
.s
w
*s •
Dead
Organic
Matter
/
nitrogen
^
>
fixation
•^•i
1 t
I
Leaching
Water
denitrification
Figure 10-25. A simplified watershed nitrogen cycle. Only the major pathways are
shown. The boxes represent major pools of nitrogen in terrestrial
ecosystems, and the lines represent the major pathways and processes
affecting nitrogen transformations. The wavy line represents the soil
surface.
Source Skeffington and Wilson (1988)
are the same as those comprising direct deposition to terrestrial ecosystems recently described
by Lindberg et al (1986) (also see Section 10 6)
Concentrations of NO3" and NH4+ in precipitation vary widely throughout North
America, depending largely on the proximity of sampling sites to sources of emissions
Galloway et al (1982) report mean concentrations of NO3" and NH4+ of 2 4 /xeq/L and
2 8 jweq/L, respectively, for a site in central Alaska In the Sierra Nevada Mountains of
10-135
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California, mean concentrations of NO3" and NH4+ for the period 1985 to 1987 were
5.0 and 5 4 jieq/L, respectively (Williams and Melack, 1991a) In a comparison of nitrogen
deposition at lake and watershed monitoring sites in the northern United States and southern
Canada, Linsey et al. (1987), found NO3" concentrations ranging from 15 to 40 /*eq/L and
NE^ concentrations from 10 to 50 /ieq/L in areas considered remote but influenced by
prairie dust and long-range acidic deposition, neither ion dominated over the other In some
areas closer to anthropogenic nitrogen sources (e g , in northeastern United States and
southeastern Canada), volume-weighted mean NO3" concentrations range from 30 jweq/L
(e.g., in the Adirondack and Catskdl mountains of New York) to 50 /weq/L (e g , in the
eastern Great Lakes region), whereas mean NH4+ concentrations range from 10 to 20 j«eq/L
in the same areas (Stensland et al, 1986) Ammonium concentrations are highest
(ca. 40 juieq/L) in the agricultural areas of the midwestern United States
Deposition of nitrogen will depend on the concentration in precipitation, the volume of
water falling as precipitation, and the amount of nitrogen in dry deposition (see
Section 10 4 of this report, see also Sisterson et al, 1990) The last of these values (dry
deposition) is difficult to measure, and is often estimated as a fraction (e g , 30 to 40%) of
wet deposition (Baker, 1991) Given the range of concentrations mentioned in the previous
paragraph, and the volumes of precipitation falling in different regions of North America,
estimates of nitrogen deposition rates range from less than 0 2 kg/ha/year in Alaska to
12 kg/ha/year in the northeastern United States (Table 10-19)
Generally NO3" dominates over NH4+ at sites close to emission sources (Linsey et al,
1987, Altwicker et al, 1986) Dissolved organic nitrogen concentrations are highly variable
in precipitation, but often amount to 25 to 50 % of inorganic nitrogen deposition values
(Linsey et al., 1987, Manny and Owens, 1983, Feller, 1987)
10.8.2.2 Transformations
Because the majority of nitrogen deposition falls first on some portion of the watershed,
the transformations that nitrogen undergoes within the watershed (e g , in soils, by microbial
action, and in plants) will play a major role in determining what forms and amounts of
nitrogen eventually reach surface waters Much of the following discussion is, therefore,
focused on terrestrial processes that alter the forms and rates of nitrogen supply It is these
10-136
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processes that, to a large degree, determine whether nitrogen deposition will ever reach
lakes, streams, and estuaries, and, therefore, they are very important in controlling the
effects of nitrogen deposition Many of these same processes occur also within surface
waters, and a specific discussion of these processes, and therr importance, follows the
discussion of nitrogen transformations
Nitrogen Assimilation
Nitrogen assimilation is the uptake and metabolic use of nitrogen by plants
(Figure 10-25) Assimilation by both terrestrial and aquatic plants will play a role in
determining whether nitrogen deposition affects aquatic systems Assimilation by the
terrestrial ecosystem controls the form of nitrogen eventually released into surface waters, as
well as affecting the acid/base status of soil and surface waters Terrestrial assimilation is a
major form of nitrogen removal in watersheds, and may in fact be sufficient to prevent all
atmospherically-derived nitrogen from reaching surface waters (Vitousek and Reiners, 1975)
Nitrogen is the most commonly limiting nutrient in forest ecosystems in North America
(Cole and Rapp, 1981) Because the primary use of nitrogen in plant biomass is the
formation of amino acids, and reduced nitrogen is the most energetically favorable form of
nitrogen for incorporation into amino acids, uptake of NH4+ is generally favored over uptake
of NO3" by terrestrial plant species This demand for NH4+ over NO3~ and the high cation
exchange capacity, typical of most temperate forest soils, combine to create the common
pattern of low NH4+ concentrations in waters draining forested watersheds in the United
States The form of nitrogen used by terrestrial ecosystems strongly affects the acidifying
potential of nitrogen deposition (Figure 10-26) Ammonium uptake is an acidifying process
(i e , uptake of NH4+ releases one mole of hydrogen per mole of nitrogen assimilated)
NH4+ + R OH = R NH2 + H2O + H+ (10-10)
The biological uptake of NO3", on the other hand, is an alkalinrzing process (i e , uptake of
NO3" consumes one mole of hydrogen per mole of mtiogen assimilated).
10-137
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In
In
In
In
NC
-1 >
>3 W
' >
Organic
Matter
N(
r\t
+1
• >
^3 N
it ^\
^1+1 N°3 -1
-1
MLJ+ +1
_J NH4
N
i
Decomposition
f
3 3 NH4 Deposition, Fertilizers
1 +1
w Denitrrfiers -> M_ N«O C
Plnnt" " •-•Ł>•-Ł
Figures Represent H+
Transfers to Soil or Water
ut
-1+ Leaching
Figure 10-26. The effect of nitrogen transformations on the watershed hydrogen ion
budget. One hydrogen ion is transferred to the soil solution or surface
water (+1) or from the soil solution or surface water (-1) for every
molecule of nitrate or ammonium that crosses a compartment boundary.
For example, nitrification follows the pathway for ammonium uptake into
organic matter (+1), and is leached out as nitrate (+1), for a total
hydrogen ion production of +2 for every molecule of nitrate produced.
Source Skeffington and Wilson (1988)
R OH + NO3" + H+ = R NH2 + 2O2
(10-11)
Nitrification
Nitrification is the oxidation of NH4+ to NO3", and is mediated by bacteria and fungi
in both the terrestrial and aquatic portions of watersheds It is an important process in
controlling the form of nitrogen released to surface waters by watersheds, as well as in
controlling the acid/base status of surface waters (Figure 10-25) Nitrification is a strongly
acidifying process, producing two moles of hydrogen for each mole of nitrogen (NH4+)
nitrified (Figure 10-26):
2O2 = NO3" + 2HH
H2O
(10-12)
10-138
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Because nitrification in forest soils commonly transforms NH4+ into NO3~, the acidifying
potential of deposition (the maximum potential for acidification that is attributable to
nitrogen) is often defined as the sum of NH4+ and NO5", assuming that all nitrogen will
leave the watershed as NO3" (e g , Hauhs et al , 1989)
In most soils, nitrification is limited by the supply of NH4+ (Likens et al , 1970,
Vitousek et al , 1979), creating a high demand for NH4+ on the part of nitrifying soil
microbes This microbial demand for NH4+, coupled with the demand for NH4+ on the
part of terrestrial plants (discussed above), leads to surface water concentrations of NH4+
that are almost always unmeasurable Nitrification rates may also be limited by inadequate
microbial populations, lack of water, allelopathic effects (toxic effects produced by inhibitors
manufactured by vegetation), or by low soil pH Of these other potential limiting factors,
soil pH plays an obviously vital role in any discussion of the acidification of surface waters
by nitrogen deposition Nitrification has traditionally been thought of as an acid-sensitive
process (Dnscoll and Schaefer, 1989, Aber et al , 1989), but high rates of nitrification have
been reported from very acid soils (i e , pH < 4 0) in the northeastern United States
(Vitousek et al , 1979, Novick et al , 1984, Rascher et al , 1987) and in Europe
(Van Breemen et al , 1982) In the southeastern United States, Montagnim et al (1989)
were unable to find any effect of pH on nitrification, or to stimulate nitrification by buffering
acid soils In a survey of sites across the northeastern United States, McNulty et al (1990)
found no correlation between nitrification rates and soil pH, but found a strong association
(r2 = 0 77) with rates of nitrogen deposition The weight of evidence suggests that
nitrification will proceed at low soil pH values as long as the supply of NB^+ is sufficient
Demtrification
Demtnfication is the biological reduction of NO3 to produce gaseous forms of reduced
nitrogen (N2, NO, or N2O) (Payne, 1981) Demtnfication is an anaerobic process (i e , it
proceeds only in environments where oxygen is absent) whose end product is lost to the
atmosphere (Figure 10-22) In terrestrial ecosystems, denitnfication occurs in anaerobic
soils, especially boggy, poorly drained soils, and has traditionally been considered a
relatively unimportant process outside of wetlands (Post et al , 1985) It has been suggested,
however, that denitnfication could be an episodic process, occurring after such events as
10-139
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spring snow melt and heavy rain storms, when soil oxygen tension is reduced (Melillo et al ,
1983). No single equation can describe the denitnrlcation reaction, because several end
products are possible However, denitnfication is always an alkalimzing process, consuming
one mole of hydrogen for every mole of nitrogen denitrified (Figure 10-26) Denitnfication
can be involved in the production or consumption of N2O, a product that may have
considerable significance as a greenhouse gas (Matson and Vitousek, 1990, Hahn and
Crutzen, 1982) In a review of the effects of acidic deposition on denitnfication in forest
soils, KHemedtsson and Svensson (1988) conclude that denitnfication rates are often limited
by the availability of anerobic soil zones, and may, therefore, be relatively insensitive to
increases in nitrogen deposition It has been suggested that the production of N2O may
increase in acidified soils (Knowles, 1982), but few field data are available to test this idea
Rates of N2O production in soil waters have been shown to increase markedly after forest
clear-cutting (Bowden and Bormann, 1986, Melillo et al , 1983), and in areas of both high
nitrogen deposition and intensive forest management, N2O production may be of concern
Nitrous oxide production is strongly influenced by soil temperature, soil NO3" concentration,
and soil moisture, Davidson and Swank (1990) suggest that one or more of these factors may
commonly limit N2O production in natural systems
Nitrogen Fixation
Gaseous atmospheric nitrogen (N^ can be fixed to produce NH4+ by a wide range of
single-celled organisms, including blue-green algae (cyanobactena), and various aerobic and
anaerobic bacteria Symbiotic nitrogen-fixing nodules are present on the roots of some early
successional forest species (Boring et al , 1988) In headwater streams, nodules on rooting
structures of ripanan vegetation (e g , Alnus sp ) can also be important nitrogen fixers
(Binkley, 1986). Ordinarily, nitrogen fixation has no direct effect on the acid/base status of
soil or surface waters
N2 + H2O + 2R OH = 2R NH2 + 3/2O2 (10-13)
10-140
-------
Nitrogen fixation in excess of biological demand, however, can lead to nitrification or
mineralization of organic nitrogen, and, ultimately, lead to acidification of soil or surface
waters (Franklin et al , 1968, Van Miegroet and Cole, 1985)
Mineralization
Mineralization is the bacterial decomposition of organic matter, releasing NH4+ that
can subsequently be nitrified to NO3" Mineralization is an important process in watersheds,
as it recycles nitrogen that would otherwise be lost from the system through death of plants,
or as leaf litter (Figure 10-22) In a comparative study of mineralization in soils,
Nadelhoffer et al (1985) found nitrogen mineralization rates ranging from 50 to
100 kg/ha/year under deciduous tree species, and from 32 to 66 kg/ha/year under coniferous
species These rates should be compared to nitrogen deposition rates of 5 to 12 kg/ha/year
for high deposition areas of the Northeast Nadelhoffer et al (1985) also report estimated
rates of nitrogen uptake that were 5 to 20% higher than rates of mineralization, suggesting
that mineralization can supply the majority, but not all, of the nitrogen needed for plant
growth in these forests
The effect of mineralization on the acid/base status of draining waters will depend on
the form of nitrogen produced The conversion of organic nitrogen (e g , from leaf litter) to
NH4+ consumes 1 mole of hydrogen per mole of nitrogen produced (Figure 10-26), and can
be thought of as the reverse of the reaction in Equation 10-10 Organic nitrogen, which is
mineralized and subsequently oxidized (nitrified) to NO3~ (Equation 10-12), produces a net of
1 mole of hydrogen per mole of NO3" produced Because the production of organic nitrogen
(i e , assimilation) can either produce or consume hydrogen (depending on whether NO3" or
NH4+ is assimilated), the net (ecosystem) effect of mineralization depends both on the
species entering the watershed and on the species leaving the watershed (Figure 10-26)
In ecosystems where plant growth is limited by the availability of nitrogen,
mineralization is also limited by nitrogen, in the sense that additions of nitrogen to the leaf
litter will speed decay and increase the rate at which nitrogen is immobilized by decomposers
(Mekllo et al , 1984, Taylor et al, 1989) Nitrogen limitation of decomposition is in part
due to the low nitrogen content typical of litter, resulting from the retranslocation of nitrogen
out of leaves during senescence
10-141
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10.8.2.3 Nitrogen Saturation
Much of the debate over whether aquatic systems are being affected by nitrogen
deposition centers on the concept of nitrogen saturation of forested watersheds Nitrogen
saturation can be defined as a situation where the supply of nitrogenous compounds from the
atmosphere exceeds the demand for these compounds on the part of watershed plants and
microbes (Aber et al, 1989, Skeffington and Wilson, 1988) Under conditions of nitrogen
saturation, forested watersheds that previously retained nearly all of nitrogen inputs, due to a
high demand for nitrogen by plants and microbes, begin to have higher loss rates of nitrogen
These losses may be m the form of leaching to surface waters or to the atmosphere through
denitnfication These two potential loss pathways have profoundly different impacts on the
acid/base status of watersheds and surface waters (see following discussion), and their
relative importance in advanced stages of nitrogen saturation will be a decisive characteristic
determining the seventy of the impact of nitrogen saturation
Aber et al. (1989) have proposed a hypothetical tune course for a watershed response to
chronic nitrogen additions (Figure 10-27), describing both the changes in nitrogen cycling
that are proposed to occur, as well as the plant responses to changing levels of
nitrogen availability Aber et al (1989) include in their hypothetical time course a trajectory
for the loss of nitrogen to surface water runoff (Figure 10-27), which suggests a simple
response (nitrogen leaching) in the later stages of nitrogen saturation One of the objectives
of this document is to establish whether stages equivalent to those shown in Figure 10-27 can
be described for surface waters, and to determine whether the response of surface waters to
advanced stages of nitrogen saturation is as simple as suggested in Figure 10-27
Stage 0 of the Aber et al (1989) conceptual model is the pretreatment condition, where
inputs of nitrogen from deposition are at background levels and watershed losses of nitrogen
are negligible (Figure 10-27) In Stage 1, increased deposition is occurring, but effects on
the terrestrial ecosystem are not evident For a limiting nutnent such as nitrogen, a
fertilization effect might result in increased ecosystem production and tree vigor at Stage 1
Retention of nitrogen is very efficient, and, on an annual basis, little or no nitrogen would be
lost to surface waters that drain Stage 1 watersheds Many forested watersheds in the United
States would be considered to exist at this stage
10-142
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N Mineralization
c
^
CD
Nitrification
N Inputs
CD
DC
0
Staq<
}/^«<
1 > x N2C
J^s'L^-^ Emiss
i i
Additions Satuiation Decline
Begin
90 1 23
I
0)
cc.
0
NPP
Foliar Biomass
Foliar N Concentration
Fine Root Mass
Nitrate Assimilation
^s
Stage
Additions
Begin
0
I I
Saturation Decline
1
Figure 10-27. Hypothetical time course of forest ecosystem response to chronic nitrogen
additions—top: relative changes in rates of nitrogen cycling and nitrogen
loss, bottom: relative changes in plant condition (e.g., foliar biomass and
nitrogen content, fine root biomass) and function (e.g., net primary
productivity and nitrate assimilation) in response to changing levels of
nitrogen availability.
Aberetal (1989)
10-143
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In Stage 2 of the Aber et al (1989) hypothetical tune course, negative effects occur,
but they are subtle, nonvisual, and/or require long time scales to detect Only in Stage 3 do
visible effects on the forests occur, resulting in major environmental impacts Aber et al
(1989) emphasize that different species and environmental conditions could alter the tuning of
effects illustrated in Figure 10-27
A number of factors may contribute to a watershed's progression through the stages of
nitrogen loss, including elevated nitrogen deposition, stand age, and high soil nitrogen pools
High rates of nitrogen deposition play a clear role, as the ability of forest biomass to
accumulate nitrogen must be finite At very high, long-term rates of nitrogen deposition, the
ability of forests and soils to accumulate nitrogen will be exceeded, and the only remaining
pathway for loss of nitrogen (other than runoff) is denitnfication As mentioned earlier, high
rates of nitrogen deposition may favor increased rates of denitnfication, but many watersheds
lack the conditions necessary for substantial denitnfication (e g , low oxygen tension, high
soil moisture, temperature) Another important factor in nitrogen loss from watersheds is the
age of the forest stands A loss in the ability to retain nitrogen is a natural outcome of forest
maturation, as demand for nitrogen on the part of more slowly growing tree species may
plateau in later stages of forest development or decline as forests achieve a "shifting-mosaic
steady state" (Bormann and Likens, 1979) Uptake rates of nitrogen into vegetation are
generally maximal around the tune of canopy closure for conifers, and somewhat later (and
at higher rates) in deciduous forests due to the annual replacement of canopy foliage in these
ecosystems (Turner et al, 1990) Large soil nitrogen pools imply that soil microbial
processes that are ordinarily limited by the availability of nitrogen are instead limited by
some other factor (e g , availability of labile organic carbon), and large soil nitrogen pools
contribute to the likelihood that watersheds will leach NO3" (Johnson, 1992, Joslin et al ,
1992). Nitrogen saturation can be seen to occur in a sequence beginning with the fulfillment
of vegetation nitrogen demand, followed by the fulfillment of soil microbial nitrogen
demand; the existence of large soil nitrogen pools suggests that the second of these
requirements may be easily met. The possible importance of all three factors (deposition,
stand age, and soil nitrogen) in shifting watersheds from one stage of nitrogen loss to another
will be discussed later in the context of surface water evidence of watershed nitrogen
saturation
10-144
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The loss of nitrogen from watersheds can also be seen to occur in stages, which
correspond to the stages of terrestrial nitrogen saturation described by Aber et al (1989)
The most obvious characteristics of these stages of nitrogen loss are changes in the seasonal
and long-term patterns of surface water NO3" concentrations, which reflect the changes in
nitrogen cycling that are occurring in the watershed The nitrogen cycle at Stage 0 is
dominated by forest and microbial uptake, and the demand for nitrogen has a strong
influence on the seasonal NO3" pattern of receiving waters The "normal" seasonal NO3"
pattern in a stream draining a watershed at Stage 0 would be one of very low, or
immeasurable, concentrations during most of the year, and of measurable concentrations only
during snowmelt (in areas where snow packs accumulate over the winter months), or during
spring rain storms The small loss of NO3" during the dormant season is a transient
phenomenon, and results because snowmelt and spring rains commonly occur in these
environments before substantial forest and microbial growth begin in the spring (e g , winter
mineralization of soil organic nitrogen may be an exception to this inactivity [Foster et al,
1989]) As a result, some of the nitrogen stored in soils and/or snowpack may pass through
the watershed during extreme hydrologic events and may result in a pulse of elevated NO3"
concentration The key surface water characteristics of Stage 0 watersheds are very low
NO3" concentrations during most of the year, and maximum spring concentrations of NO3"
that are smaller than concentrations typical of deposition
At Stage 1, the seasonal pattern typical of Stage 0 watersheds is amplified It has been
suggested that this amplification of the seasonal NO3" signal may be the first sign that
watersheds are proceeding toward the chronic stages (i e , Stages 2 and 3 in Figure 10-27) of
nitrogen saturation (Dnscoll and Schaefer, 1989, Stoddzird and Murdoch, 1991), and this
suggestion is consistent with the changes in nitrogen cycling that are thought to occur at
Stage 1 A conceptual understanding of these changes derives from the most common
definition of nutrient limitation Implicit in the definition of nutrient limitation is the idea
that "the current supply rate (of a nutrient) prevents the vegetation from achieving maximum
growth rates attainable wthm other environmental constraints" (emphasis added [Binkley
et al , 1989]) During the cold season, these environmental constraints can be severe, and
maximum attainable growth rates are clearly much lower than m the warm months Much of
this discussion is couched in terms of forest trees, but the same arguments also apply to soil
10-145
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microbial communities (e g , decomposers, nitafiers), which may be as important as
vegetation in controlling nitrogen loss from watersheds (Binkley et al, 1989)
Overall limitation of forest growth (in the early stages of nitrogen saturation) is
characterized by a seasonal cycle of limitations by physical factors (e g , cold and diminished
light during late fall and winter) and nutrients (primarily nitrogen, during the growing
season). The effect of increasing the nitrogen supply (e g , from deposition) is to postpone
the seasonal switch from physical to nutrient limitation during the breaking of dormancy in
the spring, and to prolong the seasonal nitrogen saturation that is characteristic of watersheds
at this stage. At Stage 1, this switch is enough delayed that substantial NO3" may leave the
watershed during extreme hydrologic events in the spring Watershed loss of nitrogen at
Stage 1 is still a seasonal phenomenon, and the annual nitrogen cycle is still dominated by
uptake, but NO3" leaching is less transient than at Stage 0 The key characteristics of Stage
1 watersheds are episodes of surface water NO3" that exceed concentrations typical of
deposition (e.g , Figure 10-28) Elevated NO3" during episodes may result from preferential
elution of anions from melting snow (Jeffries, 1990, Johannessen and Hennksen, 1978) or
from the contribution of nitrogen mineralization to the soil pool of NO3" that may be flushed
during high-flow periods (Rascher et al, 1987, Schaefer and Dnscoll, in press)
In Stage 2 of watershed nitrogen loss, the seasonal onset of nitrogen limitation is even
further delayed, with the effect that biological demand exerts no control over winter and
spring nitrogen concentrations, and the period of nitrogen limitation during the growing
season is much reduced. The annual nitrogen cycle, which was dominated by uptake at
Stages 0 and 1, is instead dominated by nitrogen loss (through leaching and demtnfication) at
Stage 2, sources of nitrogen (deposition and mineralization) outweigh nitrogen sinks (uptake)
The same mechanisms that produce episodes of high NO3" during extreme hydrologic events
at Stage 1 also operate at Stage 2 But more importantly, NO3" leaching can also occur at
Stage 2 during periods when the hydrologic cycle is characterized by deeper percolation
If biological demand is sufficiently depressed during the growing season, nitrogen begins to
percolate below the rooting zone, and elevated groundwater concentrations of NO3" result
Nitrification becomes an important process at Stage 2 (Aber et al, 1989, Figure 10-27),
lowered biological demand leads to a buildup of NH4 in soils, and nitrification may be
stimulated. This is a pivotal change in the nitrogen cycle because nitrification is such
10-146
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I o-Total
• - Monomerlc
A- Nonlobile
Jan
May Sept Jan
1984
May Sept
1985
Jan
May
1986
Time(months)
Figure 10-28. Temporal patterns in the chemical characteristics of stream water at
Pancake-Hall Creek in the Adirondaclks. Sulfate and base cation
concentrations are relatively invariant, whereas nitrate concentrations
undergo strong seasonally driven by snowmelt. Increases in inorganic
monomeric aluminum result when acid neutralizing capacity values fall
below zero.
Source Dnscoll et al (1989a)
10-147
-------
a strongly acidifying process (Figure 10-26) The key characteristics of Stage 2 watersheds
are elevated base-flow concentrations of NO3" that result from high groundwater
concentrations (e g , Figure 10-29). Episodic NO3" concentrations are as high as Stage 1,
but the seasonal pattern at Stage 2 is damped by an increase in base-flow concentrations to
levels as high as those found in deposition
In Stage 3, the watershed becomes a net source of nitrogen rather than a sink
Nitrogen retention mechanisms (e g , uptake by vegetation and microbes) are much reduced,
and mineralization of stored nitrogen may add substantially to nitrogen leaving the watershed
in runoff or in gaseous forms As in Stage 2, nitrification rates are substantial The
combined inputs of nitrogen from deposition, mineralization, and nitrification can produce
concentrations of NO3" in surface waters that exceed inputs from deposition alone The key
characteristics of Stage 3 watersheds are these extremely high NO3" concentrations and the
lack of any coherent seasonal pattern in NO3" concentrations
Conceptually, the stages of watershed nitrogen loss can be thought of as occurring
sequentially, as a single watershed progresses from being strongly nitrogen deficient to
strongly nitrogen sufficient This is consistent with the conceptual model presented by Aber
et al. (1989; Figure 10-27), and can be supported by two lines of evidence, presented in the
following sections of this paper The first line of evidence comes from "space for tune
substitutions" (in the sense of Pickett, 1989), where the occurrence of various stages across
a gradient of present-day nitrogen deposition is used a surrogate for the temporal sequence
that a single site might undergo if it were exposed to chronically elevated levels of nitrogen
deposition. This technique is commonly applied to current environmental problems where a
good historical record is not available (Sullivan, 1991) The second line of evidence comes
from long-term temporal trends at single sites, where increases in nitrogen efflux from
watersheds (observable as increasing trends in NO3" concentration) and changes in the
seasonal pattern of NO3" concentration can be directly attributed to the combined effects of
chronic nitrogen deposition and other factors (e g , forest maturation) The few cases where
individual sites have been observed to progress from Stage 0 to Stage 1 and/or Stage 2 of
watershed nitrogen loss are especially useful in establishing that nitrogen saturation occurs as
a temporal sequence in areas of high nitrogen deposition These lines of evidence are
discussed in the following sections
10-148
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0
=3.
W
C
I
o
O
JFMAMJ J ASONDJ FMAMJJ ASOND
1988 1989
Figure 10-29.
Temporal patterns in chemical characteristics of stream water at
Biscuit Brook in the Catskill Mountains. All chemical variables
undergo strong seasonally, with strong dependence on stream
discharge. Values for the ratio of nitrate to nitrate + sulfate
approach 0.5 during episodes, and indicate that nitrate is nearly as
important an acidifying influence as sulfate during high-flow events.
Source Murdoch and Stoddard (in press a)
10-149
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10.8.2.4 Processes Within Lakes and Streams
All of the transformations and processes discussed above (primarily in the context of
terrestrial ecosystems) also take place in lakes, streams, and estuaries The emphasis on the
transformations that occur in the watershed, before nitrogen reaches surface waters, results
from the necessity to establish a linkage between nitrogen deposition and nitrogen effects in
aquatic systems, but should not be taken to suggest that nitrogen transformations within
aquatic systems are of minor importance m the nitrogen cycle In a very real sense, nitrogen
cycling within the terrestrial ecosystems controls whether nitrogen deposition will reach
aquatic systems (and in what concentrations), whereas nitrogen cycling within lakes, streams,
and estuaries controls whether the nitrogen will have any measurable effect
Assimilation by aquatic plants is a key process in the potential eutrophication of surface
waters by nitrogen, and may also play a role m their acid/base status The following
discussion of nitrogen assimilation in aquatic systems will deal mainly with the algal and
microbial community in phytoplankton (microscopic algal and bacterial species suspended in
the water column) and penphyton (algal species growing attached to surfaces) Although
macrophytes (macroscopic algal species) are also important in the assimilation of nitrogen,
the biomass of phytoplankton and smaller microbes is potentially most reactive to changes in
nitrogen supply Algal uptake is a major component of the eutrophication process, and forms
the basis of trophic production m streams and lakes It can also play a large role in the
acid/base status of lakes Uptake of NO3" in lakes is an alkalinizing process, consuming
1 mole of hydrogen per mole of nitrogen assimilated (Kelly et al, 1990)
Like terrestrial plants, aquatic plants favor the uptake of NH4+ over the uptake of
NC>3~; NH4 uptake is energetically favorable because NO3" must first be reduced before it is
physiologically available to algae (Reynolds, 1984) In some circumstances, organic forms
of nitrogen are also available for uptake by aquatic plants (reviewed by Healey, 1973) The
preferences by algae for the different forms of nitrogen can be related to the history of
availability of nitrogen species In some algal species, the synthesis of the enzyme (nitrate
reductase) required to utilize an NO3" pool can be induced by high concentrations of NO3" in
the absence of NH4+ (Healey, 1973) The production of nitrate reductase appears to be
repressed by the presence of NH4+ (Eppley et al , 1979)
10-150
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The potential uptake rate of inorganic nitrogen is related to ambient inorganic nitrogen
concentrations (e g , Syrett, 1953), that is, cells transferred from nitrogen-deficient media to
nitrogen-sufficient media show higher rates of uptake than cells that are grown and remain in
nitrogen-sufficient media McCarthy (1981) summarized several studies that consistently
showed that potential (saturated) NH4+ uptake rates were greatly enhanced in
nitrogen-deficient cells This relationship is now used along with various other indices as
a basis to identify the degree of nitrogen limitation in phytoplankton (Vincent, 1981, Suttle
and Harrison, 1988) Under nitrogen-replete conditions, saturated uptake rates are low, but
increase with increasing nitrogen deficiency
A crucial difference between aquatic and terrestriail ecosystems with respect to nitrogen
is that nitrogen additions do not commonly stimulate growth in aquatic systems, as seems to
be the case in terrestrial systems, and nitrogen limitation may in fact be the exception in
aquatic systems rather than the rule Determining whether nitrogen limitation is a common
occurrence in surface waters will play a large role in determining whether nitrogen
deposition affects the trophic state of aquatic ecosystems
The effects of nitrogen supply on uptake and growth rates in phytoplankton and
penphyton is the subject of volumes of literature, a summary of which is beyond the scope
of this section However, certain aspects of the limitation of algal growth by the supply of
nitrogen and other nutrients will be discussed later as it relates to enrichment effects from
nitrogen deposition For other details on algal nutation, the reader is referred to reviews by
Goldman and Gilbert (1982), Button (1985), Kilham and Hecky (1988), and Hecky and
Kilham (1988)
Demtnfication plays a much larger role m nitrogen dynamics in aquatic ecosystems
than it does in terrestrial ones In streams, rivers, and lakes, bottom sediments are the mam
sites for demtnfication (see Seitzinger, 1988a), although open-water denitafication has also
been reported (Keeney et al , 1971) In lake and stream sediments, the main source of
NO3", although potentially available from the water column, is NO3" produced when organic
matter is broken down within the sediments, and the resulting NH4+ is subsequently oxidized
(Seitzinger, 1988a) Demtnfication is an especially important process in large rivers and
estuaries, and will play a large role m discussions of nitrogen loading to estuaries and near-
coastal systems (see Section 10 8 4 2) In a recent review of demtnfication in freshwater
10-151
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and estuarine systems, Seitzmger (1988a) reported demtnfication rates that were 7 to 35 % of
nitrogen inputs in large rivers, and 20 to 50% of inputs in estuaries Demtnfication in
aquatic ecosystems is an alkalimzing process, consuming 1 mole of hydrogen for every mole
of N03~ denitrified
2
Estimates of demtnfication rates range from 54 to 345 jttmol/m /h in streams with high
2
rates of organic matter deposition, 12 to 56 ^mol/m /h in (nutnent-poor) oligotrophic lakes,
2 2
42 to 171 jtmol/m /h in eutrophic lakes, and 77 to 232 jwmol/m /h in estuanes (see
Seitzinger, 1988a) These values are in the range where demtnfication can deplete NO3"
pools Rudd et al (1990) have reported an increase in the rate of denitriflcation from less
*) 2
than 0.1 /jmol/m /h to over 20 /^mol/m /h ui an okgotrophic lake when nitric acid was added
m a whole-lake experimental acidification, suggesting that freshwater demtnfication may be
limited by NO3" availability Demtnfication can account for 76 to 100% of nitrogen flux at
sediment-water interfaces m nvers, lakes, and estuanes (Seitzmger, 1988a) In the Potomac
and Delaware nvers, where organic sediment deposition is extreme due to sewage inputs, the
loss represents 35 and 20%, respectively, of external nitrogen inputs In estuanes, it can
represent a 50% loss In the deep mud of slow-flowing streams, the process can effectively
reduce NO3" concentrations in the water column by as much as 200 jweq/L over a 2 km
length of stream (Kaushik et al, 1975, Chatarpaul and Robinson, 1979) This depletion
amounts to 75 % of the daily input of NO3" during a growing season, and it has been
sufficient to consider denitnfication as a method for NO3" removal in the management of
some slow-moving streams having a deep organic substrate (Robinson et al, 1979)
Nitrogen fixation counteracts demtnfication losses of nitrogen from surface waters and
is fundamental to replenishing fixed forms of nitrogen in all aquatic ecosystems It is
thought to be the main process responsible for maintaining surplus inorganic nitrogen m lakes
and streams and is fundamental to the fact that primary production m most lakes and streams
is limited by phosphorus (Schindler, 1977) In estuanes, however, there is a higher loss of
nitrogen relative to that fixed or imported The loss may be due to high rates of
denitrification (Seitzmger, 1988a), which creates relative nitrogen deficiencies
Rates of nitrogen fixation are generally related to trophic status m freshwater Howarth
et al. (1988a) show that fixation in low-, medium-, and high-nutrient lakes is generally
<0.02} 0.9 to 6.7, and 14 3 to 656 9 mmol mtrogen/m2/year, respectively Fixation is also
10-152
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closely correlated with the abundance of blue-green algae (Wetzel, 1983), which suggests
that the algae, rather than bactena, dominate nitrogen fixation in lakes Although nitrogen
fixation does occur in sediments, that source is of minor importance compared to that in the
water column Only in very nutrient-poor lakes, where nitrogen loading from all other
sources is small, can nitrogen fixation in sediments gain some significance (e g., 32% and
6% of total inputs in Lake Tahoe, CA, and Mirror Lake, NH, respectively, Howarth et al ,
1988a)
Unlike the nitrogen fixation community in lakes, nitrogen fixers in estuaries are
2
dominated by bactena, producing rates of 0 1 to 111 mmol mtrogen/m /year (Howarth et al ,
1988a) The highest rates occur in deep organic sediments, but even these are a relatively
small percentage of total nitrogen inputs to estuaries (reviewed by Howarth et al , 1988a)
As in terrestrial watersheds, rates of nitrification in lakes and streams are often limited
by low concentrations of NH4+ Supply rates of NH4+ from watersheds are often low
(except in cases of point-source pollution), and nitrifying organisms have little substrate with
which to work Two exceptions to this generality are cases where NH4+ deposition is
extremely high, such as near agricultural areas, and cases where NH4+ is produced within
the aquatic system Experiments on whole lakes and in mesocosms in Canada have
confirmed the acidifying potential of ammonium additions from deposition to surface waters
(Schindler et al , 1985, Schiff and Anderson, 1987) Ammonium deposition is especially
deceptive because in the atmosphere, ammonium can combine as a neutral salt with SO42",
resulting in precipitation with near-neutral pH values, as seen in the Netherlands
(Van Breemen and Van Dijk, 1988) Once deposited, however, the ammonium can be
assimilated, leaving an equivalent amount of hydrogen, or it can be nitrified, leaving twice
the amount of hydrogen There is some evidence from Canadian whole-lake experiments that
nitrification in lakes is an acid-sensitive process, Rudd et al (1988) presented data indicating
that nitrification was blocked at pH values less than 5 4 in an experimentally acidified lake,
leading to a progressive accumulation of NH4+ in the water column
High NH4+ concentrations may also result in lakes whose deeper waters become anoxic
during periods of stratification (usually late winter or late summer) Production of
(by decomposition) can be substantial under anaerobic conditions, and NH4+ may accumulate
in the anoxic water Nitrification of this NH4+ occurs when lakes mix during spring or fall,
10-153
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supplying the oxygen necessary for nitrifying organisms to survive (Wetzel, 1983)
In estuaries, the processes of nitrification (aerobic) and demtnfication (anaerobic) may be
closely coupled at the sediment surface, with mineralization in the anaerobic sediments
supplying NH4+ to nitnfiers at the sediment/water interface (Jenkins and Kemp, 1984)
Except in cases where the overlying water becomes anoxic (as may be common in the
summer months), the nitrifying organisms supply NO3" back to the sediments for subsequent
denitnfication. In both cases described above (the annual cycle in lakes and the
sediment/water interface cycle in estuaries), the main influence of nitrification is to recycle
nitrogen within the system and to supply NO3" to either denitnflers or to nitrogen-deficient
algae.
In lakes, streams, and estuaries, water is in constant movement, and, to a large extent,
the effects of nitrogen cycling on biota are regulated by the local hydrology In lakes,
oxidation and reduction reactions are perceived to occur as cycles in the sense that water has
a residence time lasting from a few weeks in small ponds to many years in large lakes
Nitrogen species are assimilated, they contribute to biological productivity, the organic forms
are subsequently mineralized, and the resulting inorganic forms enter various oxidizing and
reducing pathways mediated by a microbial community within a single body of water One
or more complete cycles can be followed within a single lake before export downstream
In streams, and to some extent in estuaries, nitrogen dynamics are more closely
dependent on the physical movements of water As nitrogen compounds are cycled among
the biotic and abiotic components of the stream ecosystem, they are subject to downstream
transport. Among stream ecologists, this coupling between nutrient cycles and water
movement is termed "nutrient spiraling" (e g , Elwood et al, 1980, Newbold et al, 1983)
According to this concept, nitrogen cycling occurs in most streams, but little or no recycling
occurs in any one place Nitrogen is instead regenerated or transformed at one point in the
stream and transported downstream before subsequent reutilization or retransformation
(Stream Solute Workshop, 1990) The movement of water can increase nutrient uptake rates
and growth rates in freshwater algae (Whitford and Schumacher, 1961, 1964) by continually
resupplying nutrients at cell walls This constant replenishment prevents steep concentration
gradients from becoming established, as can happen in less active lake water (Gavis, 1976)
It also maintains high rates of production and nutrient assimilation Biomass eventually
10-154
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sloughs from substrata, and drifts as fine paniculate organic matter (Meyer and Likens,
1979) for settlement, decomposition, and mineralization downstream Very high flows
associated with intense precipitation events are physically disruptive and can increase the
concentration of particulates transported downstream (Bilby and Likens, 1979, Holmes et al,
1980) Efficiencies of nutrient uptake also decrease wiih increasing flows because of reduced
contact tune that a given ion has with the reactive substrate (Meyer, 1979)
One important consequence of nutrient spiraling ui streams is that any block in the
nitrogen cycle upstream can have potential effects on nitrogen conditions downstream
Mulholland et al (1987), for example, have presented experimental evidence that leaf
decomposition (mineralization) in streams is inhibited at low pH values Because
mineralization of organic matter is an important process in resupplying nitrogen to organisms
downstream, the existence of acidic headwaters could influence biotic conditions in
downstream portions of streams where acidification is riot important
10.8.3 The Effects of Nitrogen Deposition OKI Surface Water Acidification
The acidification processes of lakes and streams are conventionally separated into
chrome (long-term) and episodic (event-based) effects A great deal of emphasis in the past
decade has been placed on chronic acidification in general, and on chronic acidification by
sulfate in particular (e g , Galloway et al , 1983, Sullivan et al, 1988, Brakke et al, 1989)
This emphasis on SO4 " has resulted largely because sulfur deposition rates are often higher
than those for nitrogen (sulfur deposition rates are approximately twice the rates of
nitrogen deposition in the Northeast, Stensland et al, 1986) and because NO3" appears to be
of negligible importance m surface waters sampled during summer and fall index periods
(Linthurst et al , 1986) As mentioned previously, summer and fall are seasons when
watershed demand for nitrogen is very high, creating a low probability that nitrogen, m any
form, will be leached into soil and surface waters unless the watersheds have achieved
nitrogen saturation Under conditions of low nitrogen deposition (or high nitrogen demand),
nitrogen leaking from terrestrial ecosystems, as described earlier, is more likely to be a
transient (or seasonal) phenomenon than a chronic one As a result, the primary impact of
nitrogen in surface water acidification will be observed during high-flow seasons, and
particularly during snowmelt It has been estimated that 40 to 640 % more streams in the
10-155
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eastern United States (Florida to the Northern Appalachian Plateau) are acidic during spring
episodes than are acidic during spring base flow, whereas the number of acidic Adirondack
lakes is estimated to be three tunes higher during the spring than during the fall (Eshleman,
1988).
Surface waters are conventionally considered acidic if their acid-neutralizing capacity
(ANC) is less than zero The ANC of a lake or stream is a measure of the water's capacity
to buffer acidic inputs, and results from the presence of carbonate and/or bicarbonate
(or alkalinity), Al, and organic acids in the water (Sullivan et al, 1989) The main purposes
of this section are to evaluate the evidence for chronic acidification by nitrogen deposition in
North America, and to determine what role nitrogen deposition plays in episodic
acidification.
10.8.3.1 Chronic Acidification
In the United States, the most comprehensive assessment of chronic acidification of
lakes and streams comes from the National Surface Water Survey (NSWS) conducted as part
of the National Acid Precipitation Assessment Program The NSWS surveyed the acid/base
chemistry of both lakes and streams using an "index period" concept The goal of the index
period concept was to identify a single season of the year that exhibited low temporal and
spatial variability and that, when sampled, would allow the general condition of surface
waters to be assessed (Linthurst et al, 1986) In the case of lakes, the index period selected
was autumn overturn (the period when most lakes are mixed uniformly from top to bottom),
and in streams, the chosen index period was spring base flow (the period after spring
snowmelt and before leaf-out) (Messer et al, 1988) Because of the strong seasonably of the
nitrogen cycle in forested watersheds (described earlier), the choice of index penod plays a
very large role in the assessment of whether nitrogen is an important component of
acidification
The results of the Eastern Lake Survey (Linthurst et al, 1986), based on a probability
sampling of lakes during fall overturn, suggest that nitrogen compounds make only a small
contribution to chronic acidification in North America Hennksen (1988) has proposed that
the ratio of NO3" NO3"+SO42" in surface waters be used as an index of the influence of NO3"
on chrome acidification status This index assesses the importance of nitrogen relative to the
10-156
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2
importance of SO4 ", which is usually considered more important in chronic acidification (see
above) A value greater than 0 5 indicates that NO3" has a greater influence on the chronic
2
acid/base status of surface waters than does SO4 " Hennksen (1988) summarized the ratios
for acid-sensitive sites worldwide, these results are repeated in Table 10-20 In general,
Hennksen's results show that NO3" can be almost as important as SO42" in some parts of
Europe, but that ratios are low in the United States(see also Hennksen and Brakke, 1988)
One problem with Hennksen's approach, however, is that he compares data collected
intensively (i e , through multiple samples per year) with survey data collected during a
single index penod The data presented for Adirondack lakes in Table 10-20, for example,
were collected monthly over a 2-year penod (Dnscoll and Newton, 1985), and the apparent
difference between the Adirondacks and the rest of central New England (from the regional
survey data) could well result from comparing fall values to annual mean values Annual
mean values include high spring NO3" concentrations in runoff waters and will, therefore, be
higher than concentrations measured only in the autumn As a result, the ratio values
reported in Table 10-20 for the Adirondacks are an indication that NO3" may be important in
chronic acidification (i e , NO3" makes up about 15% of acid anions), but the low ratios
reported for the Eastern Lake Survey are not informative Unfortunately, no regional lake
survey with representative annual, or spring, values exists for the United States, and
questions concerning the role of NO3" in chronic lake acidification remain unanswered for
areas outside of the Adirondacks
2
Values of NO3" NO3"+SO4 " ratios are also available for streams from the National
Stream Survey (NSS) (Kaufmann et al, 1988), as well as from other regional stream surveys
(e g , Stoddard and Murdoch, 1991) Median values for each of the regions covered in these
surveys are given in Table 10-21 The NSS data have Ihe advantage of having been collected
dunng a spring base-flow index penod This penod is been shown to be a good index of
mean annual condition for streams (Messer et al, 1988, Kaufmann et al, 1988), but is not
an estimate of worst case condition, as concentrations taken during spring snowmelt would
be The Catskill regional data included in Table 10-21 are from a stream survey that
included multiple samplings per year (Stoddard and Murdoch, 1991) Several stream regions
10-157
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TABLE 10-20. CONCENTRATIONS OF NITRATE, SULFATE, AND RATIOS
OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN RUNOFF WATERS
IN ACIDIFIED AREAS OF THE WORLD3
Concentration (jieqfL)
Location
West Germany
Lange Bramke
Lange Bramke
Bavenscher Wald
Rachclsee
Gr Arbersee
KI. Arbersee
Poland
The Giant Mountains
Maly Staw
Wielki Staw
Czechoslovakia
Tatra Mountains
av 53 lakes
Jameke
Popradake
Vyshe Wahlenbugoro
Vyshe Furkotake
Bohemia
Came
Certovo
Prasilske
Plesne
Laka (man-made)
Zdarske (man-made)
Krusne hory Mountauis
Sumava Mountains
Liz
Albrechtec
Norway
Birkenes
Storgama
Sweden
Stromyra
Scotland
av 22 lakes in
the Galloway area
Year
1977
1984
1985
1985
1985
1986
1986
1984
1980-82
1980-82
1980-82
1980-82
1986
1986
1986
1986
1986
1986
1986
April '86
April '86
1973-86
1973-86
1984-85
1979
pH
58
62
45
47
45
55
47
6 1
44
66
56
63
45
42
45
47
55
65
52
589
622
452
456
654
497
N03"
16
49
77
98
93
13
40
37
2
40
44
42
93
85
40
41
45
0
118
136
36
9
12
17
21
S042'
233
230
135
118
108
92
140
97
171
111
74
110
152
182
120
203
61
156
1216
390
358
140
77
180
103
Ratio
NO3' NO3" + S042"
006
0 18
036
045
046
0 12
022
027
001
026
037
028
038
032
025
0 17
042
000
009
026
009
006
0 13
009
0 17
Sampling
Methodb
Intensive
Intensive
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Unknown
Intensive
Intensive
Intensive
Unknown
10-158
-------
TABLE 10-20 (cont'd). CONCENTRATIONS OF NITRATE, SULFATE, AND
RATIOS OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN RUNOFF
WATERS IN ACIDIFIED AREAS OF THE WORLD3
Concentration (/teq/L)
Location
United States
Adirondacks
Big Moose Lake
Cascade Lake
Darts Lake
Memam Lake
Lake Rondaxe
Squash Pond
Townsend Pond
Windfall Pond
Bubb Lake
Constable Pond
Moss Lake
Black Pond
Clear Pond
Heart Lake
Otter Lake
West Pond
Woodruff Pond
Eastern Lake Survey
Southern Blue Ridge
Florida
Upper Midwest
Upper Great Lakes
Wisconsin
Peninsula, Michigan
Northeastern Minnesota
Maine
Southern New England
Central New England
Canada
Experimental Lakes
Area, Ontario
Sudbury, Ontario
Kekimkujik,
Nova Scotia
Year PH
1980s 5 1
65
52
64
59
46
52
59
6 1
52
64
6 8
70
64
55
52
69
1985
-
-
-
-
-
-
-
-
-
1980s
1980s
1980s
N03"
24
29
24
26
23
24
27
26
16
17
26
4
1
5
9
10
2
3
1
07
06
1 0
06
09
02
08
03
1
2
2
3
S042'
140
139
139
141
134
131
154
141
131
149
141
130
139
106
138
111
147
32
94
57
50
57
78
62
75
141
101
78
252
152
78
Ratio
N03" NO3" + S042"
0 15
0 17
0 15
0 16
0 15
0 15
0 15
0 16
0 11
0 10
0 16
003
000
005
006
008
001
009
001
001
001
002
001
001
000
001
000
001
001
001
004
Sampling
Methodb
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Monthly
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Fall index
Intensive
Intensive
Intensive
aNO3" = Nitrate ion
SO42"= Sulfate ion
Sampling methods are listed as unknown, monthly, intensive (more frequent than monthly), or based on a
single fall index sample
Median value for regional population of lakes
Source Hennksen (1988)
10-159
-------
TABLE 10-21. CONCENTRATIONS OF NITRATE, SULFATE, AND RATIOS
OF NITRATE TO THE SUM OF NITRATE AND SULFATE IN STREAMS OF
ACID-SENSinVE REGIONS OF THE UNITED STATES. VALUES ARE
MEDIANS FOR REGION (FIRST AND THIRD QUARTILES IN PARENTHESES)2
Concentration (/ieg/L)
Location
National Stream Survey
Poconos/Catskills
Northern Appalachians
Valley and Ridge
Mid-Atlantic Coastal Plain0
Southern Blue Ridge
Piedmont
Southern Appalachians
Ozarks/Ouachitas
Florida
Catskill Regional Survey
Median value for 51 streams
Year PH
1986 6 96
6 60
705
598
699
6 80
733
662
548
1984-86 6 60
NO3"
6
(2-18)
30
(12-41)
10
(3-31)
-
8
(2-16)
2
(0-5)
16
(3-32)
1
(1-4)
5
(1-10)
29
(14-47)
so42-
169
(154-184)
171
(135-347)
154
(84-294)
-
17
(10-27)
48
(19-63)
58
(30-104)
59
(48-83)
22
(9-30)
138
(125-151)
NO3" NO3" + SO42"
003
(0 01-0 10)
0 14
(0 02-0 19)
009
(0 01-0 22)
-
028
(0 08-0 44)
003
(0-0 20)
032
(0 04-0 40)
002
(0-0 06)
0 19
(0 10-0 25)
0 17
(0 09-0 26)
= Nitrate ion
SC>4 = Sulfate ion
K *?
Values for pH are for entire region (Kaufmann et al , 1988), medians for NO3", SO4 ", and the
NOj" NC«3~ + SC>4 " ratio exclude sites with potential agricultural or other land-use impacts (Kaufmann et al ,
1991)
''The influence of agricultural and land use practices could not be ruled out for any of the sites in the
Mid-Atlantic Coastal Plain (Kaufmann et al , 1991)
dFrom Stoddard and Murdoch (1991)
exhibit ratios as high as those reported for the Adirondacks by Hennksen (1988) Several
regions in the southeastern United States exhibit high ratios in part because their current
SO42" concentrations are relatively low The Southern Blue Ridge, in particular, has the
lowest NO3" concentrations found in the NSS, and the relatively high NO3" NO3"+SO42"
ratios in this region could be considered misleading The stream data do suggest that the
Catskills, Northern Appalachians, Valley and Ridge Province, and Southern Appalachians all
show some potential for chronic acidification due to NO3" In all of the stream regions in
10-160
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Table 10-21, as well as the lake regions in Table 10-20, however, chronic acidification is
2.
more closely tied to SO4 " than to NO3"
The data presented thus far in this section establish which regions of the country show
potential problems with chronic acidification by NO3", but do not indicate whether the source
of the NO3" is atmospheric deposition As described earlier, several watershed processes
(e g , mineralization, nitrification, nitrogen fixation) may combine to produce NO3" and may
be responsible, at least in part, for high NO3" concentrations observed in surface waters
On a regional scale, it is not possible to attribute surface water NO3" to any single source,
but two efforts have been made to relate rates of nitrogen deposition to rates of nitrogen loss
from watersheds Data from the NSS (Kaufmann et al , 1991) suggest a strong correlation
between concentrations of stream water nitrogen (NO3~ + NH4 ) at spring base flow and
levels of wet nitrogen deposition (NO3~ + NH4+) in each of the NSS regions
(Figure 10-30) The only exception to this relationship is the Pocono/Catskill region, where
nitrogen deposition is highest (6 kg/ha/year), but where stream water nitrogen concentrations
fall below what is expected, based on results from the other regions The median stream
water NO3" value for the CatsMLs alone (from Stoddard and Murdoch, 1991, Table 10-21) is
29 /^eq/L, and fits the relationship much more closely, suggesting that watersheds in the
southern portion of this region (the Poconos) are retaining nitrogen more strongly than the
northern portion Dnscoll et al (1989a) collected inpul /output budget data for a large
number of watersheds in the United States and Canada, and summarized the relationship
between nitrogen export and nitrogen deposition at all of the sites (Figure 10-30) The
authors stress that the data illustrated in Figure 10-30 were collected using widely differing
methods and over various tune scales (from 1 year to several decades) Given the numerous
possible sources of NO3" and the watershed pathways through which nitrogen may be cycled,
the relationships illustrated in Figure 10-30 should not be over-interpreted, nor should they
be construed as an illustration of cause and effect However, the relationships do show that
watersheds in many regions of North America are retaining less than 75 % of the nitrogen
that enters them, and that the amount of nitrogen being leaked from these watersheds is
higher in areas where nitrogen deposition is highest This pattern is consistent with what we
would expect if large areas of the eastern United States were experiencing the early stages of
nitrogen saturation Furthermore, both analyses suggest a threshold value of nitrogen
10-161
-------
I 40-
f 30-
'o"
=E 20-
I .
i
(a)
0-
100 200 300 400 500
Wet NO + NH/ Deposition (eq/ha/year)
?
i
cc
z
(b)
350-
300-
250-
200-
150.
100-
50.
0 -
0
» 0
0 0
°o
o
o
° 0
° 0 ° °0
o
0 O
o o o o
^oooo-) ^OOO o
100 200 300 400 500 600
Rats of Nitrogen Wet Deposition (eq/ha/year)
Figure 10-30. Nitrogen deposition and watershed nitrogen loss, (a) Relationship
between median wet deposition of nitrogen (nitrate ions plus ammonium
ions) and median surface water nitrogen (nitrate ions plus ammonium
ions) concentrations for physiographic districts within the National
Stream Survey that have minimal agricultural activity. [Subregions are
Poconos/Catskills (ID), Southern Blue Ridge Province (2As), Valley and
Ridge Province (2Bn), Northern Appalachians (2Cn), Ozarks/Ouachitas
(2D), Southern Appalachians (2X), Piedmont (3A), Mid-Atlantic Coastal
Plain (3B), and Florida (3C)]. From Kaufmann et al. (1991).
(b) Relationship between wet deposition of nitrogen (nitrate ions plus
ammonium ions) and rate of nitrogen export for watershed studies
throughout North America. Sites with significant internal sources of
nitrogen (e.g., from alder trees) have been excluded.
Source Dnscoll et al (1989a), additional data from Barker and Witt (1990), Edwards and Helvey (1991),
Kelly and Meagher (1986), Katz et al (1985), Buell and Peters (1988), Weller et al (1986), Owens
et al (1989), Feller (1987), Stoddard and Murdoch (1991)
10-162
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deposition (ca 3 kg/ha/year) above which substantial watershed losses of nitrogen might
begin to occur
Chronic acidification due to nitrogen deposition is much more common in Europe than
in North America (Hauhs et al , 1989) Many sites show chronic increases in nitrogen
export from their watersheds (e g , Hennksen and Brakke, 1988, Hauhs, 1989), and at sites
with the highest stream water NO3" concentrations (i e , Lange Bramke and Dicke Bramke in
West Germany), NO3" concentrations no longer show the seasonally that is expected from
normal watershed processes (Hauhs et al, 1989) Hennksen and Brakke (1988) have
reported regional chronic increases in surface water NO3" in Scandinavia in the past decade
These increases in NO3" concentration are associated with increasing concentrations of Al,
which is toxic to many fish species (Hennksen et al , 1988, Brown, 1988) There is some
evidence that NO3" has a greater ability to mobilize toxic Al from soils than does SO42"
(James and Rilia, 1989) Chronic acidification attributable to ammonium deposition has also
been demonstrated in the Netherlands (Van Breemen and Van Dijk, 1988, Schuurkes, 1986,
1987) As descnbed earlier, ammonium in deposition can be nitrified to produce both NO3"
and H+, which are subsequently leaked into surface waters Rates of NO3" and NH4+
deposition are much higher in Europe (in some places deposition is >2,000 eq/ha/year,
Rosen, 1988) than in the United States (Table 10-19), and it has been suggested that chronic
nitrogen acidification is more evident in Europe than in North America because nitrogen
saturation (see discussion above) is further progressed in Europe
10.8.3.2 Episodic Acidification
In a recent comprehensive examination, Wigington et al (1990) reported that acidic
episodes have now been observed in a wide range of geographic locations in Scandinavia
(Norway, Sweden, Finland), Europe (United Kingdom, Scotland, Federal Republic of
Germany, Czechoslovakia), and Canada (Ontario, Quebec, Nova Scotia), as well as in the
United States They noted that in the United States, episodes have been registered in surface
waters in the Northeast, Mid-Atlantic, Mid-Atlantic Coastal Plain, Southeast, Upper
Midwest, and West regions In the Mid-Atlantic Coastal Plain and Southeast regions, all of
the episodes cataloged to date have been associated with rainfall In contrast, most of the
10-163
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episodes in the other regions are related to snowmelt, although rain-driven episodes
apparently can occur in all regions of the country
The regional importance and seventy of episodic acidification have not been quantified,
that is, the regional information on chronic acidification that was gained from the NSWS has
no parallel in episodic acidification As a result, all of the information we currently have
about the importance of episodes, and the influence of nitrogen deposition on episodes,
comes from site-specific studies It is important to stress that even within a given area, such
as the Northeast, major differences can be evident in the occurrence, nature, location (lakes
or streams), and timing of episodes at different sites
Eshlenian (1988) has used a simple stream mixing model (Johnson et al, 1969) to
predict the number of streams in the NSS that would be acidic during spring episodes, based
on their spring base-flow chemistry In addition, Eshleman used an empirical model relating
fall index period lake chemistry to spring episodic chemistry, using data from the U S
Environmental Protection Agency's (EPA's) Long-Term Monitoring project (Newell et al,
1987), to predict the number of Adirondack lakes that undergo episodic acidification His
results are repeated in Table 10-22 Eshleman's approach has been criticized (see discussion
below), largely because it assumes that all lakes, regardless of their baseline ANC, undergo
the same relative depression in ANC during episodes (i e , that the relationship between fall
and spring ANC is linear) This assumption ignores any effect of increased NO3" during
episodes, which may be greater in low ANC lakes (Schaefer et al , 1990, Schaefer and
Driscoll, in press). Given this criticism, Eshleman's estimates of the number of episodically
acidified systems should probably be considered conservative
A number of processes contribute to the tuning and seventy of acidic episodes (Dnscoll
and Schaefer, 1989) The most important of these processes are
• dilution of base cations (Galloway et al, 1980) by high discharge,
• increases in organic acid concentrations (Sullivan et al, 1986) during penods of high
discharge,
• increases in SO4 " concentrations (Johannessen et al, 1980) during penods of high
discharge, and
• increases in NO3" concentrations (Galloway et al, 1980, Dnscoll and Schafran,
1984, Schofield et al, 1985) during penods of high discharge
In addition to these factors, which produce the chemical conditions charactenstic of episodic
events, the likelihood of an acidic episode is also influenced by the chemical conditions
10-164
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TABLE 10-22. ESTIMATES OF THE NUMBER AND PROPORTION OF
CHRONICALLY AND EPISODICALLY ACIDIC LAKES AND STREAM REACHES
IN THE EASTERN UNITED STATES. CHRONIC CONDITIONS BASED ON
RANDOM SAMPLE OF SYSTEMS DURING INDEX CONDITIONS (SPRING
BASE FLOW OR FALL OVERTURN). EPISODIC CONDITIONS ESTIMATED
FROM TWO-BOX MIXING MODEL (FOR STREAMS), OR EMPIRICAL
RELATIONSHIPS BETWEEN FALL INDEX PERIOD AND SPRING SNOWMELT
CHEMISTRY (FOR LAKES)
Index Conditions (ANC? < 0)
Subregion
Stream Subregions
Poconos/Catskills
Southern Blue Ridge
Valley and Ridge
Northern Appalachian Plateau
Ozarks/Ouachitas
Southern Appalachians
Piedmont
Mid- Atlantic Coastal Plain
Florida
Lake Subregions
Adirondacks
Number
209
0
636
499
0
121
0
1,334
678
138
Proportion ( %)
64
0
49
58
0
25
0
11 8
392
107
Episodic Conditions (ANC? < 0)
Number
746
39
1,126
3,224
75
364
0
3,132
963
459
Proportion (%)
230
22
86
372
1 8
74
0
278
557
356
aANC = Acid-neutralizing capacity
For streams, all data are from the upper end of sampled stream reaches (Kaufmann et al , 1988), except for
the Southern Blue Ridge, where data from lower ends of stream i caches were used
Source Eshleman (1988)
before the episode begins Episodes are more likely to be acidic, for example, if the base-
flow ANC of the stream or lake is low In this way, acid anions, especially SO42", can
contribute to the seventy of an acidic episode, even though they do not increase during the
event, by lowering the base-flow ANC of the stream or lake (Stoddard and Murdoch, 1991)
In many cases, all of these processes will contribute to episodes in a single aquatic
system Dilution, for example, probably plays a role in all episodic decreases in ANC and
pH in all regions of the United States (Wigmgton et al , 1990) Dilution results from the
increased rate of runoff, and channeling of runoff through shallower soil layers, that occurs
10-165
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during storms or snowmelt, the shorter contact time produces runoff with a chemical
composition closer to that of atmospheric deposition than is typical of base-flow conditions
(e.g , Driscoll and Newton, 1985, Peters and Murdoch, 1985, Stoddard, 1987a) Because
precipitation is usually more dilute than stream or lake water, storm runoff produces surface
waters that are more dilute than during non-runoff periods In a sense, dilution sets the
*J
baseline condition to which the effects of organic acids and atmospherically derived SO4 "
and NC*3~ are added
Little information exists about the effects of changes in organic acids during episodes
Driscoll et al. (1987a) and Eshleman and Hemond (1985) concluded that organic acids did
not contribute to snowmelt episodes in the Adirondacks or in Massachusetts, respectively
At Harp Lake in Canada, organic acidity is believed to remain constant (Servos and Mackie,
1986) or decrease (LaZerte and Dillon, 1984) during snowmelt episodes Haines (1987) and
McAvoy (1989) have documented increases in organic acidity during rain-caused episodes in
coastal Maine and in Massachusetts
2. 9
Storage of SO4 " in watersheds, and subsequent release of SO4 " during episodic events,
is well documented in many parts of Europe (Wigrngton et al, 1990), but has not been
commonly found in the United States Sulfate episodes have been described for the Leading
Ridge area of Pennsylvania (Lynch et al, 1986) and at Filsen Creek in Minnesota (Schnoor
et al., 1984), but are not widespread Sulfate does contribute to episodic acidity, however,
in the sense that concentrations may remain high during events, and contribute to a lower
baseline ANC; the effects of other factors, such as increased NO3", will be in addition to any
constant effect of SO42" in lowering the baseline ANC (Stoddard and Murdoch, 1991)
The main goal of this section is to determine when increases in NO3" concentrations
play a significant role in episodic acidification In the Adirondacks, for example, strong
NO3" pulses in both lakes (Galloway et al, 1980, Driscoll and Schafran, 1984) and streams
(Driscoll et al, 1987b) are apparently the primary factor contributing to depressed ANC and
pH during snowmelt. Schaefer et al (1990) examined the same empirical relationships used
for the Adirondack lakes by Eshleman (1988, Table 10-22) and concluded that the magnitude
of the episodes experienced by lakes depends strongly on their base cation concentration
They concluded that lakes with high base cation concentrations (and, therefore, high ANC
values) undergo episodes that are largely the result of dilution by snowmelt Low ANC
10-166
-------
lakes, on the other hand, undergo episodes that result lairgely from increases in NO3"
concentrations At intermediate ANC levels, lakes are .affected by both base cation dilution
and NO3" increases, and, therefore, these lakes may undergo the greatest increases in acidity
during snowmelt episodes (Figure 10-31) The relationship between spring and fall lake
chemistry is, therefore, not linear, as assumed by Eshleman (1988), and the number of lakes
that become acidic during spring episodes is probably larger than predicted in Table 10-22
Dnscoll et al (1989a,b) report on a detailed study of nitrogen dynamics in
Pancake-Hall Creek in the Adirondack Mountains This stream is highly acidic, with low
9
and invariant concentrations of base cations, and high and invariant concentrations of SO4 "
9
(Figure 10-28) Nitrate concentrations were lower than SO4 " concentrations, and exhibited a
distinct seasonal pattern, peak concentrations approached 100 /neq/L Short-term changes in
NO3" were highly correlated, and chemically consistent, with changes in the concentrations
i o _L_
of acidic cations (H and Al ) (Dnscoll et al , 1989a) As mentioned earlier, although
dilution of base cations and increases in NO3" appear to be the primary causes of episodic
acidification in Pancake-Hall Creek, these episodes are excursions from an already low
9
baseline ANC, which can be largely attributed to high SO4 " concentrations
Stoddard and Murdoch (1991) have concluded that increases in NO3", base cation
dilution, and high baseline SO4 " concentrations all contribute to acidic episodes in Catskill
Mountain streams (Figure 10-29) In Biscuit Brook, an intensively-studied stream in the
9
Catskills, concentrations of NO3" approach those of SO4 " during episodes (Murdoch and
Stoddard, in press a) Values for the ratio of NO3" NO3" + SO4 ", as presented in
Tables 10-20 and 10-21, illustrate both the general importance of NO3" to the acid/base
dynamics of this stream, and the increase in importance of NO3" during high-flow events
(Figure 10-29)
Researchers at the Hubbard Brook Experimental Forest in New Hampshire have been
studying the links between atmospheric deposition, watershed processes, and stream water
chemistry since 1963 (Likens et al , 1977) In reference Watershed #6, stream water NO3"
concentrations undergo strong seasonal cycles, with peak concentrations as high as 85
i 9
Both NO3" and H concentrations increase during snowmelt at Hubbard Brook, and SO4 "
concentrations decrease slightly (Johnson et al , 1981, Likens, 1985)
10-167
-------
u
0)
"5
o
V)
O)
XJ
c
1
Q
O
\OJ
100 .
80 ,
60 -
40 -
20 -
n -
•_
»
• 1
•
- •
• • *
1
(a)
-40 0 40 80 120 160 200 240
Baseline ANC (|j.eq/L)
CO C
5 9
=
S
1 -
DC o o-
(b)
-50 0 50 100 150 200 250
Baseline ANC (neq/L)
Figure 10-31. Effect of baseline acid-neutralizing capacity and episodic conditions in
Adirondack lakes, (a) Relationship between baseline acid-neutralizing
capacity and the springtime depression in acid-neutralizing capacity
(baseline acid-neutralizing capacity—minimum acid-neutralizing capacity)
for 11 lakes sampled in 1986 and 1987. (b) The relative contributions of
base cations and nitrate to the springtime acid-neutralizing capacity
depressions in Adirondack lakes. Lakes at intermediate acid-neutralizing
capacity values undergo the largest springtime depressions in acid-
neutralizing capacity. Lakes with lower baseline acid-neutralizing
capacity are affected more by nitrate pulses, and lakes with higher
baseline acid-neutralizing capacity are affected more by base cation
dilution. Solid lines represent best-fit relationships.
Source. Schaeferetal (1990)
10-168
-------
The highest recorded NO3" concentrations in streams draining undisturbed watersheds in
the United States come from the Great Smoky Mountains in Tennessee and North Carolina
Nitrate concentrations in Raven Fork (Jones et al , 1983), Chngman's Creek, and Cosby
Creek (Elwood et al , 1991) range from 50 to 100 /*eq/L, and in all cases are comparable to,
or higher than, SO4 " concentrations In a survey of stream chemistry at a large number of
sites in the Smokies, Silsbee and Larson (1982) reported NO3" concentrations ranging from
0.2 to 90 jiieq/L, NO3" concentrations were highest at higher elevations and in areas of old-
growth spruce-fir forest that have never been logged. In many cases, NO3" concentrations in
streams of the Smoky Mountains are higher than nitrogen concentrations in deposition,
suggesting both that rates of biological nitrogen uptake are low, and that mineralization rates
are high (Joslm et al, 1987) Unfortunately, few data are available to suggest the original
source of nitrogen now being mineralized in this region Unless nitrogen fixation rates have
been historically quite high, at least some of the NO3" now being leaked from watersheds in
the Smokies must have originated as atmospheric deposition The data of Silsbee and Larson
(1982) suggest strongly that forest maturation is linked to the process of NO3" leakage from
Great Smoky Mountain watersheds, mineralization of soil nitrogen appears to be high only in
old-growth forests (Elwood et al , 1991)
In Canada, the influence of NO3" on episodic acidification is less universal Molot
et al (1989) and Dnscoll et al (1989a) report on numerous episodic events in 15 streams in
the Harp, Dickie, and Plastic lake watersheds Most of these events were driven by base
cation dilution, only one event was dominated by increases in NO3" concentration The
authors conclude that NO3" plays at least a small role in most episodes, and that NO3"
increases play a greater role in acidic systems than in nonacidic ones
Small increases in NO3" concentrations during hydrologic events have been recorded at
sites in a few remaining areas of North America, including northeastern Georgia (Buell and
Peters, 1988), where maximum concentrations were approximately 12 jweq/L Several
studies have reported the existence of NO3" episodes in the western United States, including
the North Cascades (Loranger and Brakke, 1988) and the Sierra Nevada (Melack and
Stoddard, 1991) In general, the maximum concentrations of NO3" observed in the West are
less than 15 jiteq/L, substantially lower than in most of the eastern United States Lakes in
the mountainous West, however, tend to be much more dilute, and, therefore, more sensitive
10-169
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to acidic deposition than in the East Thirty-nine percent of lakes in the Sierra Nevada, for
example, have ANC values less than 50 ^eq/L, as do 26% of the lakes in the Oregon
Cascades and 17% of the lakes in the North Cascades (Landers et al, 1987) Combined
f\
with base cation dilution and small concentrations of SO4 ", the NO3" increases observed
during episodes at Emerald Lake, in the Sierra Nevada, have been sufficient to drive the
ANC to zero on two occasions in the past 4 years (Williams and Melack, 1991b) Data from
the outflow at Emerald Lake in 1986 and 1987 (Figure 10-32) indicate that minimum ANC
values are coincident with maximum concentrations of NO3" and diluted base cation
concentrations. It should be noted, however, that at no time has the pH of Emerald Lake
fallen below 5 5, a level commonly considered the threshold for injury to fish populations,
and that ANC values of zero can be caused by base cation dilution alone (a natural process)
The state of episodic acidification in the Sierra Nevada (and the rest of the West) remains,
therefore, uncertain, because few data exist and the data that are available indicate ANC
depressions to a value of 0 /xeq/L, but not below
Finally, there are some areas of North America where no significant affect of NC^" on
episodic acidification has been observed Morgan and Good (1988) report data on
10 streams in the New Jersey Pine Barrens, and found mean annual NO3" greater than
1 jueq/L only in disturbed streams (in residential and agricultural watersheds) Swistock
et al (1989) and Sharpe et al (1984, 1987, 1989) reported data on episodic acidification of
several streams in the Laurel Hill area of southwestern Pennsylvania and found that NO3"
played only a minor role in stream acidification and fish kills Baird et al (1987) examined
episodic acidification during snowmelt at Cone Pond, NH, and were unable to detect any
NO3" in inlet water Cosby et al (1991) have examined 7 years of data from two streams in
Virginia, and found no evidence of NO3" episodes, NO3" concentrations are always less than
15 /teq/L in these streams Swank and Waide (1988) reported data from seven undisturbed
watersheds at the Coweeta Hydrologic Laboratory in North Carolina, where the
volume-weighted mean concentrations of NO3" were less than 1 5 |iieq/L
Some broad geographic patterns in the frequency of episodes in the United States are
now evident. Acidic episodes driven by NO3" are apparently common in the Adirondack and
Catskill Mountains of New York, especially during snowmelt, and also occur in at least some
streams in other portions of the Northeast (e g , at Hubbard Brook) Nitrate contributes on a
10-170
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42 H
Silicate (|o,mol/L)
i I i r
ANC
-------
smaller scale to episodes in Ontario, and may play some role in episodic acidification in the
western United States There is little current evidence that NO3" episodes are important in
the acid-sensitive portions of the southeastern United States outside of the Great Smoky
Mountains We have no information on the importance of NO3" in driving episodes in many
of the subregions covered by the NSS, including those that exhibited elevated NO3"
concentrations at spring base flow (e g , the Valley and Ridge Province and Mid-Atlantic
Coastal Plain), because temporally-intensive studies have not been published for these areas
As was the case with chrome acidification discussed earlier, the mere presence of NO3"
in acidic episodes should not be construed as proof that nitrogen deposition is having an
acidifying effect on surface waters, many other sources of nitrogen exist in watersheds
There is currently little direct evidence linking nitrogen deposition with those acidic episodes
that are driven by increases in NO3" concentrations, at least partially because the type of data
necessary to link deposition to stream water pulses of NO/ are extremely difficult to collect
High concentrations of NO3" during snowmelt may simply result when NO3" stored in the
snowpack during the winter months is released while the forest is still dormant The reduced
biological activity typical of the winter months creates less demand for nitrogen, and
snowpack NO3" may simply run off without entering the nitrogen cycle of the forest or
watershed Several mechanisms, however, will amplify the signal produced by atmospheric
deposition of nitrogen to snowpacks In areas with large snowpacks (e g , much of the
Northeast and all of the mountainous West), ions have been shown to drain from the pack in
the early stages of snowmelt, leading to concentrations that are much higher than the average
concentration of the snowpack itself (e g , Jeffries, 1990) This differential elution of acid
anions (like NO3~) during the initial stages of snowmelt has been shown to be responsible for
the elevated NO3" concentrations observed in parts of Scandinavia (Johannessen and
Hennksen, 1978), Canada (Jeffries, 1990), the Adirondacks (Mollitor and Raynal, 1982), the
Midwest (Cadle et al, 1984), and in the Sierra Nevada (Williams and Melack, 1991b)
Ammonium deposited to the snowpack as either wet or dry deposition can be subsequently
nitrified to NO3" in soils, or while still in the snowpack, and can produce NO3"
concentrations elevated over those calculated from NO3" deposition alone (Galloway et al,
1980; Schofield et al , 1985, Cadle et al , 1987, Schaefer and DnscoU, in press) Rates of
dry deposition of nitrogen compounds to the snowpack are difficult to measure, but
10-172
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potentially important, controls on NO3" concentrations in snowmelt water (Galloway et al,
1980, Cadle et al, 1987) Jeffries (1990) presents a recent review of snowpack storage and
release of pollutants during snowmelt
Some evidence does exist that mechanisms other than atmospheric deposition contribute
to NO3" episodes, at least on a small scale Rascher et al (1987), for example, have shown
that mineralization of organic matter in the soil during Ihe winter months, and subsequent
nitrification, contribute substantially to snowmelt NO3" concentrations at one site in the
Adirondacks Schaefer and Dnscoll (in press) have suggested that a similar phenomenon
contributes to NO3" pulses during snowmelt at 11 Adirondack lakes, and that the contribution
from mineralization is greater in low-ANC and acidic lakes Stottlemyer and Toczydlowski
(1990) also report that mineralization contributes to snowmelt NO3" at a site on the upper
peninsula of Michigan It is not currently known how widespiead this phenomenon is
Because maximum NO3" concentrations are very similar among a large number of streams,
Murdoch and Stoddard (in press b) concluded that mineralization probably does not
contribute substantially to NO3" episodes in the CatskiU Mountains due to differences in soil
quality, depth, and moisture, mineralization rates are expected to differ among watersheds,
and would produce variability in concentrations of NO3" among streams There also remains
some question of whether NO3" produced from mineralization nonetheless results from
atmospheric deposition because mineralization recycles nitrogen from leaf litter
Mineralization during the winter may simply recycle nitrogen from the leaf fall of the
previous autumn, some portion of the nitrogen incorporated into leaves in the summer
undoubtedly originates as atmospheric deposition In addition, chronic nitrogen deposition
has probably contributed to forest growth in the past (through fertilization), and nitrogen now
being mineralized may be the result of such "excess" storage of nitrogen in forest biomass
Earlier in this document (see Section 10 8 2 3) it was suggested that the seventy and
duration of NO3" episodes can be expected to increase as forests become more nitrogen
sufficient (see also Dnscoll and Schaefer, 1989, Stoddard and Murdoch, 1991) Some of the
best information on whether atmosphenc deposition is contnbuting to NO3" episodes may,
therefore, come from an examination of long-term trends in surface water NO3"
concentrations
10-173
-------
There is some evidence that the occurrence and seventy of NO3" episodes are
increasing. Smith et al (1987a) examined trends in NO3" at 383 stream locations in the
United States between 1974 and 1981, and reported increases at 167 sites, especially east of
the 100th meridian Many of the increasing trends could be attributed to increased use of
fertilizers in agricultural areas, particularly in the Midwest In addition to agricultural
runoff, Smith et al (1987a) identified atmospheric deposition as a major source of NO3" in
surface waters, particularly in forested basins of the East (e g , New England and the Mid-
Atlantic) and Upper Midwest Despite widespread use of fertilizers in most of the regions
covered by the Smith et al study, they found a high degree of correlation between stream
basin yield of NO3" and rates of nitrogen deposition
Historical data are available from 19 large streams in the Catskill Mountains, some of
which have been monitored since early in this century (Stoddard and Murdoch, 1991,
Stoddard, in review) Trend analyses indicate that NO3" concentrations have increased in all
of the streams (Table 10-23), with the majority of the increase occurring in the past two
decades (1970s and 1980s) (Murdoch and Stoddard, in press b, Stoddard, 1991) These
increases are not attributable to other anthropogenic sources of nitrogen, and are similar to
trends observed in eight headwaters streams monitored in the 1980s (Murdoch and Stoddard,
in press: a, Murdoch and Stoddard, in press b) At four historical Catskill sites where
stream discharge data are available, the relationship between NO3" concentrations and
discharge have changed over the course of the past 4 decades (Figure 10-33) In all cases,
the relationships are steeper in the 1980s than in the past, indicating that most of the increase
in NO3" has occurred at high flows (i e , episodic NO3" concentrations have increased more
than base-flow NO3" concentrations) The composite average atmospheric NO2
concentrations have been downward for the past 10 years Stream concentrations, however,
are based on nitrate deposition, not atmospheric concentrations of NO2
Trends in lake water NO3" concentrations that are similar to the Catskill stream trends
have been reported for Adirondack lakes (Dnscoll and Van Dreason, in press, Table 10-24)
Nine out of 17 Adirondack lakes exhibited significant increases m NO3" concentrations,
whereas only 1 exhibited a significant decrease (Table 10-24) It is not statistically possible
to determine whether episodic NO3" concentrations are mostly responsible for the trends in
Adirondack lakes because the data record is short (1982 to 1989) Plots of temporal NO3"
10-174
-------
TABLE 10-23. SLOPES OF NITRATE TRENDS (/teq/L/year) IN
CATSKILL STREAMS BEFORE 1945, BETWEEN 1945 AND 1970,
AND BETWEEN 1970 AND 1990. SLOPES FOR EACH PERIOD ARE
CALCULATED FROM BEST-FIT REGRESSION LINES
(ANALYSIS OF COVARIANCE ON RANKS, SEE TEXT FOR DETAILS)
FITTED TO DATA FROM THE ENTIRE PERIOD OF RECORD.
ALL TRENDS ARE SIGNIFICANT AT P LESS THAN 0.05. MEDIAN VALUES
AND SAMPLE SIZES FOR EACH PERIOD ARE GIVEN IN PARENTHESES.
[— = Data insufficient for analysis.]
Site
Batavia Kill
Bear Kill above Grand Gorge
Bear Kill above Hardenbergh Falls
Beaver Killb
Birch Creek above Pine Hill
Birch Creek at Pine Hill
Bush Kill
Bushnellville Creek
Esopus Creek above Big Indian
Esopus Creek below Big Indian
Esopus Creek at Coldbrook
Little Beaver Killb
Manor Kill
Neversmk River
Rondout Creek
Schohane Creek at Prattsville
Stony Clove Creek
West Kill
Woodland Creekb
Before 1945
+024
(11, n = 235)
-
+034
(18, n = 253)
+005
(4, n = 270)
_
-001
(11, n = 287)
+0 11
(4, n = 235)
+004
(4, n = 267)
+008
(4, n = 246)
-0 16
(7, n = 59)
+024
(7, n = 352)
+000
(4, n = 268)
-0 12
(11, n = 251)
_
_
+064
(7, n = 238)
-000
(4, n = 272)
+0 19
(7, n = 227)
+002
(4, n = 272)
Change in Nitrate Concentration
1945-1970
+021
-
(27, n = 9)
-
+010
+060
(4, n = 12)
+068
(6, n = 11)
+000
(7, n = 248)
+025
-
-001
(7, n = 64)
-008
(11, n = 784)
+001
-055
(14, n = 306)
+033
(7, n = 185)
+000
(7, n = 12)
-0 13
(14, n = 712)
+008
_
+008
Between 1970 and 1990
+028
(21, n = 70)
+070
(38,n = 92)
—
+ 176
(14, n = 10)
+268
(16, n = 75)
+073
(19 n = 63)
+228
(19, n = 94)
+ 157
(17, n = 10)
-
+ 1 98
(21, n = 93)
+200
(19, n = 886)
+085
(5, n = 10)
+097
(17, n = 96)
+ 128
(14, n = 104)
+ 1 79
(8, n = 43)
+ 1 93
(21, n = 805)
+377
(24, n = 10)
_
+395
(25, n = 10)
Data available for fewer than 2 years in one or more time periods at this site Truids were not calculated dunng these time periods at this
site, but median values and sample sizes are listed
Data for these sites are available only for periods before 1945 and from 1977 to 1979 Trends reported for the periods of missing data are
based on regression lines for the entire data set, median values cannot be listed
Source Murdoch and Stoddard (in press b)
10-175
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(a)
Schoharte Creek at Prattsvitle
01 1 10 10O
Stream Discharge (m /s)
(b)
Neversink River at Claryville
1950-59
1960-69
1 10 1OO
Stream Discharge (m /s)
Esopus Creek at Coldbrook
(d)
Rondout Creek at Lowes Corners
1 10
Stream Discharge (m /s)
Figure 10-33. Relationship between nitrate concentration and stream discharge for four
Catskill streams during four most recent decades, (a) Schoharie Creek at
Prattsville, (b) Neversink River at Claryville, (c) Rondout Creek at Lowes
Corners, and (d) Esopus Creek at Coldbrook. Regression lines for each
decade are from least-squares regression of concentration on the log of
stream discharge, and all regressions are significant (p < 0.05). All sites
indicate that nitrate concentrations at high discharges are higher in the
1970s and 1980s than in previous decades.
Source- Murdoch and Stoddard (in press b)
patterns, however, suggest that base-flow values are relatively unchanged, whereas spring
values are increasing (Figure 10-34)
A cautionary note in the interpretation of long-term nitrogen trends is introduced by
examination of long-term data from streams at the Hubbard Brook Experimental Forest
(HBEF). Data from control Watershed #6 through 1977 suggested a strongly increasing
trend in NCV (Schindler, 1987) and have been used to suggest that the HBEF watersheds are
10-176
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TABLE 10-24. TRENDS IN NITRATE CONCENTRATIONS FOR ADIRONDACK
LONG-TERM MONITORING LAKES. SLOPES ARE CALCULATED FROM
BEST-FIT REGRESSION LINES (USING ANCOVA ON RANKS)
FITTED TO DATA
Lake Name
Arbutus Lake
Barnes Lake
Big Moose Lake
Black Lake
Bubb Lake
Cascade Lake
Clear Pond
Constable Pond
Dart Lake
Heart Lake
Lake Rondaxe
Little Echo Pond
Moss Lake
Otter Pond
Squash Pond
West Pond
Windfall Lake
na
96
51
105
104
88
105
104
106
88
103
88
84
105
93
100
106
88
Change in NO3"
(/xeq/L/year)
+ 105
+ 003
+ 016
+ 004
-0 11
-050
+051
+ 1 26
+034
+088
+018
+001
000
+ 150
+ 1 14
+009
014
PC
< 00001
069
036
079
053
004
<0 0001
00003
007
< 00001
004
012
094
< 0.0001
008
056
082
aNumber of individual observations, the period of record for most sites is from June 1982 to August 1989
Slope of analysis of covanance (ANCOVA) model Positive slope indicates an increase in nitrate ions (NO3"),
negative number indicates decrease
Significance of regression coefficient for date in ANCOVA model
Source Loftis et al (1989), Dnscoll and Van Dreason (in press)
10-177
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0)
=3L
100-r
90-
80-
70-
60-
50-
40-
30--
20--
10-,-
0
(a) Constable Pond
Trend in All Data
Trend in Spring Data
|
co
60
50
40
30
20
10
0
(b) Heart Lake
Trend in All Data
Trend in Spring Data
1982 1983 1984 1985 1986 1987 1988 1989
Figure 10-34. Temporal patterns in lake water nitrate concentration for two
Adirondack lakes: (a) Constable Pond, and (b) Heart Lake. Both sites
exhibit increasing trends in nitrate ion (Table 10-24). The strongly
seasonal behavior of nitrate hi these lakes suggests that most of the
increase has occurred in spring episodic nitrate concentrations.
Source Dnscoll and Van Dreason (in press)
undergoing nitrogen saturation (Agren and Bosatta, 1988) Examination of the entire 23-year
record (1965 to 1983) from Watershed #6, however, shows no long-term trend (Likens,
1985; Dnscoll et al, 1989a) and emphasizes the importance of examining nitrogen processes
10-178
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in a truly long-term context Pools of nitrogen associated with soils and forests at HBEF,
and elsewhere, are very large (ca 340,000 mol/ha at HBEF, up to 520,000 mol/ha at other
sites in the eastern United States, Federer et al, 1989) and long-lived (the turnover rate for
nitrogen at HBEF is estimated at 80 years), small changes in the long-term cycling of
nitrogen within this system will have profound effects on stream water chemistry (Dnscoll
et al , 1989a) Although the data reported here for the Catsktlls can be considered truly
long-term (up to 65 years of record), data for the Adirondacks (Dnscoll and Van Dreason, in
press) and other areas of the United States (Smith et al, 1987a) span only 1 to 2 decades,
and should be interpreted with caution
Many of the data discussed above suggest that NO3" episodes are more severe now than
they were in the past These surface water nitrogen increases have occurred at a tune when
nitrogen deposition has been relatively unchanged in the northeastern United States (Husar,
1986, Simpson and Olsen, 1990, Bowersox et al, 1990) If we accept the idea that an
increase in the occurrence of NO3" episodes is evidence that nitrogen saturation of watersheds
is progressing, then current data suggest that current levels of nitrogen deposition
(5 to 10 kg/ha/year) are too high the for the long-term health of aquatic systems in the
Adirondacks, the Catskills, and possibly elsewhere in the Northeast It is important to note
that this supposition is dependent on our acceptance of NO3" episodes as evidence of nitrogen
saturation At this point, no measurements of changes in nitrogen cycling have been made to
support this
Similar logic would suggest that levels of nitrogen deposition in the Sierra Nevada
(ca 2 kg/ha/year) may be at the upper limit of the levels that would be protective of the
long-term health of sensitive, high elevation aquatic systems in the West The discrepancy
between the levels of nitrogen deposition that produce signs of nitrogen saturation in the
Northeast and the West is a good illustration of the need to set deposition levels in terms of a
"critical load" to specific systems The deposition levels measured in the eastern and
western United States are within the range of nitrogen critical loads (3 to 14 kg/ha/year)
suggested by European work in regions of silicate soils of varying sensitivity (Schulze et al,
1989) The Northeast, because of deeper soils and aggrading forests, may be able to absorb
higher rates of deposition without serious damage than areas of the mountainous West, where
soils are thin and forests are often absent The abilities of these regions to absorb nitrogen is
10-179
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a function of the capacities of their watersheds to retain nitrogen Because these capacities
differ from one region to another, the critical loads of nitrogen that will produce signs of
degradation also vary from region to region These differences are at the heart of the critical
loads concept of setting deposition limits
10.8.4 The Effects of Nitrogen Deposition on Eutrophication
The term "eutrophy" generally refers to a state of nutrient enrichment (Wetzel, 1983),
but is commonly used to refer to conditions of increased algal biomass and productivity,
presence of nuisance algal populations, and a decrease in oxygen availability for
heterotrophic organisms Eutrophication is the process whereby lakes, estuaries, and marine
systems progress toward a state of eutrophy In lakes, eutrophication is often considered to
be a natural process, progressing gradually over the long-term evolution of lakes The
process can be significantly accelerated by the additional input of nutrients from
anthropogenic sources The subject of eutrophication has been extensively reviewed by
Hutchinson (1973), the National Research Council (1969), and Likens (1972)
Establishing a link between nitrogen deposition and the eutrophication of aquatic
systems depends on a determination of two key conditions The first condition is that the
productivity of the system is limited by nitrogen availability Our current concept of nutrient
limitation stems from Liebig's Law of the Minimum (Von Liebig, 1840), which can be
paraphrased to suggest that, at any single point in time, ecosystem productivity will be
limited by whatever necessary environmental element is in shortest supply When that
necessary environmental element is nitrogen, then the system can be said to be nitrogen
limited. The second condition is that nitrogen deposition be a major source of nitrogen to
the system In many cases, the supply of nitrogen from deposition is minor when compared
to other anthropogenic sources, such as pollution from either point or nonpoint sources
10.8.4.1 Freshwater Eutrophication
It is generally accepted that the productivity of fresh waters is limited by the availability
of phosphorus, rather than the availability of nitrogen (reviewed by Hecky and Kilham,
1988). Although conditions of nitrogen limitation do occur in freshwater systems (discussed
below), they are often either transitory, or the result of high inputs of phosphorus from
10-180
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anthropogenic sources At high rates of phosphorus input, phosphorus will cease to be in
short supply, and whatever nutnent is then least abundant (often nitrogen) will become
limiting Although additions of nitrogen from deposition will lead to increased productivity
in these situations, the primary dysfunction is an excess supply of phosphorus, and these
situations will not be discussed further Often when nitrogen limitation does occur, it is a
short-lived phenomenon because nitrogen-deficient conditions favor the growth of blue-green
algae (e g , Smith, 1982), many of which are capable of nitrogen fixation Because
nitrogen-fixing species are not limited by the availability of fixed nitrogen (e g , NH4 ,
NO3"), they may thrive under conditions where other species are nitrogen limited, and
effectively increase rates of nitrogen input to the system by fixation of gaseous nitrogen
High rates of nitrogen fixation may lead to situations where nitrogen can no longer be said to
be limiting, and the system often returns to a state of phosphorus limitation In lakes,
nitrogen fixation may be considered a natural mechanism that compensates for deficiencies in
nitrogen, and contributes to the long-term evolution and ubiquity of phosphorus limitation
(Schindler, 1977)
Nitrogen limitation can occur naturally (i e , in the absence of anthropogenic
phosphorus inputs) in lakes with very low concentrations of both nitrogen and phosphorus, as
are common in the western United States and in the Northeast (Suttle and Harrison, 1988)
Suttle and Harrison (1988) and Stockner and Shortreed (1988) have suggested that
phosphorus concentrations are too low in these systems to allow blue-green algae to thrive
because they are poor competitors for phosphorus at very low concentrations (e g , Schindler
et al, 1980, Smith and Kalff, 1982) Thus, diatom communities dominate phytoplankton
and penphyton communities in these extremely nutrient-poor (ultraoligotrophic) systems, and
rates of nitrogen fixation do not increase because blue-green algae do not become
established, regardless of relative nitrogen or phosphorus deficiency In these systems, the
two nutrients are often closely coupled and constant shifts between nitrogen and phosphorus
deficiency may occur without obvious changes in community structure In these situations,
additional loading of nitrogen from anthropogenic deposition is likely to have only a small
effect on primary productivity because the system quickly becomes phosphorus limited In a
literature survey of 62 separate nutnent limitation studies in lakes, Elser et al (1990) found
that simultaneous additions of nitrogen and phosphorus produced the largest growth response
10-181
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in 82% of the experiments These results underline the likelihood that a lake limited by one
nutrient may quickly become limited by another if the lake becomes enriched with the
original limiting nutrient
Estimations of nutrient limitation in lake ecosystems follow three major lines of
reasoning' (1) evidence from ambient nutrient concentrations and the nutritional needs of
algae, (2) evidence from bioassay experiments at various scales, and (3) evidence from
nutrient dynamics and input/output studies (Hecky and Kilham, 1988, Howarth, 1988)
Much of the acceptance of the idea that freshwater lakes are primarily phosphorus
limited stems from the close correlations between phosphorus concentrations and lake
productivity or algal biomass (usually measured as chlorophyll concentration) that have been
observed m a large number of lake studies (e g , Dillon and Rigler, 1974, Schindler, 1977,
1978, reviewed by Reynolds, 1984, Peters, 1986) More recently, researchers have begun to
question the ubiquity of the phosphorus chlorophyll relationship, and to identify some of the
factors that lead to the large variability observed in this relationship in nature (e g , Smith
and Shapiro, 1981, Smith, 1982, Pace, 1984, Hoyer and Jones, 1983, Prairie et al , 1989)
Notably, researchers have found that the relationship is not linear, as previously supposed,
but sigmoidal (McCauley et al, 1989), and that the slope of the relationship is significantly
affected by nitrogen concentrations, particularly at high concentrations of phosphorus
(> 10 /ieq/L) that are likely to be caused by anthropogenic inputs McCauley et al (1989)
found that nitrogen had little effect on the phosphorus chlorophyll relationship at low
concentrations of phosphorus This effect is expected in nutrient-poor lakes, where the
primary effect of nitrogen additions would be to push lakes into a phosphorus-deficient
condition
Arguments based on ambient nutrient concentrations stem from the early work of
Redfield (1934), who examined the concentrations of nutrients within the cells of nutrient-
sufficient algae from marine systems worldwide, and found surprisingly consistent results for
the ratio of carbon to nitrogen to phosphorus concentrations (106 16 1), deviations from
these ratios are taken to be evidence that one nutrient or another is limiting to algal growth
(e.g., nitrogen: phosphorus [N P] ratio values below 16 1 suggest nitrogen limitation, values
above 16:1 suggest phosphorus limitation) Because the relative supply rates of phosphorus
and nitrogen will determine whether one or the other nutrient is in short supply, it has been
10-182
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suggested that the ratio of the two nutrients (i e , total nitrogen total phosphorus) can be used
as an index of nutrient limitation (Chiaudam and Vighi, 1974, Rhee, 1978, Schindler, 1976,
1977, 1978) Various researchers have extended interpretation of the Redfield ratio to
include ambient nutrient concentrations ui water (Redfield's original work was with
intracellular concentrations), and applied the nutrient ratio criteria to waters supplying lakes
to determine the likely limiting conditions that these waters will produce (e g , Schindler,
1977, Smith and Shapiro, 1981, Frame et al, 1989) This method has the potential to
illustrate regional patterns and has gamed some support from the results of bioassay
experiments (see below) This idea has been refined recently to exclude from the ratio those
forms of nitrogen and phosphorus that are not biologically available (e g , especially organic
forms of nitrogen), with the result that good predictions of nutrient limitation can now be
made from ratios of total dissolved inorganic nitrogen (DIN) to total phosphorus (TP)
(Morns and Lewis, 1988)
Moms and Lewis (1988) conducted nutrient addition bioassays on natural assemblages
of phytoplankton from many lakes, and compared their results to DIN TP values measured in
the lakes at the same tune as the experiments were conducted They found that lakes with
DIN TP values less than 9 (using molar concentrations) could be limited by either nitrogen
or phosphorus (often additions of both nutrients were required to stimulate growth), whereas
lakes with DIN TP values less than 2 were always limited by nitrogen The discrepancy
between the 16 1 Redfield ratio and the 2 1 ratio suggested by Morns and Lewis (1988) may
result from measuring ambient, rather than cellular, nutnent concentrations and from the
variety of critical nitrogen phosphorus (N P) ratios exhibited by different species in nature
(Suttle and Harrison, 1988)
If a critical DIN TP value less than 2 is applied to lakes from the Eastern Lake Survey
(Linthurst et al, 1986) and Western Lake Survey (Landers et al, 1987), it is possible to
estimate the number of nitrogen-limited lakes in some regions of the United States
(Table 10-25) Lakes with total phosphorus concentrations greater than 2 0 /*eq/L have been
excluded from this analysis because many of them may have expenenced anthropogenic
inputs of phosphorus (Vollenweider, 1968, Wetzel, 1983) This test is, therefore, a
conservative one for nitrogen limitation, both because the DIN TP value chosen (< 2) is a
conservative measure of nitrogen limitation (Morns and Lewis, 1988) and because some
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TABLE 10-25. ESTIMATED NUMBER AND PROPORTION OF NITROGEN-
LIMITED LAKES IN SUBREGIONS OF THE UNITED STATES SAMPLED BY THE
NATIONAL SURFACE WATER SURVEY. ESTIMATES ARE BASED ON MOLAR
RATIOS OF TOTAL INORGANIC NITROGEN CONCENTRATIONS (NITRATE +
AMMONIUM) TO TOTAL PHOSPHORUS CONCENTRATIONS
Subregion
Eastern Lake Surveya
Adirondacks (1A)
Poconos/CatsMls (IB)
Central New England (1C)
Southern New England (ID)
Northern New England (IE)
Northeastern Minnesota (2A)
Upper Peninsula, Michigan
(2B)
Northcentral Wisconsin (2C)
Upper Great Lakes Area (2D)
Southern Blue Ridge (3A)
Florida (3B)
Western Lake Surveyb
California (4A)
Pacific Northwest (4B)
Northern Rockies (4C)
Central Rockies (4D)
Southern Rockies (4E)
Number of
Lakes in
Subregion
1,684
1,986
2,003
2,667
2,388
2,132
1,698
1,707
6,147
538
8,053
2,806
2,200
3,335
2,970
2,195
Estimated Number Proportion
of Nitrogen- of Population
Limited Lakes Nitrogen-Limited (%)
164
2285
549
144.7
91 3
3162
305 8
2482
13454
115
25
535 8
609 1
7399
7887
4552
10
11 5
27
54
3 8
148
180
145
21 9
2 1
00
19 1
277
222
266
207
"Data from Kanciruk et al (1986), excluding lakes with total phosphorus > 2 /-imol/L
Data from Eilers et al. (1987), excluding lakes with total phosphorus > 2 /nmol/L
lakes with naturally high concentrations of phosphorus may be excluded, these lakes are
more likely to be nitrogen-limited than lakes with low phosphorus concentrations The
proportions of lakes that can be considered nitrogen-lunited vary widely from region to
region, with the greatest number being found, as expected, in the West The highest
10-184
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proportion was found in the Pacific Northwest (27 7% of lakes exhibited low DIN TP
ratios), but all subregions of the West contained substantial numbers of nitrogen-limited
lakes The smallest proportions were found in the Southeast (2 5 % of the lakes in the entire
region exhibited low DIN TP ratios) and the Northeast (5 %) One surprise in this analysis is
the number of lakes in the Upper Midwest that appear to be nitrogen-limited, taken as a
whole, this region had 19% of its lakes with DIN TP ratios less than 1
A more direct indication of nutrient limitation than is available from nutrient ratios can
be gained from bioassay experiments, where a small volume of natural lake water is enclosed
and various known concentrations of potentially limiting nutrients are added (e g , Melack
et al , 1982, Setaro and Melack, 1984, Stoddard, 1987b) A growth response (usually
measured as an increase in biomass) in treatments containing an added nutrient constitutes
evidence of limitation by that nutrient The results of such experiments are available for
only a few selected nutrient-poor lakes, however, and indicate a variety of responses
including strong phosphorus limitation (Melack et al, 1987), limitation by phosphorus and
Fe (Stoddard, 1987b), simultaneous nitrogen and phosphorus limitation (i e , the two
nutrients are so closely balanced that addition of one alone simply leads to limitation by the
other, Gerhart and Likens, 1975, Suttle and Harrison, 1988, Dodds and Pnscu, 1990), and
limitation primarily by nitrogen (Morns and Lewis, 1988, Goldman, 1988) No clear pattern
of nitrogen or phosphorus limitation develops from an examination of these few studies
The potential for nitrogen deposition to contribute to the eutrophication of freshwater
lakes is probably quite limited Eutrophication by nitrogen inputs will only be a concern in
lakes that are chronically nitrogen limited This condition occurs in some lakes that receive
substantial inputs of anthropogenic phosphorus, and in many lakes where both phosphorus
and nitrogen are found in low concentrations (e g , Table 10-25) In the former case, the
primary dysfunction of the lakes is an excess supply of phosphorus, and controlling nitrogen
deposition would be an ineffective method of water quality improvement In the latter case,
the potential for eutrophication by nitrogen addition (e g , from deposition) is limited by low
phosphorus concentrations, additions of nitrogen to these systems would soon lead to
nitrogen-sufficient, and phosphorus-deficient, conditions Increases in nitrogen deposition to
some of the regions in Table 10-24 would probably lead to measurable increases in algal
10-185
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biomass in those lakes with low DIN TP ratios and substantial total phosphorus
concentrations, but the number of lakes that meet these criteria is likely to be quite small
10.8.4.2 Estuaries and Coastal Waters
Estuarine and coastal water ecosystems exist at the transition between freshwater
systems and the open ocean These transition zones share some characteristics with both
freshwater and marine systems, but they also have some unique properties that lead to
different responses to NOX deposition and a correspondingly different set of concerns They
are at the end of a long series of nitrogen transport and transformation processes involving
interactions with vegetation, soils, groundwater, small streams, lakes, and rivers At each
step in this series, the processes vary temporally and spatially and may be subject to a variety
of human influences This transition zone integrates complex and fluctuating processes that
are distributed over what are sometimes very large watersheds
The transition zones between fresh- and saltwater systems are subject to natural
processes that are not observed elsewhere in aquatic systems, such as tidal flows and salinity
changes They are also subject to substantial human influence Estuaries provided natural
ports and are among the most productive ecosystems on the planet (Begon et al , 1986)
Tims, they became an obvious location for cities, with accompanying demands for
wastewater disposal The history of human use of estuaries and lands around estuaries make
it more difficult to isolate the effects of a particular anthropogenic contaminant on ecosystem
characteristics. The conservative approach used above to assess the impact of nitrogen
deposition on freshwater eutrophication (excluding all systems with anthropogenic impacts
other than atmospheric deposition) is not possible for estuaries and coastal waters, all
estuanne systems, and most coastal waters, have been subjected to human impacts, often for
several centuries
Estuaries are bodies of water, more or less isolated from the rest of the ocean, where
fresh water and salt water mix This generally produces a salinity gradient, and often leads
to stratification of water, with the heavier salt water below a layer of fresh water Estuaries
are also subject to tidal effects and may be strongly influenced by river flows
In combination, these forces tend to produce quite complex water circulation patterns with
significant biological consequences For example, water currents within Chesapeake Bay
10-186
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concentrate and circulate the dinoflagellate Gyrodmium uncatenum, which is responsible for
red tides in that estuary (Tyler et al , 1982) Circulation patterns within estuaries may also
influence patterns of habitat use by fish (e g , Pietrafesa et al, 1986)
Boynton et al (1982) described a classification of estuaries into four categories that
were designed to reflect the primary factors influencing algal production and the variability
that exists among estuaries
• Fjords have deep basin waters and shallow underwater sills connecting them with the
sea, providing slow exchange with adjacent sea waters,
• Lagoons are shallow, well-mixed, slowly flushed, and only slightly influenced by
riverine inputs,
• Embayments are deeper than lagoons, often stratified, only slightly influenced by
freshwater inputs, and have good exchange with the ocean, and
• River-Dominated Estuaries are a more diverse group of systems, all of which exhibit
seasonally depressed salinities due to riverine inputs and variable degrees of
stratification
The physical and chemical structure of estuaries will strongly shape the movement and
transformation of nitrogen compounds Aston (1980) has provided a list of features of
estuaries that have a controlling influence on the geochemistry of contaminants and nutnents
(1) The tidal mixing of fresh and sea waters on a semidiurnal or diurnal tune scale, with
corresponding changes in the volume of water in an estuary, produces temporal changes
in the contributions of nutnents and dissolved gases from marine and freshwater
sources For example, estuaries are generally enriched in nutnents relative to ocean
waters due to the local influences of land drainage and often pollution
(2) The circulation, and especially the stratification, of some estuaries can create vertical
and horizontal variations of the concentrations of nutnents and dissolved gases within
an estuary
(3) Estuanne topography may give nse to particularly restncted circulations (e g , in
fjords, where the mixing of external sea water with the estuanne waters is greatly
reduced), and the restncted mixing leads to unuseal chemical environments (e g ,
oxygen-deficient waters)
(4) The circulation patterns in coastal waters and estuaries lead to the deposition of various
types of sedimentary material The deposition and resuspension of sediments may
10-187
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influence the budgets of dissolved constituents, including nutrients and gases, in
estuarine waters
(5) Chemical reactions occurring during the mixing of river water with sea water may lead
to the removal or addition of the dissolved nutrients Also, the changes in temperature
and salinity during estuarine mixing influence the solubility of dissolved gases, and thus
influence their removal or addition in an estuary
(6) Biological production and metabolism have significant influences on the occurrence and
distribution of nutrients and some gases (e g , CO2 and oxygen) in estuarine waters
The biological communities in estuaries tend to be species-poor because few species are
able to tolerate the extremes in environment to which they are exposed What species
do thrive, however, are often productive
In fact, estuaries may be extremely productive Fisheries yields in estuaries are higher
per unit area than in lakes (Nixon, 1988) This appears not to be related to primary
production, but rather to the efficiency of utilization of the primary production The input of
nutrients from outside the ecosystem may be a major determinant of overall fisheries
production levels (Day et al, 1982) The economic importance of estuaries may be simply
indicated by McHugh's (1976) estimate that in 1970, 69% (by weight) of fish landings in the
United States were estuary dependent
Estuaries and coastal waters receive substantial amounts of weathered material (and
anthropogenic inputs) from terrestrial ecosystems and from exchange with sea water As a
result, they tend to be very well buffered, acidification is not a concern in any of these areas
The same load of weathered material and anthropogenic inputs that makes estuaries and
coastal areas insensitive to acidification, however, makes them very prone to the effects of
eutrophication Eutrophication of these areas has some very specific and damaging
consequences, especially the creation of anoxic bottom waters, blooms of nuisance algae, and
replacement of economically important species by less-desirable ones (e g , Mearns et al ,
1982; Jaworski, 1981) Eutrophication, for example, has been suggested as the causal factor
in the disappearance of the striped bass (Morone saxatths) fishery in Chesapeake Bay (Price
et al, 1985), the increasing spatial extent of anoxic bottom waters during the summer season
is the proposed mechanism (e g , Officer et al, 1984) Anoxia is also thought to have had
disastrous effects on surf clams (Spisula solidissimd) in the New York Bight (Swanson and
Parker, 1988) and the blue crab (Callmectes sapidus) habitat in Chesapeake Bay (Officer
et al., 1984). In 1971, blooms of the red tide dinoflagellate Ptychodiscus brevis in the Gulf
10-188
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of Mexico were responsible for the deaths of approximately 100 tons of fish daily, the high
nutrient concentrations typical of eutrophic conditions have been linked to many blooms of
nuisance algae (Paerl, 1988)
Establishing a link between nitrogen deposition and the eutrophication of estuaries and
coastal waters depends on a determination (as it does in fresh water—see above) of two key
conditions The first condition is that the productivity of these systems is limited by nitrogen
availability The second condition is that nitrogen deposition be a major source of nitrogen
to the system In many cases, the supply of nitrogen fiom deposition is minor when
compared to other anthropogenic sources, such as pollution from either point or nonpomt
sources
Few topics in aquatic biology have received as much attention in the past decade as the
debate over whether estuanne and coastal ecosystems are limited by nitrogen, phosphorus, or
some other factor (reviewed by Hecky and Kilham, 1988) In a seminal paper published in
1971, Ryther and Dunstan (1971) used evidence of ambient nutrient concentrations and the
results of bioassay experiments to conclude that nitrogen limited the productivity of waters
along the south shore of Long Island and in the New York Bight They noted that, during
blooms of algae in these areas, inorganic nitrogen concentrations often decreased to levels
below detection, whereas inorganic concentrations of phosphorus remained high From this
evidence, they deduced that phosphorus could not be a limiting factor, but that nitrogen could
be They conducted bioassay experiments, suspending in small bottles single-species cultures
of either Nannochlons atomus or Skelatonema costatum, the two algal species that were
dominant in the blooms in each location, in filtered sea water with additions of either
ammonium or phosphorus Ryther and Dunstan (1971) found that both species increased
dramatically in ammonium-enriched bottles, but that phosphorus-enriched bottles were no
different than controls, and that this response was consistent at a large number of sites
throughout the south shore of Long Island and in the New York Bight They concluded that
"nitrogen is the critical limiting factor to algal growth and eutiophication in coastal marine
waters" (Ryther and Dustan, 1971)
Since the publication of this influential paper, many researchers have accepted the
notion that coastal waters and estuaries are limited primarily by nitrogen (e g , Boynton
et al, 1982, Nixon and Pilson, 1983), to the point where nitrogen limitation in marine
10-189
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waters, and phosphorus limitation in fresh waters, has become near dogma (Hecky and
Kilham, 1988). More recently, some oceanographers have begun to question the ubiquity of
nitrogen-limitation in estuanne and coastal marine waters (e g , Smith, 1984, Howarth,
1988), and it seems clear that evidence for nutrient limitation in these systems must be
analyzed on a case-by-case basis Experiments to confirm widespread nitrogen limitation in
estuaries have not been conducted, and nitrogen limitation cannot be assumed to be the rule
(Hecky and Kilham, 1988)
Estimations of nutrient limitation in estuaries and coastal marine ecosystems follow the
same three major lines of reasoning as arguments about freshwater nutrient limitation (see
Section 10.8 4 1). (1) evidence from ambient nutrient concentrations and the nutritional
needs of algae, (2) evidence from bioassay experiments at various scales, and (3) evidence
from nutrient dynamics and input/output studies (Hecky and Kilham, 1988, Howarth, 1988)
As explained earlier, arguments based on ambient nutrient concentrations stem from the
early work of Redfield (1934), who examined the concentrations of nutrients within the cells
of nutrient-sufficient algae from marine systems worldwide, and found surprisingly consistent
results for the ratio of carbon to nitrogen to phosphorus concentrations (106 16 1, using
molar concentrations), deviations from these ratios are taken to be evidence that one nutrient
or another is limiting to algal growth (e g , molar N P ratio values below 16 1 suggest
nitrogen limitation, values above 16 1 suggest phosphorus limitation) Various researchers
have extended interpretation of the Redfield ratio to include ambient nutnent concentrations
in water (Redfield's original work was with intracellular concentrations), and applied the
nutrient ratio criteria to waters supplying estuaries and coastal systems to determine the likely
limiting conditions that these waters will produce (e g , Ryther and Dunstan, 1971, Jaworski,
1981). The biotic response (i e , biostimulation) is not measured using this approach, but is
instead inferred from geochemical principles, in this sense, the nutrient-ratio approach
measures potential nutnent limitation rather than actual limitation Boynton et al (1982)
summarized nutrient ratio information for a number of estuanne systems, these results are
repeated in Table 10-26 At the time of maximum primary productivity, a majonty of the
estuaries they surveyed (22 out of 27) had N P ratios well below the Redfield ratio and may
have been nitrogen limited
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TABLE 10-26. MOLAR RATIOS OF DISSOLVED INORGANIC NITROGEN
TO DISSOLVED INORGANIC PHOSPHORUS IN A VARDZTY OF
ESTUARIES*
Estuary
Pamlico River, NC
Roskeeda Bay, Ireland
Narragansett Bay, RI
Bedford Basin, Nova Scotia
Beaufort Sound, NC
Chincoteague Bay, MD
Western Wadden Sea, Netherlands
Eastern Wadden Sea, Netherlands
Peconic Bay, NY
Mid-Patuxent River, MD
Southeastern Kaneohe Bay, HI
St Margarets Bay, Nova Scotia
Central Kaneohe Bay, HI
Long Island Sound, NY
Lower San Francisco Bay, CA
Upper San Francisco Bay, CA
Baratana Bay, LA
Victoria Harbor, Bntsh Columbia
Mid-Chesapeake Bay, MD
Duwamish River, WA
Upper Patuxent River, MD
Baltic Sea
Loch Etive, Scotland
Hudson River, NY
Vostock Bay, USSR
Apalachicola Bay, FL
High Venice Lagoon, Italy
DIN DIP Ratio at
Time of Maximum
Productivity
02
03
05
08
10
12
13
15
15
1 8
20
02
28
3 9
60
60
62
62
76
85
92
15
Redfield Ratio N.P = 161
18
20
20
20
48
Annual Range
in DIN DIP Ratio
0-3
0-1
05-14
05-8
05-16
1-10
1 3-120
15-56
1-4
1 8-53
Not reported
1-7
Not reported
1-6
45-85
05-16
6-16
6-15
7-225
8-16
9-61
Not reported
12-125
16-30
5-22
5-22
48-190
aDIN = Dissolved inorganic nitrogen, DIP = Dissolved inorganic phosphorus
N P = Nitrogen phosphorus
Source Boynton et al (1982)
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The data in Table 10-26, as well as from many other studies, suggest that N P ratios
vary widely within a single system from season to season D'Elia et al (1986), for example,
report ratios for the Patuxent River estuary that vary from over 20 1 during the winter to less
than 1:1 dunng the summer This variability suggests that estuanne algae may be limited by
different nutrients at different seasons
The ambient nutrient ratio approach has been criticized widely because it ignores
several factors known to be important to algal growth The use of only inorganic nutrient
species in the ratios, for example, has been criticized because many algal species are known
to utilize organic forms, especially of phosphorus (Howarth, 1988), the nutrient ratios listed
for freshwater systems (see freshwater eutrophication section, above) were based on
concentrations of total inorganic nitrogen and total phosphorus because these are thought to
be better estimators of the nutrient species actually available to algae (Morns and Lewis,
1988) Algal growth may also be more dependent on the supply rates of nutrients than on
their ambient concentrations (Goldman and Gilbert, 1982, Healey, 1973), many species of
algae may, therefore, not be limited by nutrients whose ambient concentrations are so low as
to be undetectable Broecker and Peng (1982) have echoed the earlier conclusions of
Redfield himself (1958) in pointing out that biologically mediated nitrogen fixation, and loss
rates of nitrogen from the surface waters of marine ecosystems, interact with terrestrial
nutrient inputs and tend to push the N P ratio in the particulate (i e , living) fraction of water
toward a "geochemicaUy balanced" ratio (i e , the Redfield ratio of 16 l[see
Section 10 8 4 1]) Thus ratios within the biologically active portion of the ecosystem
(particularly the algae) may approach 16 1 despite much lower ratios in the abiotic portion of
the ecosystem Taken as a whole, the evidence for nitrogen limitation from ambient nutrient
concentrations in estuaries and coastal waters must be considered equivocal
A second, and more direct, line of evidence for nutrient limitation in estuaries and
coastal waters comes from bioassay experiments These experiments have been conducted in
both freshwater and marine systems at a number of scales from small single-species cultures
(Level I experiments), to small enclosures of natural algal assemblages (Level n), to
intermediate-sized enclosures (mesocosms) of natural assemblages (Level HI), to
whole-system (so far largely limited to whole lakes) treatments (Level IV, levels as defined
by Hecky and Kilham, 1988) These experiments, therefore, progress along a gradient of
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"naturalness" from studies substantially different from the real world (Level I) to those that
simulate natural conditions very closely (Levels in and IV) Interpretation of the results of
these experiments, therefore, follows the same gradient, with more confidence being placed
in the results of studies at the upper (i e , more natural) end of the gradient (Hecky and
Kilham, 1988, Howarth, 1988) The results of Level I experiments on single-species
cultures of algae, like the original experiments of Ryther and Dunstan (1971), are especially
difficult to interpret because threshold N P ratios for individual species are known to vary
substantially Sutfle and Harrison (1988) report limitation at ratios from 7 1 to 45'1 for
single species At all scales, the experimental procedure used for experimental nutrient
additions is fairly similar, with various nutrients being added either alone or in combination,
and the growth in treated enclosures being compared to growth in control enclosures
Level I and Level n experiments have been conducted in a wide variety of estuaries and
coastal waters (e g , Thomas, 1970, Ryther and Dunstan, 1971, Vince and Vakela, 1973,
Smayda, 1974, Goldman, 1976, Graneli, 1978) and often suggest nitrogen limitation Two
studies have suggested seasonal changes from nitrogen limitation to phosphorus limitation
(D'Eha et al , 1986, McComb et al , 1981), in both cases, nitrogen-deficient conditions were
found during the peak of annual productivity in the summer The results of experiments at
Levels I and n suggest that nitrogen limitation is at least a common, if not ubiquitous,
phenomenon in coastal and estuanne waters This interpretation has been challenged by
Smith (1984) and Hecky and Kilham (1988) because the experiments were conducted at such
an unrealistic spatial scale In particular, Level I and n experiments measured only the
short-term response of algae present at the tune the experiments were run, they did not allow
natural mechanisms such as species replacement and nitiogen fixation to take place
Only a few examples of Level m bioassays exist for estuanne and coastal ecosystems
The best known of these have been conducted at the Marine Ecosystem Research Laboratory
(MERL) at the University of Rhode Island The MERL tanks are large (13-m3), relatively
deep (5-m) cylinders, with natural sediments and filtered seawater inputs They are designed
to mimic the environment of Narragansett Bay, including the mixing, flushing, temperature,
and hght regimes (Nixon et al, 1984) In the original experiments conducted in the MERL
tanks, nutrients were added with ratios that matched those of sewage entering Narragansett
Bay, but at concentrations that ranged from 1 to 32 tunes those in the bay itself, the
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experiments were run for 28 mo Algal abundance, primarily diatoms, increased with the
level of nutrient enrichment, but not on a 1 1 basis Productivity increased only by a factor
of 3 5 in the 32-time treatment, suggesting that something other than nutrients was limiting
for at least a portion of the experiment (Oviatt et al, 1986) Oviatt et al (1989) have
suggested that, in treatments with high levels of nutrient enrichment, grazing by zooplankton
controlled algal abundances to low levels, and that the upper limit to productivity was set by
self-shading in the algal community Further experiments conducted with varying nutrient
ratios suggested that diatoms in the low-nutrient (one-tune) treatments were limited by silica,
and not by either nitrogen or phosphorus (Doenng et al, 1989) Sewage inputs to many
estuaries, including Narragansett Bay, are deficient in silica (Officer and Ryther, 1980), and
silica concentrations often fall to very low levels during winter diatom blooms in this area
(Pratt, 1965) Taken as a whole, the results of the MERL experiments suggest a complex
picture for Narragansett Bay, where no nutrient is strongly limiting to algal biomass through
much of the year, and where algal abundances during winter blooms are controlled ultimately
by the concentrations of silica
In another Level HI bioassay experiment, D'Eka et al (1986) simulated the
3
environment of the Patuxent River estuary, a tributary to Chesapeake Bay, in 0 5-m
enclosures Their results had a strong seasonal component Supplements of nitrogen, either
as NCV or as NH4+, stimulated growth during the low-flow, late-summer season This
corresponds to the tune period when N P ratios in the estuary are low (1 1 or lower)
Phosphorus additions stimulated growth during the late-winter, high-flow season, when N P
ratios typically exceed 20 1. Peaks in algal abundance occurred in the summer, when anoxic
conditions in bottom waters in Chesapeake Bay are common, and when algae appear to be
nitrogen-deficient
Thus far, only one Level IV experiment has been conducted in estuaruie waters, and
only preliminary results are available Sewage treatments supplying nutrients to the
Himmerfjard basin, a brackish fjord in the Stockholm archipelago on the eastern coast of
Sweden, have been deliberately altered to produce varying levels of phosphorus and nitrogen
loads since 1983 (Graneli et al, 1990) Between 1983 and 1985, phosphorus removal at the
plant was deliberately reduced to produce a 10-fold increase in orthophosphate, and
additional sewage inputs were routed into the basin to increase total nitrogen inputs by 30 to
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40 % At the same tune as nutrient manipulations were being earned out, measurements
were made of nitrogen cycling in the basin, and algal bioassays were conducted to determine
nutrient limitation Preliminary results suggest that nitrogen is limiting at low nutrient
concentrations (i e , typical of near-coastal regions unaffected by anthropogenic inputs), and
that limiting nutrients in areas affected by anthropogenic inputs are determined by the supply
ratios of nitrogen and phosphorus (Graneli et al , 1990) Because small changes in the
supply of either phosphorus or nitrogen in the Himmerfjard basin have caused changes in the
identity of the limiting nutrient (i e , increases in phosphorus quickly lead to nitrogen
limitation, and vice versa), the authors suggest that management of both nitrogen and
phosphorus is necessary to reduce eutrophication in the basin
The remaining line of evidence used to infer nutrient limitation in estuanne and coastal
marine ecosystems comes from studies of nutrient dynamics, and especially of input/output
budgets In many ways, the results of these studies help to integrate the sometimes
contradictory results gleaned from studies of nutrient ratios and bioassay experiments at
different levels of complexity Smith (1984) summarized the studies conducted on four
subtropical bays and concluded that phosphorus is more likely to be limiting in these systems
than nitrogen, and that physical factors are often more important than either nutrient Smith
noted that in the systems that had high throughputs of water (i e , embayments according to
the Boynton et al [1982] criteria, see earlier description), incoming ratios of nutrients were
matched very closely by the ratios in the outgoing water This suggests that algal growth is
having little effect on nutrient levels, and that nutrients do not limit productivity In systems
that flush more slowly (i e , lagoons or fjords in the Boynton et al [1982] classification),
any deficiencies in nitrogen in the incoming water can be made up by nitrogen fixation on
the ocean bottom, and phosphorus is, therefore, more Likely to be limiting
The question of why nitrogen deficiencies in marine systems are not simply made up by
nitrogen fixation, as suggested by Smith (1984), is central to the issue of whether estuaries
and coastal waters are primarily limited by nitrogen or not In lakes (see the description in
Section 10 8 4 1), conditions of nitrogen deficiency often produce blooms of planktonic
blue-green algae, which fix atmospheric nitrogen and act to return the algal community to a
condition of nitrogen sufficiency (Schindler, 1977, Flett et al , 1980) Only when N P ratios
are extremely low and blue-green algae are unable to fix enough nitrogen to bring the ratio
10-195
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up to the Redfield proportions do lakes remain nitrogen limited (Howarth et al, 1988a)
Why, then, doesn't the same phenomenon (nitrogen fixation by blue-green algae) occur in
nitrogen-deficient marine systems9 A major difference in the biogeochemistry of lakes and
estuaries is that nitrogen fixation by free-living algae (phytoplankton) rarely occurs in
estuaries, even when the N'P ratios of incoming water suggest severe nitrogen limitation
Howarth et al (1988b), for example, surveyed a large number of estuaries along the Atlantic
coast of the United States and found no instances in which nitrogen-fixing blue-green algae
made up more than 1 % of the algal biomass A number of explanations for this lack of
nitrogen fixation in estuaries have been proposed, including shorter water residence tunes
(faster flushing rates) than lakes, greater turbulence than in lakes, and lower concentrations
of micronutnents (especially Fe and molybdenum) needed for the biochemical pathways in
nitrogen fixation (Howarth, 1988, Howarth et al, 1988b) Of these, only the last argument
really holds true in a comparison of lakes and estuaries Howarth and Cole (1985) and Cole
et al. (1986) have determined that the high concentrations of sulfate in marine systems
interfere with the assimilation of molybdenum by marine algae, and propose that low rates of
molybdenum availability are, in turn, limiting to rates of nitrogen fixation in many systems
Molybdenum limitation, however, has not been experimentally demonstrated in many marine
environments In fact, many nutrient addition bioassays conducted in benthic environments
have shown that the availability of organic matter and of oxygen-depleted microenvrronments
tightly control marine microbial nitrogen fixation potentials (Paerl et al, 1987, Paerl and
Prufert, 1987). Because the enzymes needed for nitrogen fixation are readily inactivated by
oxygen, rates of fixation may be limited by energy availability (i e , the supply of carbon
reductant) and ambient oxygenation By and large, nitrogen-deficient marine waters are
depleted in readily oxidrzable organic matter and are well oxygenated When high rates of
nitrogen fixation do occur in marine systems, they are usually associated with
bottom-dwelling (benthic) algae (Howarth, 1988), these habitats are relatively enriched with
organic matter and support localized oxygen-depleted microenvrronments (Paerl et al, 1987)
Iron is also required for nitrogen fixation, and may limit rates of nitrogen fixation in some
freshwater lakes (Wurtsbaugh and Home, 1983), concentrations of Fe in seawater are often
much lower than in fresh water, and although little direct evidence of limitation of nitrogen
fixation by low Fe concentrations exists, it is certainly a likely condition (Howarth et al,
10-196
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1988b) It is difficult at this point in the debate over marine nitrogen fixation to state
anything definitively beyond the fact that nitrogen fixation is not common in marine waters
(Carpenter and Capone, 1983, Howarth et al, 1988a) One possible conclusion from the
debate among researchers in this field (e g , Howarth et al, 1988b, Paerl et al, 1987) is that
planktonic nitrogen fixers may be limited by micronutnent availability, whereas benthic
nitrogen fixers are limited by availability of organic carbon and high ambient oxygen levels,
but both factors, as well as others, probably operate in both environments Light, for
example, appears to play a role in clear, tropical lagoons (Potts and Whitton, 1977, Wiebe
et al, 1975) because benthic nitrogen-fixing algae in these environments require light for
photosynthesis The presence of benthic nitrogen fixation in Smith's (1984) subtropical
lagoons may help explain the apparent contradiction between his predictions of phosphorus
limitation and experimental results suggesting nitrogen limitation in slowly flushed systems
Nixon and Pilson (1983) have summarized the results of numerous input/output studies
in estuaries and coastal waters and related the inputs of various nutrients to algal biomass
Then- results for nitrogen are repeated in Figure 10-35 and are supported by a similar
analysis conducted by Boynton et al (1982) for algal productivity The relationship between
nitrogen inputs and mean algal biomass in marine systems is certainly much weaker than the
relationship between phosphorus and biomass in lakes (e g , Schindler, 1978), but is
nonetheless suggestive of a general pattern of nitrogen limitation in these systems
(Figure 10-35) Seasonal effects on nutrient ratios, grazing by zooplankton, and physical
factors such as light, circulation patterns, and turbidity all lend uncertainty to the
relationship Perhaps the most important aspect of the relationship is the apparent strong
dependence of annual maximum chlorophyll concentrations (Figure 10-35b) on nitrogen
inputs (r = 0 57, p < 0 0001) Many of the most severe impacts of eutrophication are
experienced during summer algal blooms, these seem to be more strongly dependent on
nitrogen than biomass in other seasons (e g , D'Elia et al, 1986).
In summary, there does seem to be confirmatory evidence of nitrogen limitation in
many estuanne and coastal marine ecosystems This conclusion is a general rule, rather than
an absolute one, and other limiting factors certainly occur in some locations, and during
some seasons In general, ratios of nitrogen to phosphorus in inputs to estuaries and coastal
waters are much lower than in lakes (Hecky and Kdham, 1988, Howarth, 1988), and this
10-197
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Nitrogen Input (nmol/L/year)
Figure 10-35. Concentrations of (a) mean algal chlorophyll and (b) annual maximum
chlorophyll, in the midregion of various estuaries (1 to 15) and in the
Marine Ecosystem Research Laboratory experimental ecosystems
(A to G) as a function of the input of dissolved inorganic nitrogen.
1 - Providence River estuary, RI; 2 - Narragansett Bay, RI; 3 - Long
Island Sound; 4 - Lower New York Bay; 5 - Delaware Bay; 6 - Patuxent
River estuary, MD; 7 - Potomac River estuary, MD; 8 - Chesapeake Bay;
9 - PamUco River estuary, NC; 10 - Apalachicola Bay, FL; 11 - Mobile
Bay, AL; 12 - Barataria Bay, LA; 13 - North San Francisco Bay, CA;
15 - Kaneohe Bay, HI. Note change in scale on vertical axis.
Source Nixon and Pilson (1983)
10-198
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probably contributes strongly to the apparent difference between lakes and marine systems in
their nutrient limitation These low ratios, however, result largely from sewage inputs
(Ryther and Dunstan, 1971, Jaworski, 1981, Howarth, 1988), and whether atmospheric
deposition of nitrogen contributes to eutrophication in these systems will depend strongly on
the relative inputs of nitrogen from these two sources As stated in the introduction to this
section, any question of negative impacts on estuaries and coastal waters from nitrogen
deposition depends both on a determination of nitrogen limitation and on a determination that
atmospheric deposition is a major contributor of nitrogen to these ecosystems
Anthropogenic sources of nitrogen to estuaries and coastal waters include point sources
(such as sewage plant outfalls), fertilizer and animal wastes in runoff, and atmospheric
deposition (predominantly due to NOX from combustion and ammonium from agricultural
activity) Atmospheric deposition may be supplied directly to the surfaces of estuaries or
coastal waters or may be supplied indirectly to the watershed and subsequently transported to
the coast by river flow As discussed earlier, nitrogen can be deposited in a variety of
forms, two of the contentious issues in determining the impact of NOX on estuanne
ecosystems are estimating the total deposition and the uncertainty in the relative proportion
contributed by the different forms, especially between dry and wet deposition (e g , Fisher
et al , 1988a)
Runoff inputs to estuaries may be the most variable of the nitrogen inputs They vary
with watershed area, precipitation rates, land-use patterns (especially the use of fertilizer),
and rates of atmosphenc deposition Spring runoff represents a major input of nutrients to
estuanne and coastal systems Runoff inputs vary seasonally (e g , Jaworski, 1981) and
from year to year (e g , Boynton et al , 1982; Jaworski, 1981) Nitrate inputs to estuaries
increase markedly during flooding conditions (Biggs and Cronin, 1981), and are at least
partially responsible for the finding that nitrogen is less likely to be limiting in the winter and
spring than in the summer (above)
Point sources of nutnents may be particularly important near urbanized areas Sewage
inputs contribute more than half of the inorganic nitrogen content to a number of major
estuaries in the United States Long Island Sound (67%), New York Bay (82%), Rantan
Bay (86%), San Francisco Bay (73%), and Delaware Bay (50%) (Nixon and Pilson, 1983)
10-199
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Natural and anthropogenic sources of nitrogen to coastal waters may result in the same
form of nitrogen (e g , NO3") being transported by the same route (e g , river input) Their
effects will, therefore, be indistinguishable, and it becomes impossible to assign
"responsibility" for a problem to a particular source This has obvious consequences for
policy decisions because, for example, there are many possible regulatory actions that could
all result in the reduction of nitrate input to a particular estuary It may be more cost
effective, for example, to increase the efficiency of nitrogen removal in sewage treatment
than to reduce NOX emissions, even if NO3" inputs from atmospheric deposition are
increasing.
The first published attempt to determine the relative importances of nitrogen from
deposition, and nitrogen from runoff, was that of Correll and Ford (1982) for the Rhode
River estuary, a tributary to the Chesapeake Bay Correll and Ford assumed in their analysis
that all atmospheric nitrogen deposited on the watershed was retained, and that the only
atmospheric inputs of nitrogen to the estuary were those that fell directly on the water
surface This estimate should, therefore, be considered a lower limit to the importance of
atmospheric deposition because some terrestrial watersheds do show retention capacities
lower than 100% (see discussion of nitrogen saturation, above) Correll and Ford (1982)
conclude that, on an annual basis, atmospheric and watershed sources of nitrogen to the
Rhode River are approximately equal During the summer and fall, a period when the
Chesapeake Bay undergoes substantial anoxia, precipitation inputs of nitrogen may slightly
exceed those from watershed runoff It is important to note that the watershed of the Rhode
River estuary is small relative to the estuary itself (the watershed is less than six times the
size of the estuary) These results should be extrapolated with caution to situations where
watershed sizes may be orders of magnitude larger than those of the waters that dram them
The entire Chesapeake Bay, for example, is approximately one-fifteenth the size of its
watershed, and the relative importance of nitrogen falling directly on the water surface
would, therefore, be smaller relative to terrestrial inputs
Paerl (1985) has determined that NO3"-ennched rain falling on the waters of Bogue
Sound (an embayment), the Continental Slope, and the Gulf Stream (all off the east coast of
North Carolina) increased algal biomass as much as fourfold, and that rain falling directly on
the ocean surface accounted for as much as 10 to 20% of the volume of water supplied to
10-200
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these near-coastal areas More recent work (Paerl et al, 1990) indicates that rainfall
additions as low as 0 5 % by volume stimulated algal primary production and biomass in
these nitrogen-limited waters Paerl (1985) and Paerl et al (1990) did not estimate the
proportion of the total nitrogen inputs to these areas that entered as precipitation, but they do
suggest that algal blooms initiated by direct inputs of nitrogen from large ram storms could
be sustained by NO3'-enriched runoff from nearby land masses Terrestrial inputs of
nitrogen (from runoff) usually lag rainfall by 4 to 5 days in this region These studies appear
to be unique in showing a direct link between nitrogen deposition and algal productivity, but
do not provide enough information to estimate the overall importance of deposition to the
maintenance of high algal biomass in these waters
10.8.4,3 Evidence for Nitrogen Deposition Effects in Estuarine Systems—Case Studies
Complete nitrogen budgets, as well as information on nutrient limitation and seasonal
nutrient dynamics, have been compiled for two large estuaries, the Baltic Sea and
Chesapeake Bay, and for the Mediterranean Sea In the case of the Mediterranean,
Loye-Pilot et al (1990) suggest that 50% of the nitrogen load originates as deposition falling
directly on the water surface In the case of the Baltic and Chesapeake, deposition of
atmospheric nitrogen has been suggested as a major contributor to the eutrophication of the
estuaries (see below) Data for other coastal and estuanne systems are less complete, but
similarities between these two systems and other estuanne systems suggest that their results
may be more widely applicable The discussion in this document is limited to these two
"case studies," with some speculation about how other estuaries may be related
The Baltic Sea
The Baltic Sea is perhaps the best-documented available case study of the effects of
nitrogen additions in causing estuanne eutrophication Like many other coastal waters, the
Baltic Sea has expenenced a rapidly increasing anthropogenic nutrient load, it has been
estimated that the supply of nitrogen has increased by a factor of 4, and phosphorus has
increased by a factor of 8, since the beginning of the century (Larsson et al , 1985) The
first observable changes attributable to eutrophication of the Baltic were declines in the
concentration of dissolved oxygen in the 1960s (Rosenberg et al , 1990) Decreased
10-201
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dissolved oxygen concentrations result when decomposition in deeper waters is enhanced by
the increased supply of sedimenting algal cells from the surface water layers to the
sediments. In the case of the Baltic, the spring algal blooms that now result from nutrient
enrichment consist of large, rapidly sedimenting algal cells, which supply large amounts of
organic matter to the sediments for decomposition (Enoksson et al , 1990) Since the 1960s,
researchers in the Baltic have documented increases in algal productivity, increased incidence
of nuisance algal blooms, and periodic failures and unpredictability in fish and Norway
Lobster catches (Fleischer and Stibe, 1989, Rosenberg et al, 1990)
It has now been shown by a number of methods that algal productivity in nearly all
areas of the Baltic Sea is limited by nitrogen Nitrogen-to-phosphorus ratios range from
6:1 to 60.1 (Rosenberg et al, 1990), but the higher ratios are only found in the remote, and
relatively unimpacted, area of the Bothman Bay (between Sweden and Finland) Productivity
in the spring (the season of highest algal biomass) is fueled by nutrients supplied from deeper
waters during spring overturn (Granek et al, 1990), deep waters are low in nitrogen and
high hi phosphorus, resulting in N P ratios near 5 (Rosenberg et al , 1990), suggesting
potential nitrogen limitation when deep waters are mixed with surface waters Low N P
ratios in deep water result from denitnfication in the deep sediments (Shaffer and Ronner,
1984). Primary productivity measurements in the Kattegat (the portion of the Baltic between
Denmark and Sweden) correlate closely with uptake of NO3", but not of phosphate ions
(Rydberg et al, 1990) Level n and IDE nutrient enrichment experiments conducted in near-
shore areas of the Baltic, as well as in the Kattegat, indicate nitrogen limitation at most
seasons of the year (Graneli et al, 1990) Growth stimulation of algae has also been
produced by addition of ram water to experimental enclosures, in amounts as small as 10%
of the total volume (Graneli et al, 1990), rain water in the Baltic is enriched in nitrogen, but
is phosphorus-poor In portions of the Baltic where freshwater inputs keep the salinity low,
blooms of the nitrogen-fixing blue-green alga Aphamzomenon flos-aquae are common
(Graneli et al , 1990), blue-green algal blooms are common features of nitrogen-limited
freshwater lakes (see Section 10 6 4 1), but are usually absent from marine waters
Nitrogen budget estimates indicate that the Baltic Sea as a whole receives
1 X 109 kg/year of nitrogen, of which 3 9 x 108 kg/year (37%) comes directly from
atmospheric deposition (Rosenberg et al, 1990) Fleischer and Stibe (1989) report that the
10-202
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nitrogen flux from agricultural watersheds feeding the Baltic have been decreasing since
about 1980, but that the nitrogen contribution from forested watersheds is increasing, they
cite both increases in nitrogen deposition and the spread of modern forestry practices as
causes for the increase It should be noted, however, that the Baltic also expenences a
substantial phosphorus load from agricultural and urban lands, and that phosphorus inputs
may help to maintain nitrogen-limited conditions (Graneli et al, 1990) If the Baltic had
received consistent nitrogen additions (e g , from the atmosphere or from agricultural runoff)
in the absence of phosphorus additions, it might well have evolved into a phosphorus-limited
system some tune ago
The physical structure of the Baltic Sea, with a shallow sill limiting exchange of water
with the North Sea (see the definition of a fjord, above) contributes to the eutrophication of
the basin by trapping nutrients in the basin once they reach the deeper waters Because the
larger algal cells that result from nutrient enrichment m the basin provide more nutrients to
the deep water through sedimentation, and because only shallow waters have the ability to
exchange with the North Sea, it is estimated that less than 10% of nutrients added to the
Baltic are exported over the sill to the North Sea (Wulff et al, 1990) Throughout much of
the year, especially during the dry months, productivity in the Baltic is maintained by
nutrients recycled within the water column (Enoksson et al., 1990) The trapping of
nutrients within the basin and recycling of nutrients from deeper waters by circulation
patterns suggest that eutrophication of the Baltic is a self-accelerating process (Enoksson
et al , 1990), with a long time lag between reductions of inputs and improvements in water
quality
Chesapeake Bay
The most complete attempts to estimate the relative importance of atmospheric
deposition to the overall nitrogen budget of an estuary or coastal ecosystem in the United
States were completed for Chesapeake Bay by the Environmental Defense Fund (EDF)
(Fisher et al , 1988a, Fisher and Oppenheimer, 1991) and by Versar, Inc (Tyler, 1988) in
1988 Neither of these reports has been published in a peer-reviewed arena, but the issue of
atmospheric contributions to the eutrophication of the Chesapeake has been widely discussed
10-203
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(and criticized), particularly after the publication of the EDF report, and bears close
examination for these reasons
Both reports conclude that atmospheric deposition makes a substantial contribution
(25 to 40% of total inputs) to the nitrogen budget of Chesapeake Bay In both cases,
nitrogen budgets for the bay were constructed via a number of steps, each of which involved
simplifying assumptions that bear further examination Both reports calculate inputs from
atmospheric deposition to the bay itself (Step #1), atmosphenc deposition to the watershed
(#2), fertilizer application in the watershed (#3), generation of animal wastes in the
watershed (#4), inputs from urban land use (#5), and point source inputs (#6) Once the
total inputs to the watershed and bay were estimated, both reports calculated the proportion
of the inputs that were retained by the watershed (Step #7) and the proportion that were
retained within the rivers and tributaries feeding the bay (#8)
The two reports had different goals, which make their results difficult to compare The
EDF report (Fisher et al, 1988a, Fisher and Oppenheimer, 1991) estimated the proportions
of both NO3" and NH4+ deposition to the total nitrogen budget of the Chesapeake (including
all forms of nitrogen, and both base flow and storm flows) The Versar report (Tyler,
1988), on the other hand, estimated only contributions of NO3", because NH4+ does not
result from the burning of fossil fuels, and excluded base-flow contributions In addition, the
Versar report used a range of values both for the watershed contributions made by each
nitrogen source (deposition, fertilizers, etc ) and for the fraction of the inputs retained by the
watershed (transfer coefficients) This results in a wide range of budget values for each of
the sources, and for the relative importance of NO3" deposition to the budget, which
complicates any comparison of the results of the two studies Nonetheless, the two reports
used similar methods in developing their budgets, and a combined discussion of the
uncertainties involved in each of the steps listed above is warranted
The results for the two budgets are presented in Table 10-27 Since the publication of
these budgets, additional information on such issues as dry deposition and retention of
nitrogen by forested watersheds has become available This new information has been
compiled to produce a third "refined" budget, which is also presented in Table 10-27 The
assumptions that were used to construct the refined budget are outlined in each of the
discussions of individual budgeting steps below
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TABLE 10-27. THREE NITROGEN BUDGETS FOR CHESAPEAKE BAY
Source of Nitrogen
EDF Budget Versar Budget Refined Budget
(kg X 108/year) (kg X 108/year) (kg X 108/year)
Direct Deposition
Nitrate Ions 0 8
Ammonium Ions 0 4
Nitrogen Load to Bay (from direct deposition) 1 3
Forests
Nitrate Ion Deposition 9 0
Ammonium Ion Deposition 4 9
Watershed Retention 0 8
In-Stream Retention 1 4
Atmospheric Nitrate Ion Load to Bay (from forests)
Nitrogen Load to Bay (from forests)
Pasture Land
Nitrate Ion Deposition 2 4
Ammonium Ion Deposition 1 3
Animal Wastes 14 5
Watershed Retention 0 7
In-Stream Retention 1 5
Atmospheric Nitrate Ion Load to Bay (from
pastures)
Nitrogen Load to Bay (from pastures)
Cropland
Nitrate Ion Deposition 2 5
Ammonium Ion Deposition 1 4
Fertilizers 15 8
Watershed Retention 0 8
In-Stream Retention 5 9
Atmospheric Nitrate Ion Load to Bay (from
cropland)
Nitrogen Load to Bay (from cropland)
Residential/Urban
Nitrate Ion Deposition 0 4
Ammonium Ion Deposition 0 3
Watershed Retention 0 3
In-Stream Retention 0 4
Atmospheric Nitrate Ion Load to Bay (from urban
areas)
Nitrogen Load to Bay (from urban areas)
Point Sources 3 4
NITRATE ION LOAD TO BAY (FROM 3 5
DEPOSITION)
TOTAL NITROGEN LOAD TO BAYb 13 94
Percent of Nitrogen from Nitrate Ion Deposition 25%
50%
07
a
07
84
a
02
02
95%
50%
95%
1 7
-a 94-'
11 8 50%
0 01-0 06
007-04
}70%
28
a
4 1-27 0
001-03
0 06-3 6
76-99%
50%
20-3 2
0 94-1 48
3 03-8 26
18-3 l%e
06
03
08
64
3 5
07
10
13
07
195
0 13
08
21
1 1
158
007
06
07 06
35% -a 62-96% 03
0% 0 01-0 14 20% 0 1
0 01-0 14 03
34
153
682
225%
846%
35%
95%°
35%
35%
50%
35%
aThe Versar Budget (Tyler, 1988) does not calculate loads of ammonium ions
bFor the Environmental Defense Fund (EDF) Budget (Fisher et al , 1988a, Fisher and Oppenheimer, 1991) and
refined budget, total nitrogen load to the bay includes both nitrate ion (NO3 ) and Nlfy The Versar Budget
(Tyler, 1988) includes only NO3"
°Watershed and in-stream retention values for pastureland in the EDF Budget apply only to animal wastes For
atmospheric deposition, the cropland retention value (70%) was ui>ed
95% retention was used for animal wastes, 85% retention was used for deposition (see text)
eThe range of contributions of NO3~ deposition to the total budget were calculated by comparing
maximum-to-maximum estimates and minimum-to-minimum estimates These combinations are more likely to
occur during extreme (e g , very wet or very dry) years
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Hie major uncertainty involved in calculating direct inputs to the Chesapeake from
atmospheric deposition (Step #1, above) is estimation of the contribution of dry deposition
(see also Section 10.2) Both reports use actual deposition monitoring data (i e , from
NADP/National Trends Network) to estimate the nitrogen load from wet deposition and then
assume that the rate of dry deposition of nitrogen in the watershed is equal to the rate of wet
deposition. As discussed earlier (see Section 10 8 2 on nitrogen inputs), the measurement of
dry deposition is a much vexed issue, and most researchers make educated guesses of rates
of dry deposition by assuming that they are some fraction of wet deposition rates The
assumption that dry deposition is equal to wet deposition is probably reasonable for areas
directly adjacent to emissions sources (Summers et al, 1986), but the ratio of dry deposition
to the sum of wet and dry deposition may fall as low as 0 2 in locations remote from
sources. For example, Barne and Sirois (1986) estimated that dry deposition contributed
21 to 30% of total NO3~ deposition in eastern Canada Baker (1991) concludes that dry
deposition of NO3" is approximately 40% of wet deposition, whereas dry deposition of NH4+
is approximately 34% of wet deposition (resulting in ratios of dry deposition to wet plus dry
deposition of 0 29 and 0 25, respectively) for areas remote from emissions In the most
complete analysis of dry and wet deposition of NO3" to date, Sisterson et al (1990) reported
ratios of dry deposition to wet plus dry deposition of 0 35 for two locations inside or near
the borders of the Chesapeake Bay watershed (State College, PA, and West Point, NY)
Based on the results of these studies, it seems that the assumption made in the two
Chesapeake Bay nitrogen budgets (i e , that dry deposition is equal to wet deposition)
probably overestimates the importance of dry deposition The 0 35 ratio is used in
constructing the refined budget in Table 10-27
The two reports (Fisher et al, 1988a, Fisher and Oppenheimer, 1991, Tyler, 1988)
also present different values for the direct contribution of wet deposition to the bay because
they use different methods to estimate the spatial pattern of deposition in the bay and its
watershed The EDF report uses wet deposition values from the nearest NADP collector,
and the Versar report extrapolates deposition values from isopleth maps of NO3" deposition
In addition, the Versar report includes direct atmospheric inputs to the tributaries of the bay,
as well as to the bay itself (Table 10-27) Aside from problems with estimating dry
deposition, it seems likely that the approach used in the Versar report for estimating
10-206
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deposition is more precise than that used in the EDF report The Versar values for wet
deposition were, therefore, used in the refined budget, after adjusting them to reflect a 35 %
contribution from dry deposition Ammonium deposition was calculated for the refined
budget by applying the ratio of NH4+ to NO3" deposition reported in the EDF report to the
estimated NO3" deposition values from the Versar report (i e , these values assume that the
spatial pattern in NH4+ deposition is the same as the spatial pattern for NO3" deposition)
The uncertainties involved in estimating nitrogen deposition to the Chesapeake Bay
watershed (Step #2) are similar to those for estimating direct deposition It seems likely that,
by assuming dry deposition is equal to wet deposition, both reports overestimate the dryfall
contnbution to deposition Differences between the estimates of wet deposition presented in
the two reports result from the same methodological differences used in estimating direct
inputs (i e , use of the nearest NADP collector versus extrapolated values from isopleth
maps) and from slight differences in the estimates of the coverage of each land-use type
The Versar method produces slightly lower estimates ol atmospheric nitrogen inputs to the
basin (Table 10-27) and, as in the case with estimates of direct deposition to the bay, the
Versar method probably produces better estimates of basin- wide deposition loads than the
EDF approach The refined budget uses the Versar values for wet NO3" deposition (adjusted
to reflect a 0 35 ratio for dry deposition, as above) and estimates of NH4+ deposition based
on the Versar spatial deposition pattern and the EDF estimate of NH^ deposition, as above
The EDF report (Fisher et al , 1988a, Fisher and Oppenheimer, 1991) uses county
agricultural reports and U S Census Bureau data to calculate the application rates of
fertilizers to the counties (and portions thereof) in the Chesapeake Bay watershed (Step #3,
above) The Versar report (Tyler, 1988) calculates the total fertilizer load (from NO3") to
the watershed by applying a correction factor to the level of fertilizer application
recommended by the U S Department of Agriculture, the conection factor was based on
local officials' best guesses of actual fertilizer application rates (e g , 30 to 60% of the
recommended rates) Because it deals only with NO3" loading, the Versar approach also
necessitates making an assumption about the proportion of nitrogen fertilizers that are applied
as NO3", as opposed to NH4+ or urea, the report assumes that 60% of the nitrogen added is
in the form of NO3", but presents no data to support this assumption Because it is more
direct in nature, the EDF approach to estimating fertilizer inputs seems to be more defensible
10-207
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than the Versar approach, and the EDF estimate is, therefore, used in the refined budget
The EDF estimate of 15 8 x 10 kg/year is near the bottom range of fertilizer loads
estimated by the Versar report (Table 10-27)
The EDF (Fisher et al, 1988a, Fisher and Oppenheimer, 1991) and Versar (Tyler,
1988) reports use the same estimate (from the EDF report) for the contribution by animal
wastes (Step #4, above) to the nitrogen budget The EDF report used county agricultural
statistics to calculate the total number of farm animals of different types in the Chesapeake
Bay watershed These population numbers were then multiplied by published estimates of
the amount of nitrogenous wastes excreted by each type of animal annually, to produce an
estimate of 19 5 X 10 kg/year As in the estimates of fertilizer NO3" inputs, the Versar
report assumed that 60% of animal nitrogenous wastes were in the form of NO^", this
estimate seem especially difficult to justify when it is used both for animal wastes and for
fertilizers, as there is no reason to expect both nitrogen sources to have the same
composition. The EDF estimate of 19 5 x 10 kg/year is used for the refined budget
In both reports, atmospheric deposition is considered to be the only source of nitrogen
to urban areas (Step #5, above) As pointed out m the Versar report (Tyler, 1988), this is
likely to be an underestimate because it ignores fertilizer applications to lawns and gardens
Because fertilizers applications are seasonal, and the area of urban land in the basin is small
(about 3% of the total), this underestimate is considered unimportant As mentioned earlier,
the EDF (Fisher et al, 1988a, Fisher and Oppenheimer, 1991) and Versar reports use
slightly different methods to calculate wet deposition The primary difference between the
two estimates of nitrogen loadmg to urban areas (Table 10-27), however, is in their estimate
of the proportion of the basin in residential and urban land use (5 X 105 ha in the EDF
report versus 8 x 10 ha m the Versar report) In neither case does the nitrogen
contribution from urban lands (<2% of the total loadmg to the watershed) play a significant
role in the budgets. The Versar estimate of deposition to urban areas is used in the refined
budget, with the same adjustments applied as for the deposition to the watershed and directly
to the bay (above)
Both reports used the same EPA estimates of point source inputs to the Chesapeake Bay
watershed (Step #6, above), the lower value presented in the Versar report (Tyler, 1988) is
the estimated proportion of point source inputs that are in the form of NO3", again assuming
10-208
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that NO3" is 60% of the total inorganic nitrogen The upper limit to the range of point
source inputs presented by the Versar report is a more recent (1988) estimate from the
Chesapeake Bay Program There seems to be little reason not to use the original EDF value
(Fisher et al, 1988a, Fisher and Oppenheimer, 1991) of 32 9 x 106 kg/year (Table 10-27),
and this value is used in the refined budget
Perhaps the greatest source of uncertainty in both nitrogen budgets is created when the
proportions of nitrogen inputs that are retained within the watershed are estimated (Step #7,
above) Both reports use a variety of methods to calculate separate transfer coefficients for
each land use type, and in some cases, for different sources of nitrogen within a single land-
use type In particular, the Versar report (Tyler, 1988) compares calculated loads
(as described in the preceding paragraphs) to calculated runoff from each land-use type (from
Smullen et al, 1982) and estimates a range of transfer coefficients from these calculated
values Because the error inherent in the calculated values is amplified when they are
compared, this method seems especially problematic Often, the calculated transfer
coefficients differ greatly from coefficients measured for smgle basins within the Chesapeake
Bay watershed The transfer coefficients for each land-use type are discussed in detail
below It should be emphasized that all of the nitrogen budgets discussed below deal only
with inorganic forms of nitrogen (i e , NO3" and NH4 ) Outputs of organic nitrogen from
watershed can be substantial (e g , Correll and Ford, 1982), and organic forms can result
from atmospheric deposition sources when watershed processes route nitrogen through the
biotic portion of the ecosystem Given this possible source of error, the nitrogen retention
values presented below should probably be considered maximum estimates
Estimating watershed retention of nitrogen in forested watersheds is difficult, primarily
because so few data are available, and the applicability of smgle watershed values to wide
areas of the Chesapeake Bay watershed is untested The Versar (Tyler, 1988) report
compares calculated deposition loads (Table 10-27) to estimates of runoff from forests (from
Smullen et al, 1982) to yield a transfer coefficient of 4 8% As discussed above, this
estimate must be considered very uncertain, because of the combined errors introduced by
comparing two calculated values The EDF report (Fisher et al., 1988a, Fisher and
Oppenheimer, 1991) found literature values that ranged from 50% (in the Mid-Appalachians)
to 97% (in the Coastal Plain), and used 80% as a "reasonable mid-range estimate" Given
10-209
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the range of possible retention values, it seems unlikely that any single number would be a
reasonable estimate for the entire Chesapeake Bay watershed Some additional nitrogen
retention values are given in Table 10-28, based on published nitrogen budgets for
watersheds in or near the Chesapeake Bay basin These are arranged according to
physiographic regions, in order to illustrate the spatial variability in watershed nitrogen
retention Of the values in Table 10-28, only those of Kaufmann et al (1991) are applicable
to broad spatial areas, because they are based on a probability sampling of streams in each
region. These values assume that NO3" concentrations at spring base flow are representative
of annual mean concentrations (Kaufmann et al, 1988, Messer et al , 1988) If the retention
coefficients for each physiographic region are weighted by the proportion of the Chesapeake
Bay watershed in each physiographic region (from SmuUen et al, 1982), an area-weighted
retention coefficient of 84 6% results, this figure was used for the refined budget
(Table 10-27). The 84 6% figure agrees remarkably well with the data presented in
Figure 10-28b (Driscoll et al, 1989a), which suggest an interpolated coefficient of 84 7% at
the levels of deposition calculated for the Chesapeake Bay watershed (8 9 kg/ha total
deposition, or 5.8 kg/ha of wet deposition)
Nitrogen retention by pasturelands is generally thought to be very high Both the EDF
(Fisher et al., 1988a; Fisher and Oppenheimer, 1991) and the Versar (Tyler, 1988) reports
estimate retention coefficients in the 94 to 99% range As with forest nitrogen retention, the
EDF estimate is based on published values from watershed studies, whereas the Versar
estimate is based on comparisons of calculated loads and calculated runoff The EDF
estimate (95%) is based primarily on a study by Kuenzler and Craig (1986, as reported in
Fisher et al, 1988a, Fisher and Oppenheimer, 1991) on pastureland in the Chowan River,
NC, watershed Similar results (94 4% retention) have been reported for unfertilized pasture
lands in Ohio by Owens et al (1989), where NO3" losses were lower from pastureland than
from nearby undisturbed forests (86% retention) Nitrogen retention coefficients reported
here were recalculated to include dry deposition (at 35% of total deposition), as was the case
for forest nitrogen budgets reported above The EDF report applies the 95 % retention rate
only to animal wastes, and uses a 70% retention coefficient for atmospheric deposition
Because they are primarily in the form of particulate organic matter, it seems reasonable to
assume that animal wastes will be more strongly retained than deposition The refined
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TABLE 10-28. RETENTION OF NITROGEN IN WATERSHEDS IN OR
NEAR THE CHESAPEAKE BAY BASIN, FROM PUBLISHED REPORTS.
ALL NITROGEN LOADS HAVE BEEN REESTIMATED BASED ON MEASURED
WET DEPOSITION, AND A 35% CONTRIBUTION TO TOTAL DEPOSITION
FROM DRY DEPOSITION
Physiographic Region
Poconos/Catskills
Biscuit Brook, NY
Northern Appalachians
Southwestern Pennsylvania
Southwestern Pennsylvania
Fernow, WV
Eastern Tennessee
Valley and Ridge
Catoctin Mountains, MD
Shenandoah National Park, VA
Mid-Atlantic Coastal Plain
Chesapeake Bay, MD
Piedmont
Northern Georgia
Southern Blue Ridge
Nitrate Ion
Nitrogen Load Export
(106 eq/year)a (106 eq/year)
-
878 214
1,192 264
1,506 607
707 36
-
593 250
557 3
1,000 10
486 11
Percent
Retention
883
757
727
780
945
595
946
785
575
995
909
990
902
977
883
Source
Kaufmann et al
(1991)
Stoddard and
Murdoch (1991)
Kaufmann et al
(1991)
Barker and Witt
(1990)
Sharpe et al
(1984)
Helvey and
Kunkle (1986)
Kelly (1988)
Kaufmann et al
(1991)
Katzetal (1985)
Shaffer and
Galloway (1983)
Kaufmann et al
(1991)
Weller et al
(1986)
Kaufmann et al
(1991)
Buell and Peters
(1988)
Kaufmann et al
(1991)
aNitrogen loads are calculated from published wet deposition estimates, extrapolated to total deposition
according to a 0 35 dry wet plus dry ratio (see text)
Retention estimates are calculated by comparing mean concentrations of piecipitation to mean concentrations in
stream water Estimates from Kaufmann et al (1991) are from the National Stream Survey (Kaufmann
et al , 1988) and are for the population of streams within each physiographic province
10-211
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budget, therefore, applies the 95% retention figure for animal wastes, and an 85% retention
coefficient (as for forests, above) for nitrogen from deposition (Table 10-27)
The ability of croplands to retain, nitrogen is generally high because most of the
nitrogen applied to crops as fertilizer is removed as biomass during harvest (Lowrance et al,
1985; Groffman et al, 1986) Both the EOF (Fisher et al, 1988a, Fisher and Oppenheimer,
1991) and the Versar (Tyler, 1988) budgets compare estimates of fertilizer and deposition
loads to estimates of runoff from croplands to calculate nitrogen transfer coefficients Use of
loads estimates from a number of sources creates a range of retention coefficients from 70%
(Fisher et al, 1988a; Fisher and Oppenheimer, 1991) to 99% (Tyler, 1988) Published
values from studies of cropland watersheds are all toward the higher end of this range
Peterjohn and Correll (1984) measured a retention coefficient of 93 2 % for a fertilized corn
field in Maryland. Groffman et al (1986) reported 100% retention of fertilizer nitrogen in a
sorghum field in the Georgia piedmont, lower retention coefficients (76 1 %) were measured
during the winter, but the planting of crimson clover (a nitrogen-fixing legume) as a winter
cover crop complicates the interpretation of these figures Lowrance et al (1985) reported
nitrogen budgets for four cropland watersheds with a variety of crops in the Georgia Coastal
Plain, with retention coefficients ranging from 97 8 to 100 % Nitrogen retention coefficients
reported here were recalculated to include dry deposition (at 35% of total deposition), as was
the case for forest and pastureland nitrogen budgets reported above A retention coefficient
of 95%, as used for the refined budget (Table 10-27) is near the middle of the range of
published values Fertilizer inputs are generally in the same inorganic forms as atmospheric
deposition, and there seems no reason to apply different retention values to fertilizer and
deposition sources of nitrogen
Published reports of nitrogen retention in urban lands are apparently unavailable The
EDF report (Fisher et al, 1988a, Fisher and Oppenheimer, 1991) simply chose a retention
coefficient midway between their cropland value (70%) and complete runoff from impervious
surfaces (100%). The Versar report (Tyler, 1988) calculates transfer coefficients from
estimated loads (from deposition) and estimated runoff, and gives a range of 62 to 96%
(Table 10-27) There is little justification for choosing any particular value The 50 % value
used for the refined budget (Table 10-27) is chosen only to provide a "ball-park" value,
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slightly higher or lower values, when applied to the relatively small atmospheric loads falling
on urban areas, will not substantially change the conclusions presented here
The final assumption that affects the nitrogen budgets concerns the proportion of
watershed runoff that is lost during transport through rivers to the bay (Step 8, above)
Demtnfication in slow-moving lotic waters can significantly reduce the load of nitrogen
delivered to estuanne waters (see Section 10 8 2 4) In the absence of any measured loss
rates, both the EDF (Fisher et al, 1988a, Fisher and Oppenheimer, 1991) and the Versar
(Tyler, 1988) reports adopt the 50% loss value suggested by the Chesapeake Bay Program
(Smullen et al, 1982) More recently, denitnfication values have been published for two
rivers, the Potomac, which supplies water directly to Chesapeake Bay, and the Delaware,
which is adjacent to the Chesapeake Bay watershed (summarized in Seitzinger, 1988a)
Seitzinger and Garber (1987) estimated that 35 % of the dissolved inorganic nitrogen (NO3~
+ NH4+) load to the Potomac River was lost through denitnfication Seitzinger (1988b)
measured denitnfication rates at six locations in the tidal portion of the Delaware River and
estimated that 20% of the dissolved inorganic nitrogen load was lost through denitnfication
Both of these studies were conducted in the relatively flat, slow-moving and tidal portions of
nvers, where denitnfication rates are likely to be maximal, due to the existence of anoxic
sediments Data from smaller streams suggest that lower rates of nitrogen retention (10 to
15%) are more likely to occur in headwater streams (Tnska et al, 1990, Duff and Tnska,
1990) In light of these lower measured rates of nitrogen loss, the 50% figure used in the
EDF and Versar budgets seems insupportable for nvenne losses, loss rates as high as 50%
have been measured only in estuanne waters (e g , Narragansett Bay, Seitzinger et al, 1984,
the Baltic Sea, Larsson et al, 1985) The refined budget uses a figure of 35%, reflecting the
only known value for a nver feeding the Chesapeake itself (Seitzinger and Garber, 1987),
and may still overestimate in-stream retention in small streams
When the three budgets are compared, they suggest a wide range in estimated
contnbutions from individual sources of nitrogen (e g , estimates of cropland inputs vary
from 0 03 x 106 kg/year for the "best case" Versar budget to 59 8 x 106 kg/year for the
EDF budget), but a surprisingly consistent percentage contnbution from atmosphenc NO3"
deposition (18 to 31 %) to the total budget (Table 10-27) All three budgets suggest that a
large amount of nitrogen enters the bay from deposition, the 15 9 X 10 kg/year estimate
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from the refined budget corresponds to a nitrogen load of 44 metric tons per day entering
Chesapeake Bay from deposition directly to the bay and the watershed The caveat presented
earlier concerning organic forms of nitrogen should probably be repeated here, the estimates
of atmospheric NO3" contributions to the bay ignore all but the inorganic nitrogen fractions
Organic nitrogen can be a substantial contributor to the nitrogen in runoff, and could
potentially have a large atmospheric deposition component Many of the estimates that went
into these budgets are relatively certain For example, we have good data on wet deposition,
and can extrapolate to total deposition with reasonable certainty given recent estimates of dry
deposition within the watershed (e g , Sisterson et al, 1990) The biggest uncertainty in
estimating atmospheric NO3" loading to the bay results from the figure for retention of
nitrogen by forested watersheds This influence results from the fact that most of the
watershed (ca. 80%) is forested, small changes in the retention coefficients can have a large
effect on the estimated load to the bay from these watersheds The retention coefficient
calculated for the refined budget (84 6%) is our current best estimate, based on regional
estimates of retention within each of the physiographic regions in the Chesapeake Bay basin,
however, it still contains considerable uncertainty The retention coefficients listed in
Table 10-28 suggest that retention can vary from less than 60% to more than 99% in
individual watersheds Many more values from individual watersheds are needed before we
can be certain how representative the values for each physiographic region are
Taken as a whole, the budgets suggest that deposition is approximately equal in
importance to point-source supplies of nitrogen, and is possibly more important than
agricultural sources of nitrogen (Table 10-27) The fact that three different approaches
(i.e., the three budgets in Table 10-27) yield similar results lends weight to the suggestion
that atmospheric nitrogen contributes substantially to the eutrophication of the Chesapeake
Bay. The detrimental effects of eutrophication have been discussed earlier (see
Section 10 8 4.1) These results are surprising, given the emphasis usually placed on
reducing point-source inputs to the bay in order to improve water quality (e g , Chinchilli,
1989, Caton, 1989). Based on the results of nutrient limitation work discussed earlier, it
seems clear that the control of nitrogen inputs is important to the control of eutrophication in
the Chesapeake Bay The results of the budget exercises discussed here suggest that any
10-214
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program for nitrogen control should include the control of nitrogen deposition, as well as
point and nonpoint sources
Some corroboration of the budgets presented here is provided by recent attempts at
calculating nitrogen mass balances for the Upper Potomac River Basin (Groffman and
Jaworski, 1991, Jaworski and Linker, 1991) These studies apply both the EDF budget
technique and an "input-output analysis matrix" to calculate nitrogen loads and nitrogen
exports attributable to various sources within the Upper Potomac watershed (approximately
18% of the entire Chesapeake Bay watershed) The latter technique combines model
estimates of edge-of-field exports of nitrogen for different land-use types with a watershed
mass balance, where measurements or estimates of loads (e g , point sources, fertilization,
etc ) are balanced against measured or estimated outputs (e g , crop harvest, river export)
and the difference is attributed either to storage of nitrogen within the watershed or to
denitnfication and volatilization (gaseous losses) When applied to the Upper Potomac River
Basin, the EDF technique estimates that 10 6 x 106 kg/year of nitrogen that leaves the
watershed originated as atmospheric deposition (45 % of the total export) The second
technique estimates that 8 2 X 106 kg/year of the nitrogen leaving the watershed originated
as deposition (or 25% of the total export) The major difference between the two estimates
is rn the total export values (23 8 x 106 kg/year and 32 1 x 106 kg/year, respectively) The
value for the input-output analysis matrix is likely to be the best estimate for the Upper
Potomac because it is based on actual mass balance estimates of river export The same
discrepancy would apparently not exist if the input-output analysis matrix technique were
applied to the entire Chesapeake watershed, as the estimates of load to the bay from the EDF
technique match current best estimates of actual loads very closely (140 X 10 kg/year for
the EDF method, 130 x 10 kg/year for best current estimates, Fisher and Oppenheimer,
1991) Given the similarities in the two estimates of Upper Potomac River export
attributable to atmospheric deposition, and the unlikelihood that estimates for total river
export for the entire Chesapeake would differ as much as the estimates for the Potomac do,
the Upper Potomac River basin study lends substantial credence to the EDF technique The
improvements made to the EDF method in this document and presented in the "refined
budget" (Table 10-28) seem, therefore, to represent the best available information on
atmosphenc nitrogen loading to the Chesapeake Bay
10-215
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Finally, atmospheric NO3" inputs to the Chesapeake Bay should be put into the context
of seasonal nitrogen limitation of algal productivity in the bay As was discussed earlier, the
bay may undergo seasonal shifts in nutrient limitation, from phosphorus limitation in late
winter and early spring to nitrogen limitation during summer and fall (e g , D'Elia et al,
1982, D'Elia et al, 1986) If atmospheric NO3~ is to have a significant effect on algal
biomass, it would need to be present during the late summer, low-flow, high-biomass period
However, much of the NO3~ load occurs during the spring, when river flows and NO3"
leakage from watersheds are high (e g , Lowrance and Leonard, 1988) In the case of the
Baltic Sea, discussed earlier, nutrients were largely trapped within the estuary by
sedimentation processes and minimal water exchange with the North Sea Does the
Chesapeake Bay act in a similar manner to trap nutrients, providing a mechanism for
springtime loads of NO3" to influence summertime productivity9 Unfortunately, few
measurements of the nutrient retention capacities of the Chesapeake Bay are available, but
some estimates have been made Smullen et al (1982) estimated, based on some
measurements of current and nutrient concentrations at the mouth of the bay and a simple
box model, that virtually all of the nitrogen entering the bay was retained Nixon (1987) and
Nixon et al. (1986) question this conclusion, and point out the such high nutrient retention
rates should result in very high nutrient concentrations in the sediments, which have not been
found. Based on estimates of sediment nutrient concentrations, Nixon et al (1986)
calculated that only approximately 5 % of nitrogen entering the bay is retained The
argument of Nixon et al (1986), however, seems to ignore the potential effect of
denitrification in maintaining low sediment nitrogen concentrations, despite high rates of
retention by the bay Fisher et al (1988b) use longitudinal profiles of nutrient concentrations
throughout the bay to estimate that 33 to 71 % of nitrogen entering the bay is retained These
lower estimates of nitrogen retention suggest that nitrogen entering the Bay during spring
runoff does have the potential to affect productivity in the Bay during the critical summer
months. They also suggest, however, that the Chesapeake Bay could return to background
nitrogen concentrations within several flushing tunes of the bay, or within several years
(Fisher et al, 1988b), if nutrient control strategies were put in place
It is impossible to determine at this point whether the Chesapeake Bay example is an
unusual one in terms of the relative importance of atmospheric nitrogen inputs Jaworski
10-216
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(1981) gives crude nitrogen budgets for four estuaries and embayments in the United States,
his results suggest that the Chesapeake Bay receives an unusually large proportion of nitrogen
(68%) from land runoff (which includes agricultural and deposition sources) Jaworski's
(1981) budgets indicate that wastewater discharges are more important in the Hudson River
(New York) and San Joaqmn River (California) estuaries (63 and 47% of inputs,
respectively, but these estimates do not include deposition), and the Potomac River estuary
has equal inputs from wastewater and land runoff Of Jaworski's four systems, the
Chesapeake Bay is the least influenced by point-source pollution, but it also receives larger
inputs from point sources than many estuaries in the United States (e g , the Apalachicola
Bay, Nixon and Pilson, 1983) If one views all estuarme and coastal waters as lying along a
gradient from high to low influence by point-source pollution, then the relative importance of
deposition to the nitrogen budget will change as one moves along the gradient. The general
applicability of the nitrogen budget results from the Chesapeake Bay will depend on where
the bay falls along this gradient
10.8.5 Direct Toxicity Due to Nitrogen Deposition
In addition to the effects of acidification and eutrophicatton, nitrogen deposition could
potentially contribute to directly toxic effects in surface waters Toxic effects on freshwater
biota result from un-iomzed NH3 that occurs in equilibrium with ionized NH4+ and
hydroxide (OH") High NH3 concentrations are associated with lesions m gill tissue, reduced
growth rates of trout fry, reduced fecundity (number of eggs), increased egg mortality, and
increased susceptibility of fish to other diseases, as well as a variety of pathological effects m
invertebrates and aquatic plants (reviewed m U S Environmental Protection Agency, 1985)
Most analytical methods for ammonium actually measure the sum of NH3 and NH4+, which
is commonly referred to as NH4+; for clarity, the sum of ammonium and NH3 will be
referred to here as total ammonia (T-NH3) No single toxic concentration for T-NH3 can be
established because the relative contribution of NH3 to T-NH3, and the toxicity of NH3, vary
with the pH and temperature (Emerson et al, 1975) and the ionic strength (Messer et al,
1984) of the water The proportion of NH3 mcreases at higher temperatures and increasing
pH Because of the variability m NH3 toxicity, new criteria have recently been developed
that calculate the toxicity as a function of pH, temperature, and ionic strength (U S
10-217
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Environmental Protection Agency, 1985) The new regulations require the calculation of a
"final chronic value" (FCV) and "final acute value" (FAV), 4-day average concentrations of
NH3 cannot exceed the FCV more often on average than once every 3 years, nor can 1-h
average concentrations exceed one-half of the FAV more often on average than once every
3 years.
Critical concentrations of NH3 that cause the various effects are wide ranging and are
related to site specific temperature and pH values For example, the concentration values at
which 50% of the test organisms die within 48 h (48-h LC50) for Daphma magna, a common
invertebrate found in lake zooplankton, range from 38 to 350 ^mol/L T-NH3 over a
temperature range from 19 6 to 25 °C and pH range of 7 4 to 8 6 (U S Environmental
Protection Agency, 1985) However, results of toxicity tests on stream insects showed that
96-h LC5Q values ranged from 128 to 421 /tmol/L T-NH3 at relatively constant chemical
conditions. The 96-h LC50 values for rainbow trout ranged from 11 4 to 78 5 jwmol/L
T-NH3 Fingerhngs tend to be less sensitive than older life stages, and lower oxygen
concentrations increased sensitivity to NH3 Variation in temperature, pH, acclimation time,
and CO2 concentrations also appeared to explain some variation in responses Effler et al
(1990) calculated FCV and FAV values for Onondaga Lake, an urban lake in Syracuse, NY,
that is heavily polluted with municipal sewage For both salmonid and nonsalmonid fishes,
the FCV values varied (with tune of year) from 1 4 to 2 9 /tmol/L One-half FAV values
for nonsalmomds varied from 3 6 to 28 6 /jmol/L (acute toxicity information for salmomds is
not given). At typical pH (pH = 8) and temperature (temperature = 20 °C) values for
Onondaga Lake, the minimum FCV value of 1 4 /jmol/L corresponds to a T-NH3 concen-
tration of 36 jttmol/L, this concentration is always exceeded in the lake (Effler et al, 1990)
Onondaga Lake is unusual in being very productive, and so tends to be warmer and
have a higher pH than many lakes At lower pH values (pH = 7) and lower temperatures
(15 °C), the percentage of T-NH3 that is free NH3 drops dramatically (Emerson et al ,
1975), so that the FCV values reported for Onondaga Lake would not be exceeded until a
T-NH3 concentration of 785 /xmol/L was reached Currently no areas of North America are
known to experience rates of NH4 deposition that are sufficient to produce such high
concentrations in surface waters Given current maximal concentrations of NH4+ in
deposition (40 /imol/L; Stensland et al, 1986) and reasonable maximum rates of dry
10-218
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deposition and evapotranspiration (dry deposition equal to 100% of wet deposition and
evapotranspiration equal to 50% of deposition), maximum NH4+ concentrations in surface
waters will be less than 160 jwmol/L If all nitrogen in deposition (NC^" + NH4+) were
ammonified, maximum potential NH4+ concentrations attributable to deposition would be
approximately 280 jimol/L, and would be unlikely to be toxic except in unusual
circumstances Because NH4+ is rapidly oxidized to NO3~ in watershed soils and under
well-oxygenated conditions in lakes and streams, the likelihood of reaching toxic
concentrations are extremely limited Toxic levels would be more likely in systems that have
oxygen deficits, high organic matter loading (which would increase oxygen demand and
contribute ammonium through mineralization processes), and direct inputs of NH3 (i e , near
feedlot operations) In such cases, it would probably be more effective to remove the local
causes of oxygen depletion and organic matter loading, than to reduce atmospheric inputs of
nitrogen It appears from the information above that the potential for directly toxic effects
attributable to nitrogen deposition in the United States is very limited
10.9 DISCUSSION AND SUMMARY
10.9.1 Introduction
Since the mid-1980s, the view has emerged that the deposition of atmospheric inorganic
nitrogen has impacted aquatic and terrestrial ecosystems (Aber et al , 1989, EUenberg, 1987,
Van Breeman and Van Dijk, 1988) It is known that in many areas of the United States, the
atmospheric input of nitrogen compounds has been significant (U S Environmental
Protection Agency, 1982, Sections 10 4 and 10 7 2), however, the impacts have generally
been unknown or considered benign Although, the e\idence linking nitrogen deposition
with ecological impacts has been tenuous, there has been a growing concern (Skeffington and
Wilson, 1988) This concern has been magnified because continuous deposition of
atmospheric concentrations of nitrogen compounds (particularly HNO3 and NO3") in North
America and most European countries has resulted in ecosystems once limited by nitrogen
receiving nitrogen in excess of plant and microbial demand These concerns have led to the
efforts in Europe to develop "critical loads" of nitrogen for various ecosystems A critical
load is defined as "a quantitative estimate of an exposure to one or more pollutants below
10-219
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which significant harmful effects on specified sensitive elements of the environment do not
occur according to present knowledge" (Nilsson and Grennfelt, 1988) The concept of
critical load has not received wide acceptance in North America Current information
indicates that "nitrogen-saturated" forests are relatively rare and limited in extent, especially
managed forests In addition, because of the great variation in both natural forest nitrogen
uptakes and management intensity, it is not reasonable to assign one critical load to all forest
ecosystems
10.9.2 Ecosystems
Ecosystems are composed of populations of "self-supporting" and "self-maintaining"
Irving plants, animals, and microorganisms interacting among themselves and with the
nonliving chemical and physical environment within which they exist (Odum, 1989, Billings,
1978; Smith, 1980) Ecosystems usually have definable limits and may be large or small
(e g., fallen logs, forests, grasslands, cultivated or uncultivated fields, ponds, lakes, rivers,
estuaries, oceans, the earth) (Odum, 1971, Smith, 1980, Barbour et al, 1980) The
environmental conditions of a particular area or region determine the boundaries of the
ecosystem as well as the organisms that can live there (Smith, 1980) Together, the
environment, the organisms, and the physiological processes resulting from their interactions
form the life-support systems that are essential to the existence of any species on earth,
including man (Odum, 1989)
Human welfare is dependent on ecological systems and processes Natural ecosystems
are traditionally spoken of in terms of their structure and functions Ecosystem structure
includes the species (richness and abundance) and their mass and arrangement in an
ecosystem. This is an ecosystem's standing stock—nature's free "goods" (Westman, 1977)
Society reaps two kinds of benefits from the structural aspects of an ecosystem (1) products
with market value such as fish, minerals, forest and pharmaceutical products, and genetic
resources of valuable species (e g , plants for crops, timber, and animals for domestication)
and (2) the use and appreciation of ecosystems for recreation, aesthetic enjoyment, and study
(Westman, 1977).
Structure within ecosystems involves several levels of organization The most visible
are (1) the individual and its environment, (2) the population and its environment, and (3) the
10-220
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biological community and its environment, the ecosystem (Billings, 1978) Ecosystems
function as energy and nutrient transfer systems Through the process of photosynthesis,
vegetation accumulates, uses, and stores carbon compounds (energy) to maintain
physiological processes and to build plant structure Carbohydrates and other compounds
accumluated and stored by plants are the basic source of food (energy and nutrients) for the
majority of animals and microorganisms Energy moves unidirectionally and ultimately
dissipates into the environment Nutrients are recycled into the system Because the various
ecosystem components are chemically interrelated, stresses placed on individual components,
such as those caused by nitrogen deposition and loading, can produce perturbations that are
not readily reversed and will significantly alter the ecosystem (Gudenan and Kueppers,
1980)
10.9.3 The Nitrogen Cycle
Nitrogen, one of the main constituents of the protein molecules essential to all life, is
recycled within ecosystems (see Section 10 1) Most organisms cannot use the molecular
nitrogen found in the earth's atmosphere It must transformed by terrestrial and aquatic
microorganisms into a form other organisms can use The transformations of nitrogen as it
moves through the ecosystem is referred to as the nitrogen cycle Mature natural ecosystems
are essentially self-sufficient and independent of external additions Modern technology, by
either adding nitrogen or removing nitrogen from ecosystems, can upset the relationships that
exist among the various components, and thus change their structure and functioning
10.9.4 Nitrogen Deposition
The removal (dry deposition) of reactive nitrogen gases from the atmosphere occurs
along several pathways leading to foliage, bark, or soil, with pathways to foliage being pre-
dominant during the growing season The prevalence of any particular type of deposition is
a function of (1) the physicochemical properties of nitrogen compounds, (2) their ambient
concentration, and (3) the presence of suitable receptor sites in the landscape (e g , leaves
with open stomata) Average canopy-level measurements (Table 10-29) exhibit the following
pattern or tendency towards dry deposition HNO3 > NH3 = NO2 > NO Although the
leaf-level data for crops are incomplete (NO and HNO3, data are not available), the leaf
10-221
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TABLE 10-29. MEAN DEPOSITION CHARACTERISTICS OF REACTIVE
NITROGEN GASES AT THE LEAF OR CANOPY SCALE OF
RESOLUTION FOR CROP OR TREE SPECIES
Compound
Summary for Crop Species
Nitric Oxide
Nitrogen Dioxide
Nitric Acid
Ammonia
Summary for Tree Species
Nitric Oxide
Nitrogen Dioxide
Nitric Acid
Ammonia
Leaf-Level Measures
Kx (mm/s)a
NDb
12
NDb
45
<03
1 1
2 1
1 8
Canopy-Level Measures
Vd (mm/s)a
13
77
198
66
NDb
24
41
22
"Means are the average for all species studied However, measurements on dormant plant materials, foliage
with low stomatal conductance, and data recorded in the dark were excluded The values listed as Kj (leaf
conductance) and Vj (deposition velocity) for particles represent the leaf-wash and throughfall measurement
techniques, respectively
ND = No data were available
conductance (K^) data for trees shows a similar pattern These patterns are consistent with
the observations of Bennett and Hill (1973), and can be partially explained by gas solubility
characteristics (Taylor et al, 1988) Particle deposition data averaged across species and
experimental techniques shows approximately three tunes greater nitrate aerosol deposition
(7.8 mm/s) than for ammonium (2 mm/s) However, the high average Vd for NO3" is
probably excessively high due to the unavoidable inclusion of nitrate from HNO3 in
measurements of nitrate deposition
With the possible exception of HNO3 vapor, deposition characteristics of reactive
nitrogen compounds are highly variable and dramatically influenced by environmental
conditions that affect stomatal conductance The tight relationship between stomatal
10-222
-------
conductance and the deposition of NO and NO2 implies that gaseous deposition of reactive
NOX is greatly reduced in the dark, when stomata close (Hanson et al, 1989, Saxe, 1986,
Hutchinson et al, 1972) Deposition of gaseous nitrogen forms is usually proportional to
ambient concentrations, but "compensation concentrations" at which no uptake occurs (i e ,
< 0 003 to 0 005 ppmv) have been reported for NO2 and NH3 Data for NO, NO2, and
HNO3 (Grennfelt et al, 1983, Johansson, 1987, Marshall and Cadle, 1989, Skarby et al,
1981), from the vegetation dormant period, show a reduced potential for deposition
Conversely, particulate nitrate and ammonium deposition do not appear to be affected by the
season of the year (Gravenhorst et al, 1983, Lovett and Lindberg, 1984)
The preceding information on gases and particles indicates that methods for measuring
gas or particulate deposition may produce dramatically different results Leaf-level measures
of deposition (Kj) for NO, NO2, and HNO3 were 4 to 10 tunes lower than estimates obtained
using micrometeorological canopy-level measurements (Vd) This discrepancy can largely be
explained once canopy area instead of ground area is factored into the canopy-based
measurements
The canopy-level Vd measurement has been criticized because it attempts to pool
environmental, physiological, and morphological characteristics into a single descriptive
measurement (i e , it attempts to do too much, Taylor et al, 1988) The result of this over
simplification is that Vd for even a single trace gas vanes substantially in space and tune
However, average K^ and Vd values for NH3 on crop species were comparable, perhaps
because crop canopies are more uniform and closer to the ground Particle deposition is
governed by a different set of principles (see Section 1C) 2 3) and the same relationships
between leaf and canopy level measurements may not be applicable
Daytime rates of NOX or NH3 deposition can also be approximated from ambient
concentrations of the gases (U S Environmental Protection Agency, 1982, Hicks et al ,
1985) and deposition constants such as those presented in Table 10-29 Hanson et al (1989)
used such information with conservative estimates of concentration to approximate total
nitrogen deposition from NO2 to various forest stands They predicted NO2-mtrogen inputs
between 0 04 and 1 9 kg nitrogen/ha/year for natural forests and inputs up to 12 kg
nitrogen/ha/year for forests in urban environments For a forested watershed, Grennfelt and
Hultberg (1986) calculated the annual deposition of NO2 plus HNO3 to be in the range from
10-223
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3.6 to 5 1 kg nitrogen/ha/year Hill (1971) estimated the removal of NO2 from the
atmosphere in Southern California to be approximately 109 kg nitrogen/ha/year
Preliminary particle deposition measurements and calculated dry deposition estimates of
reactive nitrogen gases indicate significant nitrogen inputs to terrestrial systems Barne and
Sirois (1986) estimated that dry deposition contributed 21 to 30% of total NO3" deposition in
eastern Canada Lovett and Lindberg (1986) concluded that dry deposition of nitrate is the
largest form of inorganic nitrogen deposited to oak-hickory forests in eastern Tennessee
Annual estimates of NH3 deposition have been reported (Cowling and Lockyer, 1981,
Sinclair and Van Houtte, 1982), but numerous reports of NH3 evolution from foliage under
conditions of high soil nitrogen confound simple estimates of annual NH3-mtrogen
deposition Lovett (1992) summarized research data for a number of forested sites in North
America and Norway and concluded that dry deposition of nitrogen typically occurs at annual
rates approximately equal to nitrogen deposited in precipitation
Because gaseous deposition is difficult to measure accurately or continuously at the
landscape level of resolution, estimates of dry nitrogen deposition must rely on models
Rigorous models of pollutant deposition have been developed (Hicks et al , 1985, Baldocchi,
1988; Baldocchi et al, 1987) and will be needed in the future for accurate determination of
reactive nitrogen gas and particle deposition to forest stands and ecosystems Although
progress has been made in understanding and modeling the processes that control the dry
deposition of nitrogen containing compounds, additional research will be required to
minimize errors in predictions of total dry nitrogen deposition to specific regions and under a
range of environmental conditions
Increased efforts have been made to establish both wet and dry deposition rates of
nitrogen to various types of ecosystems These current deposition data are important because
they provide a basis for evaluating potential effects against "suggested critical levels"
Although the concept of critical nitrogen loading has not been widely adopted in North
America, for reasons discussed in Sections 10 5 8, 10 5 9, and 10 6 3 1, a comparison of
total nitrogen deposition data for North America with proposed critical loads for Europe
provide a comparison of the status of terrestrial systems with respect to changes that might
be expected from elevated levels of nitrogen deposition Figure 10-19 summarizes wet
deposition data for nitrate and ammonium in the United States Because the data are for wet
10-224
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10-225
-------
deposited forms of nitrogen, they represent an underestimate of the total nitrogen deposition
to the ecoystems Table 10-14 summarizes information regarding the total (wet and dry)
deposition of nitrogen to a variety of ecosystems/forest types or regional areas in North
America and Europe
10.9.5 Effects of Deposited Nitrogen on Soils
The effects of nitrogen deposition on biological systems must be viewed from the
perspective of the amount of nitrogen in the system, the biological demand for nitrogen, and
the amount of deposition If nitrogen is deposited on a nitrogen-deficient ecosystem, a
growth increase will likely occur If nitrogen is deposited on a ecosystem with adequate
supplies of nitrogen, nitrate leaching will eventually occur Nitrate leaching is usually
deemed undesirable in that it can contaminate groundwater and lead to sod acidification
This analysis focuses on forest ecosystems, but considers and ecosystems as well
Agricultural lands are excluded from this discussion because crops are routinely fertilized
with amounts of nitrogen (100 to 300 kg/ha) that far exceed pollutant inputs even in the most
heavily polluted areas Pollutant nitrogen inputs to grasslands and and soils can be expected
to produce increased growth in some instances, despite water limitations (e g , Fisher et al,
1988c). However, these systems are obviously not subject to the soil acidification and
groundwater NO3" pollution problems that might occur in more humid areas Excess
nitrogen deposited on these ecosystems leave via either demtnfication or NH4+ volatilization
(see review by Woodmansee, 1978)
The biological competition for atmospherically deposited nitrogen among heterotrophs
(decomposing microorganisms), plants, and mtnfymg bactena, combined with the chemical
reactions between NH4 and humus in the soil, determine the degree to which vegetation
growth increase will occur and the degree to which incoming nitrogen is retained within the
ecosystem Until recently, mtnfymg bactena were thought to be poor competitors for
nitrogen, with heterotrophs being the most effective competitors and plants being
intermediate Recent studies of soil nitrogen dynamics using N (Davidson et al, 1990) and
thorough analyses of forest nitrogen budgets suggest that these assumptions and perhaps our
conceptual model of soil nitrogen cycling need modification Specifically, nitrification may
be proceeding at a significant level without the appearance of NO3" in soils or soil solution if
10-226
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TABLE 10-14. MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL NITROGEN DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Forms of Nitrogen Deposition (kg/ha)a
Site Location/
Vegetation
United States
California, Chaparral
California, Sierra Nevada
Georgia, Loblolly pine
North Carolina, Loblolly pine
North Carolina, Hardwoods
North Carolina, White pine
North Carolina, Red spruce
New Hampshire, Deciduous
New Hampshire, Deciduous
New York, Red spruce
New York, Mixed deciduous
Tennessee, Mixed deciduous
Tennessee, Oak forest #1
Tennessee, Oak forest #2
Tennessee, Oak forest #1
Tennessee, Oak forest #2
Tennessee, Oak forest
Tennessee, Loblolly pine
Washington, Douglas fir
Washington, Douglas fir
U S Regions
Adirondacks
Midwest
Northeast
Northwest
Southeast
Southeast Appalachians
Wet
Cloud Rain
82
—
37
87
48
37
87 62
70
93
73 61
48
29
32
29
69
60
45
43
29
10
63
42
21 7
166
206
42
Dry
Particles
—
10
22
05
09
36
—
—
02
08
4 1
44
44
1 3
1 2
1 8
06
1 3
"
47
29
—
—
—
3 1
Gases Total
23C
(2)
42 9
4 1 15
53
27 7
86 27
(7)
(9)
23 16
25 8
61 13
40 12
40 11
8
7
38 10
14 9
06 5
(1)
11
7 1
22
17
21
73
Reference
Rigganetal (1985)
Williams and Melack
(1991a)
Lovett (1992)
Lovett (1992)
Swank and Waide (1988)
Lovett (1992)
Lovett (1992)
Likens et al (1970)
Likens (1985)
Lovett (1992)
Lovett (1992)
Kelly and Meagher
(1986)
Kelly and Meagher
(1986)
Kelly and Meagher
(1986)
Kelly (1988)
Kelly (1988)
Lindbergetal (1986)
Lovett (1992)
Lovett (1992)
Henderson and Hams
(1975)
Dnscolletal (1989a)
Dnscolletal (1989a)
Munger and Eisenreich
(1983)
Munger and Eisenreich
(1983)
Munger and Eisenreich
(1983)
Dnscoll et al (1989a)
10-227
-------
TABLE 10-14 (cont'd). MEASUREMENTS OF VARIOUS FORMS OF
ANNUAL NITROGEN DEPOSITION TO NORTH AMERICAN AND
EUROPEAN ECOSYSTEMS
Site Location/
Vegetation
Forms of Nitrogen. Deposition (kg/ha)
Wet
Dry
Cloud Rain
Particles Gases Total Reference
Canada
Alberta (southern)
British Columbia
Ontario
Ontario (southern)
73
55
37
23
122
14
19 5 Peake and Davidson
(1990)
(5) Feller (1987)
(4) Linseyetal (1987)
37 Roetal (1988)
Federal Republic of Germany
Spruce (Southeast slope)
Spruce (Southwest slope)
Netherlands
Oak-birch
Deciduous/spruce
Scots pine
Douglas fir
Douglas fir
165
243
193
95 7U
165 Hantschel et al (1990)
243 Hantschel et al (1990)
24-56 Van Breemen and Van
Dijk (1988)
21-42° Van Breemen and Van
Dijk (1988)
17-64° Van Breemen and Van
Dijk (1988)
17-64° Van Breemen and Van
Dijk (1988)
115 Draayers et al (1989)
Norway
Spruce
United Kingdom
Spruce forest
Cotton grass moor
a— Symbolizes data not available or,
10 3 07
19 80
04 80
in the case of cloud deposition,
02 112
3-19°
13 5 23 4
40 124
not present
Lovett (1992)
Royal Society (1983)
Fowler et al (1989a)
Fowler et al (1989a)
total nitrogen deposition and are enclosed in parentheses
'Total nitrogen deposition was based on bulk deposition and throughfall measurements and does include
components of wet and dry deposition
Includes deposition from gaseous forms
10-228
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NO3~ is rapidly taken up by heterotrophs It is also clear that trees can be very effective
competitors for atmospherically deposited nitrogen in nitrogen-deficient ecosystems Finally,
the role of chemical reactions between NH4+ and humus need to be investigated, such
reactions have been shown to be very important in fertilization studies, and they may also
play a major role in unfertilized ecosystems If this is Ihe case, the fundamental assumption
that nitrogen retention is controlled primarily by biological processes may be erroneous
Nitrification and NO3" leaching become significant only after heterotroph and plant
demand for nitrogen are substantially satisfied, a condition that has been referred to as
"nitrogen-saturated" Nitrogen-saturated forest ecosystems are very rare in the United States,
but do occur in some slow-growing, high-nitrogen input areas (e g , high-elevation southern
Appalachians) Additions of nitrogen in any biologically available form (NH4+, NO3~, or
organic) to a nitrogen-saturated system will cause equivalent leaching of NO3", except in
those very rare systems where nitrification is inhibited by factors other than competition from
heterotrophs and plants Considering the effects of NO3" only will result in a substantial
underestimation of the acidification potential of atmospheric deposition in nitrogen-saturated
ecosystems
Vegetation demand for nitrogen depends on a number of growth-influencing factors
including temperature, moisture, availibility of other nutrients, and stand age Uptake rates
decline as forests mature, especially after the cessation of the buildup of nutrient-rich foliar
biomass following crown closure Thus, nitrogen-saturation tends to be more common in
older forests than in younger forests because nitrogen demand is less Processes that cause
net nitrogen export from ecosystems, such as fire and harvesting, will naturally push
ecosystems toward a state of lower nitrogen-saturation or even nitrogen deficiency. Intense
fires cause a large loss of ecosystem nitrogen capital, but frequent, low-intensity fires may
have little effect
A review of the literature on forest fertilization and nitrogen-cycling studies under
various levels of pollutant nitrogen input reveals some interesting contrasts that pertain to the
the relative roles of heterotrophs, plants, and nitnfiers discussed above Forest fertilization
has proven quite successful in producing growth increases in nitrogen-deficient forests, even
though trees typically recover only 5 to 50% of fertilizer nitrogen (Table 10-12) On an
ecosystem level, however, retention of nitrogen is usually quite high (often 70 to 90 % of
10-229
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applied nitrogen, Table 10-12), primarily due to fertilizer nitrogen retention in the litter and
soil, including nonbiological reactions between NH4+ and humus Fertilization studies differ
from pollutant nitrogen deposition in several important respects (1) pollutant nitrogen
deposition enters the ecosystem at the canopy level, whereas fertilizer is typically (but not
always) applied to the soil, (2) fertilization leads to high concentrations of NH4+ and, in the
case of urea, high pH, both of which are conducive to nonbiological reactions between soil
humus and NH4+, and (3) pollutant nitrogen deposition enters the ecosystem as a slow,
steady input in rather low concentrations, whereas the fertilizer is typically applied in one to
five large doses Both plants and nitrifying bacteria are favored by slow, steady inputs of
nitrogen, possibly giving them a competitive advantage over heterotrophs for pollutant
nitrogen inputs A review of the literature on nitrogen cycling in unfertilized forests, with
differing levels of pollutant nitrogen input supports this hypothesis Ecosystem-level
recovery of atmospherically deposited nitrogen (typically less than 50% and often 0%,
Table 10-13 and Figure 10-8) is lower than of fertilizer nitrogen (typically 70 to 90% of
applied nitrogen, Table 10-12 and Figure 10-11) It also appears that vegetation retention of
incoming nitrogen in unfertilized forests is somewhat higher than in fertilized forests,
whereas soil (heterotroph) retention of atmospherically deposited nitrogen is much lower
In forests with very low atmospheric nitrogen inputs, it appears as if the soil is being
"mined" for the nitrogen necessary to supply vegetation, an indication that plants are actually
out-competing heterotrophs for nitrogen In forests with high atmospheric nitrogen inputs,
heterotrophic nitrogen uptake appears to be minimal, perhaps because of limitations by
organic substrates or other nutrients
Because nitrification results in the creation of HNO3 within the soil, there are concerns
that elevated nitrogen inputs to nitrogen-saturated systems will result in soil acidification and
Al mobilization There are very few proven, documented cases in which excessive
atmospheric nitrogen deposition has caused soil acidification (e g , in forests in the
Netherlands subject to very high nitrogen deposition levels, 40 to 80 kg/ha/year), but there is
no doubt that the potential exists for many mature forests with low uptake rates, given high
enough inputs for a sufficiently long tune The amount of nitrogen deposition required will
vary with the ecosystem The greatest uncertainty in assessing and projecting rates of soil
10-232
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10-234
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2,000-
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0 1,000 2,000
Fertilizer Input (kg/ha)
Figure 10-8. Ecosystem recovery of fertilizer nitrogen as a function of fertilizer
nitrogen input.
Source Johnson (1992)
acidification is the estimation of weathering rates (i e , the release of base cations from
primary minerals)
Soil acidification is usually thought of as an undesireable effect, but in some cases, the
benefits of alleviating nitrogen deficiency clearly outweigh the detriments of soil acidification
(e g , the benefits of nitrogen fixation by red alder always outweigh the detriments of soil
acidification to succeeding Douglas fir stands in the Pacific Northwest)
10-235
-------
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Atmospheric N Input (kg/ha/year)
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Figure 10-11. Ecosystem nitrogen retention as a function of atmospheric nitrogen input.
Source Johnson (1992)
f\
Increased concentrations of NO3" or any other mineral acid amon (e g , SO4 or Cl") in
soil solution lead to increases in the concentrations of all cations in order to maintain charge
balance in solution Equations describing cation exchange in soils dictate that as the total
amon (and cation) concentrations increase, individual cation concentrations increase as
follows A13+ > Ca2+ andMg2+ > K+, Na+, and H+ Thus, soil-solution Al3 +
o i
concentrations increase not only as the soil acidifies (i e , as the proportion of Al on the
10-236
-------
exchange complex increases) but also as the total ionic concentration of soil solution
increases
3 -f"
There are several cases in which Al concentrations in natural waters have been
shown to be positively correlated with NO3" concentrations Ulrich (1983) noted
NO3" - A13+ pulses in soil solutions from the Soiling site in Germany during warm, dry
o I
years He hypothesized that these nitrate-induced Al pulses caused root injury and were a
major contributor to what he termed "forest decline" observed in Germany during the
mid-1980s This hypothesis is disputed by other German forest scientists who point out that
forest decline occurred on base-rich as well as base-poor soils (the base-rich soils not being
subject to A13+ pulses) (e g , Rehfuess, 1987), Van Breemen et al (1982, 1987) and
3 -J-
Johnson et al (1991) noted NO3" - Al pulses in soil solutions from forest sites in the
Netherlands and in the Smoky Mountains of North Carolina Aluminum toxicity is one of
several nitrogen-related hypotheses posed to explain what has been termed forest decline in
both countries Other hypotheses include weather extremes and climate change, Mg and
K deficiencies that occur in sites naturally low in these nutrients, and foliar damage due to
acid mist Researchers on aquatic effects of acid deposition have long noted springtime
3+ +
pulses of NO3 , Al , and H in acid-affected surface waters of the northeastern United
States (Galloway et al., 1980, DnscoU et al, 1989b)
10.9.6 Effects of Nitrogen on Ecosystems
Ecosystems respond to environmental stresses through their constituent organisms (see
Section 10 1) Plant populations, when exposed to any environmental stress, can exhibit four
different reactions (1) no response—the individuals are resistant to the stress, (2) the most
severe response—mortality of all individuals and local extmction of the extremely sensitive
populations, (3) physiological accommodation—growth and reproductive success of
individuals are unaffected because the stress is physiologically accommodated, and
(4) differential response—members of the population respond differentially, with some
individuals exhibiting better growth and reproductive success due to genetically determined
traits (Taylor and Pitelka, 1992, Garner, 1992) The primary effect of air pollution on the
more susceptible members of the plant community is that they can no longer compete
effectively for essential nutrients, water, light, and space, and are eliminated The extent of
10-237
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change that may occur in a community depends on the condition and type of community, as
well as the pollutant exposure (Garner, 1992)
Plant responses are foliar or soil mediated Subsequent to the dry and wet deposition of
nitrogen forms from the atmosphere (Section 10 4), nitrogen-containing compounds can
impact the terrestrial ecosystems when they enter plant leaves and alter metabolic processes
(Chapter 9) or by modifying the nitrogen cycle and associated soil chemical properties
(Section 10 5) Changes in biochemistry that result in reduced vigor and growth and
decrease the plant's ability to compete for light, water, space, and nutrients can be
manifested as changes in plant populations, communities, and, ultimately, ecosystems
(Chapter 9, Section 10 2) Interpretation of the effects of wet- and dry-deposited nitrogen
compounds at the ecosystem level is difficult because of the interconversion of nitrogen
compounds and the complex interactions that exist between biological, physicochemical, and
climatic factors (Section 10 2, U S Environmental Protection Agency, 1982) Nevertheless,
reactive nitrogen compounds have been hypothesized to impact ecosystems through
modifications of individual plant physiological processes upon entering plants through the
foliage, or through alterations in the nitrogen status of the ecosystem
Very little information is available on the direct effects of HNO3 vapor on vegetation,
and essentially no information is available on its effects on ecosystems Norby et al (1989)
reported that HNO3 vapor (0 075 ppmv) induced NRA in red spruce foliage The effects of
NH3, a reduced nitrogen gas, have been summarized by Van der Eerden (1982), however,
NH3 concentrations seldom reach phytotoxic levels in the United States (U S Environmental
Protection Agency, 1982) In contrast, high NH3 concentrations have been observed in
Europe (Van Dijk and Roelofs, 1988) Van der Eerden (1982) summarized available
information on the direct response of crop and tree species to NH3 fumigation and concluded
that the following concentrations produced no adverse effects 0 107 ppmv (75 /*g/m3)
yearly average, 0.858 ppmv (600 /-tg/m3) daily average, and 14 3 ppmv (10,000 j^g/m3)
hourly average Submicron ammonium sulfate aerosols have been shown to affect foliage of
Phaseolus vulgans L (Gmur et al, 1983) Three-week exposure to a concentration of
•2
26 mg/m (37 ppmv) produced leaf chlorosis, necrosis, and loss of turgor
Because current ambient concentrations of NO, NO2, and NH3 are low across much of
the United States, except in certain highly populated urban areas, significant direct effects of
10-238
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these nitrogen compounds on ecosystems seems unlikely at the current time Concentration
and effects data are unavailable for making similar conclusions regarding other reactive
nitrogen compounds like HNO3 vapor or the gaseous aitrate radical
Serious consideration is currently being given to hypotheses that excess total nitrogen
deposition may impact plant productivity directly or thiough changes in soil chemical
properties Furthermore it has been proposed that excess nitrogen deposition to ecosystems
can modify interplant competitive balances, leading to changes in species composition and/or
diversity The uptake of nitrogen and its allocation is of overriding importance in plant
metabolism and governs, to a large extent, the utilization of phosphorus, potassium, and
other nutrients, and plant growth Nitrogen is the mineral nutrient that most frequently limits
growth in both agricultural and natural systems (Chapin et al, 1987) Normally, the
acquisition of nitrogen is a major carbon expense for plants Plants expend a predominant
fraction of the total energy available to them in the form of carbohydrates in the acquisition
of nitrogen Absorption of nitrogen from the soil requires constant and extensive root
growth to meet the needs of a rapidly growing plant because soil pools of nitrogen,
ammonium, or nitrate in the immediate vicinity of the roots aie usually so small that they are
quickly depleted (Section 10 3)
Increased nitrogen deposition has been associated with changes in the following plant
and soil processes involved in nutrient cycling (1) plant uptake and allocation, (2) litter
production, (3) immobilization (includes the processes of ammomfication [the release of
ammonium] and nitrification [the conversion of ammonium to nitrate during the decay of
litter and soil organic matter]), (4) NO3" leaching, and (5) trace gas emission (Aber et al ,
1989, Figure 10-17) Changes in tree physiology include altered nutrient uptake and
carbohydrate allocation, which directly alters the rate of photosynthesis and influences
growth rate and mycorrhizae formation, and increased leaf nitrogen (Chapin et al , 1987,
Waring, 1985) Susceptibility to insect and disease attick have also been attributed to
alteration in tree physiology (Chapin et al, 1987, Waring, 1987, Shigo, 1973, Hollis et al,
1975, Weetman and Hill, 1973)
Increased nitrogen inputs can affect tree resistance to insects and disease either
positively or negatively Alleviating nitrogen deficiency may increase plant resistance to
pathogen attack, but it may also reduce the production of phenols in plant tissues, thereby
10-239
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Process altered by
nitrogen deposition
Deposition
Photosynthesis
X
Animal
Proteins
Soil
i
V
/ 4
Bacterial
Nitrogen
Rxafion
* K
Litter
Production
(Death)
-
DeattA
| \
'A *
Mlcroblal
Decomposition
,
V
\
Trace
Gas
Emissions
V
V
Figure 10-17. Nitrogen cycle (dotted lines indicate processes altered by chronic nitrogen
deposition).
Source Garner (1992)
reducing resistance to pathogen attack To date, there is little research to show how
increased nitrogen inputs affect susceptibility to pathogen attack, but the potential for either
increased susceptibility or protection is significant
The nitrogen-photosynthesis relationship is critical to the growth of trees because in the
leaves of plants with C3 photosynthesis (the pathway used by most of the world's plants),
approximately 75 % of the total nitrogen is contained in the choloroplasts and is used during
photosynthesis (Chapin et al, 1987) As a rule, plants allocate resources most efficiently
when growth is equally limited by all resources When a specific resource such as nitrogen
limits growth, plants adjust by allocating carbohydrates to the organs that acquire the most
10-240
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strongly limiting resources, however, when nitrogen is abundant, allocation is to the
formation of new leaves
Plants do not necessarily benefit from added nitrogen More nitrogen in the soil is not
mirrored by increased uptake except at low levels (Section 10 3) Among boreal and
subalpine conifers and other vegetation adapted to resouice-poor environments, nitrogen
added to the soil may not increase growth The nitrate reductase enzyme activity in roots
and shoots determines the pattern of nitrate assimilation The photosynthetic capacity of
conifer foliage is low and not greatly enhanced by increasing the nitrogen content (Waring
and Schlesinger, 1985) High leaf nitrogen content is not always an advantage when other
resources, among which are light and water, are limited When photosynthesis is measured
at light saturation, leaf nitrogen is closely correlated with photosynthetic capacity But when
light is low, photosynthesis increases very little, if at all, with increasing leaf nitrogen
(Chapm et al, 1987) In dense conifer forests, lack of sunlight makes the metabolic
conversions of nitrate inefficient because production of large amounts of carbohydrates and
other light-driven reactions become limiting (Zeevaart, 1976) When nitrogen is no longer
limiting, deficiencies of other nutrients may occur (Abet et al , 1989, Kenk and Fischer,
1988) Competition, under the above circumstances, favors deciduous tree species, plants
characteristic of resource-rich environments, rather than conifers (Waring, 1987)
Excessive NH4+ deposition (40 to 80 kg/ha/year) 1o soils in which nitnfication is
inhibited causes serious nutnonal unbalances and even toxic effects to some forests in the
Netherlands (Boxman et al, 1988) Deleterious effects of excess nitrogen deposition (40 to
80 kg/ha/year) can occur via aboveground processes as well K and Mg deficiencies in
declining Dutch forests are thought to be caused by excessive foliar leaching due to high
inputs of NH4+ (Roelofs et al, 1985)
Growth responses to increased nitrogen inputs resulted in changes in species
composition in ecosystems in the Netherlands (Van Breeman and Van Dijk, 1988) Species
respond differentially to increased nitrogen availability, creating the potential for changes in
ecosystem composition with increased nitrogen loading Changes from heathland to
grassland in Holland have been attributed to current rates of nitrogen deposition (Roelofs
et al, 1987) Ellenberg (1987) points to further species changes in Central European
ecosystems as a likely consequence of elevated nitrogen He states that "More than 50 % of
10-241
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the plant species in Central Europe can only compete on stands that are deficient in nitrogen
supply."
De Temmerman et al (1988) found increased fungal outbreaks and frost damage on
several pine species exposed to very high NH3 deposition rates (> 350 kg/ha/year)
Numbers of species and fruiting bodies of fungi have also decreased concomitantly with
nitrogen deposition in Dutch forests (Van Breemen and Van Dijk, 1988) Schulze (1989)
presents a clear progression of evidence that indicates that canopy uptake of nitrogen together
with root uptake has caused a nitrogen unbalance in Norway spruce, leading to its decline
^
Excessive nitrogen inputs to terrestrial ecosystems can cause differential competitive
advantage among plants within a heathland (Heil and Bruggink, 1987, Heil et al , 1988)
In unmanaged heathlands in the Netherlands, Calluna vulagns is being replaced by grass
species as a consequnce of the eutrophic effect of acidic rainfall and large nitrogen inputs
arising from intensive farming practices in the region Calluna is an evergreen with a long
growing season, which normally permits it to compensate for its slow growth rate so that it
competes successfully with the faster growing Molmia (grass) under normal nutrient-limiting
conditions However, a large increase in the nitrogen supply improves the competitive
advantage of Molmia, increasing its growth rate so that it becomes the dominant species in
the heathland. Roelofs et al (1987) observed that nitrophilous grasses (Molmia and
Descfjampsid) are displacing slower growing plants (Enca and Calluna) on heathlands in the
Netherlands, and suggested that a correlation existed between this change and nitrogen
loading. Van Breemen and Van Dijk (1988) found a substantial displacement of heathland
plants by grasses from 1980 to 1986 and also observed increases in nitrophilous plants in
forest herb layers Ellenberg (1988) suggested that ionic inputs (NO3~ and NH4+) influence
competition between organisms long before toxic effects appear on individual plants These
changes in the Netherlands have occurred under nitrogen loadings of between 20 and 60 kg
nitrogen/ha/year Liljelund and Torstensson (1988) have shown clear signs of vegetation
changes in response to nitrogen deposition rates of 20 kg/ha/year
Evidence is accumulating that the assumed O3-specific effects of forests within the
Los Angeles basui are not strictly the result of O3 exposures but, in part, due to the
co-deposition of oxides of nitrogen, specifically HNO3 The environment is seldom optimal
in either natural or agricultural communities It is not unusual, therefore, for plants growing
10-242
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in natural habitats to encounter multiple stresses Plant responses to multiple stresses depend
on resource (carbon and nitrogen) interactions at levels ranging from the cell to the
ecosystem (Chapin et al , 1987) At the present tune, data dealing with the response of trees
or other vegetation to the combined stresses of O3 exposure above ground and nitrate
deposition through the soil are sparse, however, when the responses of plants exposed to
O3 alone and to nitrate deposition alone are considered, it is possible to conceptualize how
exposure to the two in combination could affect vegetation Both O3 exposure and nitrate
uptake can affect the processes of photosynthesis, carbohydrate allocation, and nutrient
uptake The impact of a reduced carbon supply to the shoot or to the roots and the affect on
subsequent allocation of nitrogen, as well as other nutrients, can be deduced from
Figure 10-17
The importance of the nitrogen-photosynthesis relationship and the allocation of nitrogen
and carbon on plant growth has been discussed in the pievious section Patterns of carbon
allocation directly influence the growth rate (McLaughkn et al, 1982, U S Environmental
Protection Agency, 1986, Garner et al , 1989) The ready availability of nitrogen in the soil
and its uptake influence the process of photosynthesis by increasing carbohydrate demand and
shifting allocation (Figure 10-17) from the roots, to the shoots To increase carbohydrate
production in order to utilize increased leaf nitrogen, plants compensate by producing more
leaves
Exposure to O3 inhibits photosynthesis and increases carbohydrate demand in plants that
already have a high carbohydrate demand Ozone is the most phytotoxic of the ambient air
pollutants Many controlled studies using both herbaceous and woody vegetation have
demonstrated inhibition of photosynthesis and premature senescence of leaves by O3 exposure
(Garner et al , 1989, U S Environmental Protection Agency, 1986) Exposure of sensitive
trees to O3 decreases growth and vigor by inhibiting photosynthesis, decreasing carbohydrate
production and allocation to the roots, and interfering with mycorrhizae formation
(McLaughlin et al , 1982, Tingey and Taylor, 1982, U S Environmental Protection Agency,
1986, Garner et al , 1989)
Both increased soil nitrogen and O3 exposure can affect nutrient uptake When nitrogen
is readily available, other nutnents (e g , phosphorus and calcium) can become limiting
Decreased carbohydrate allocation to roots, a result of O3 exposure, interferes with
10-243
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mycorrhizae formation and, subsequently, nutrient uptake Limiting carbohydrate production
and nutrient availability suppresses growth (McLaughlin et al , 1982, Mooney and Winner,
1988; U.S. Environmental Protection Agency, 1986) The combined stresses resulting from
increased soil nitrogen and ambient O3 exposure, therefore, have the capability of severely
impacting plant growth
10.9.7 Nitrogen Saturation, Critical Loads, and Current Deposition
Ecosystem nitrogen saturation and the definition of the level of total nitrogen deposition
at which critical changes begin to appear in sensitive ecosystems have been the subject of
recent conferences in Europe (Nilsson and Grennfelt, 1988, Brown et al, 1988, Skefflngton
and Wilson, 1988) The Workshop held at Skokloster, Sweden, in March 1988 (Nilsson and
Grennfelt, 1988) adopted the following definition for a critical load "A quantitative estimate
of an 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 " In the
Skokloster Report (Nilsson and Grennfelt, 1988) and subsequent publications synthesizing
much of the information, nitrogen critical loads were aimed "to protect soils from long-term
chemical changes with respect to base saturation" (Nilsson and Grennfelt, 1988, Schulze
et al, 1989) The critical loads were estimated using two equations Based on the equations
and estimates of the various parameters within them, the authors calculated critical loads for
various forest ecosystems Their values ranged from a low of 3 to 5 kg nitrogen/ha/year for
raised bogs to a high of 5 to 20 kg nitrogen/ha/year It is important to recognize that the
magnitude of such a critical load will be site and species specific because it is highly
dependent on initial soil chemistries and biological growth potentials (i e , nitrogen
demands)
The aim of the nitrogen saturation concept is to make it possible to define a critical
load for nitrogen (deposition rate) at which no change or deleterious impacts will occur to an
ecosystem (Nilsson, 1986) Problems exist, however, with implementing the concept
Establishing a critical load depends on the catena used (e g , one critical load would be
required to prevent species change and another would be required to prevent community
change) (Liljelund and Torstensson, 1988)
10-244
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Skeffington and Wilson (1988) point out that intrinsic in all definitions of a critical load
is the notion that there is a load at which no long-term effects occur The complexity of the
nitrogen cycle and ecosystem diversity make defining a critical load for nitrogen very
difficult The following possible criteria may be useful for defining appropriate critical
nitrogen loads on ecosystems
• prevent nitrate levels in drinking or surface waters from rising above
standard levels,
• ensure proton production less than weathering rate,
• maintain a fixed NH3-base cation balance,
• maintain nitrogen inputs below nitrogen outputs (the nitrogen-saturation approach),
and
• minimize accelerations in the rates of ecological succession (vegetation changes due
to altered interspecific competition)
In summarizing the results of a recent conference on critical nitrogen loading, after
discussing various options for setting a critical nitrogen load, Skeffington and Wilson (1988)
concluded that "we do not understand ecosystems well enough to set a critical load for
nitrogen deposition in a completely objective fashion " Brown et al (1988) further
concluded that there was probably no universal critical load definition that could be applied
to all ecosystems, and a combination of scientific, political, and economic considerations
would be required for the application of the critical load concept
The following terrestrial ecosystems have been suggested as being at risk from the
deposition of nitrogen-based compounds
• heathlands with a high proportion of lichen cover,
• low meadow vegetation types used for extensive grazing and haymaking, and
• coniferous forests, especially those at high altitudes (World Health
Organization, 1987)
The above oligotrophic ecosystems are considered at risk from atmospheric nitrogen
deposition because plant species normally restricted by low nutrient concentrations could gain
a competitive advantage, and their growth at the expense of existing species would change
the "normal" species composition and displace some species entirely (Ellenberg, 1987,
Waring, 1987) Sensitive natural ecosystems, unlike highly manipulated agricultural systems,
may be prone to change from exposure to dry deposited nitrogen compounds because
processes of natural selection whereby tolerant individuals survive may not be keeping pace-
10-245
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with the current levels of atmospheric nitrogen deposition (World Health Organization,
1987)
There is little doubt that nitrogen deposition has had an effect on many ecosystems in
Europe. Kauppi et al (1992) report that biomass of European forest increased during the
1971 to 1990 period This is in stark contrast to earlier claims of forest decline The
authors attribute this growth increase to increases in nitrogen deposition and base their
conclusions on a comparison of the magnitude of increases in nitrogen deposition and
responses shown by European forests to nitrogen fertilizer It is logical to assume that the
same growth increase would occur in many forests in North America (especially western
North America) with increased deposition, given known nitrogen deficiencies and responses
to nitrogen fertilization (Aber et al, 1989, Gessel et al, 1973)
However, because ecosystems have a variable capacity to buffer changes caused by
elevated inputs of nitrogen, it is difficult to make general conclusions about the type and
extent of change (if any) currently resulting from nitrogen deposition in North America
More research needs to be conducted in this area to determine if the hypothesized effects of
excess nitrogen deposition are taking place and to determine the sensitivity of a wide range
of natural ecosystems to nitrogen loading
10.9.8 Effects of Nitrogen on Wetlands and Bogs
The anaerobic (oxygen-free) nature of their waterlogged soils is the feature that sets
wetlands apart Anaerobic wetland soils favor the accumulation of organic matter and losses
of mineral nitrogen to the atmosphere through demtnfication reactions (the conversion of
nitrate to gaseous nitrogen by microbes) Nitrogen deposition can impact plant and microbial
processes either directly or indirectly by acidifying the environment An increase in nitrogen
supply through atmospheric deposition or other means alters the competitive relationships
among plant species such that fast-growing mtrophilous species (species that have a high
nitrogen requirement) are favored Microbial rates of decomposition, nitrogen fixation (the
conversion of gaseous nitrogen to ammonium), nitrification (the conversion of ammonium to
nitrate), and dissimilatory nitrate reduction (conversion to gaseous nitrogen or ammonium)
are all affected Acidification below pH 4 to 5 7 blocks the nitrogen cycle by inhibiting
nitrification, and the accumulation of NH4+ in the environment represses nitrogen fixation
10-246
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(Roelofs, 1986, Schuurkes et al , 1986, 1987, Rudd et al, 1988) The proportion of N2O
produced by demtnfication reactions increases with decreasing pH below 7, and the absolute
rate of production of N2O increases with increasing eutrophication (nutrient enrichment of
the environment) (Focht, 1974) This is potentially important on a global scale because of
chemical reactions with N2O in the atmosphere that result in a loss of O3
The importance of atmospheric nitrogen deposition to the community structure (species
composition and interrelationships) of wetlands increases as rainfall increases as a fraction of
the total water budget Primary production (plant growth) in wetlands is commonly limited
by nitrogen availability Primary production is proportional to the rate of internal nitrogen
cycling, which is influenced by the quantity of minerahzable soil nitrogen as well as the
supply of nitrogen to the ecosystem from the atmosphere or surface flow Total nitrogen
inputs range from about 10 kg nitrogen/ha/year m ombrotrophic bogs (ram-fed bogs), which
receive water only through precipitation, to 750 kg mtiogen/ha/year or more in mtertidal
wetlands with large ground and surface hydrologic inputs
From studies of nine North American wetlands, bulk nitrogen deposition ranges from
5 5 to 12 kg nitrogen/ha/year and occurs in the form of NO3", NH4+, and dissolved organic
nitrogen in roughly equal proportions More recent studies, however, suggest that these rates
are too low and that the wet deposition of NO3" alone is greater than 15 kg nitrogen/ha/year
over much of eastern North America (Zemba et al, 1988) Dry deposition, which probably
accounts for greater than 50% of total deposition, adds to the total Leaf capture of nitrogen
m fog droplets is a third form of deposition that is locally important Applications of
nitrogen fertilizer in the field, ranging from 7 to 3,120 kg mtiogen/ha/year, have increased
standing biomass by 6 to 413 % Other nutrients, like phosphorus, become secondarily
limiting to primary production after nitrogen inputs reach a threshold Fertilization and
increased atmospheric deposition have increased the dominance of grass species over other
plant species in bogs, and extreme eutrophication is associated with a decrease m plant
species diversity
Single additions to vegetated wetland soils of N-labeled mineral nitrogen at rates of
about 100 kg nitrogen/ha/year indicate that up to 93 % of applied NH4 is rapidly assimilated
into organic matter within a single growing season The majority of the labeled nitrogen is
lost from the system after 3 years by the combined processes of advective transport m water
10-247
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(carried in moving water) of particulate organic matter, advective and diffusive transport of
dissolved nitrogen, and demtnfication In the absence of plants, the major fate of inorganic
nitrogen applied to wetland soils is loss to the atmosphere by demtnfication
Peat-forming Sphagnum spp are largely absent from bogs in western Europe where
bulk deposition rates are about 20 to 40 kg nitrogen/ha/year, and soft-water communities
once dominated by isoetids in the Netherlands have been converted to later successional
stages dominated by Juncus spp (rush) and Sphagnum spp or to grasslands Heathlands
dominated by shrubs have also converted to grasslands Experimental studies indicate that
ombrotrophic bogs can be maintained if nitrogen inputs are less than 20 kg nitrogen/ha/year
Increased productivity associated with eutrophication is accompanied by increased rates of
transpiration (evaporation of water from leaf surfaces), which can alter wetland hydrology
and influence the direction of wetland succession By this mechanism, one modeling study
suggests that a succession (change) from open ombrotrophic bog to forested wetland occurs
when a threshold of 7 kg nitrogen/ha/year is exceeded These estimates are consistent with
conclusions from studies of species distributions that place the limit for many species from
10 to 20 kg nitrogen/ha/year (Liljelund and Torstensson, 1988)
Fourteen percent (18 species) of the plant species from the conterminous United States
that are formally listed as endangered, and an additional 284 species listed as potentially
threatened (Code of Federal Regulations, 1987), are found principally in wetland habitats
Some of the endangered plants, like the green pitcher plant, are known to be adapted to
infertile habitats and are threatened by current levels of nitrogen deposition in parts of North
America. Plant species that are threatened by high nitrogen deposition are not confined to
wetland habitats, however, but are common across many ecosystem types (EUenberg, 1988)
10.9.9 Effects of Nitrogen on Aquatic Systems
Nitrogen deposition has not historically been considered a serious threat to the integrity
of aquatic ecosystems
Assessment of the aquatic effects of NOX depends on a close examination of the
processes by which nitrogen may enter streams, lakes and estuaries Sources of nitrogen
may include (1) atmospheric deposition directly to the water surface, (2) deposition to the
watershed that is subsequently routed to the drainage waters, (3) gaseous uptake by plants
10-248
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that is subsequently routed, by way of litter fall and decomposition, to drainage waters, and
(4) nitrogen fixation, either in the water itself, or in watershed soils In addition, numerous
processes act to transform nitrogen species into forms that are only indirectly related to the
original deposition or fixation These transformations include (1) nitrogen assimilation (the
biological uptake of inorganic nitrogen species), (2) nitrification (the oxidation of ammonium
to nitrate), (3) demtnfication (the biological reduction of nitrate to form gaseous forms of
nitrogen, N2, NO, or N2O), and (4) mineralization (the decomposition of organic forms of
nitrogen to form ammonium) The multiple sources of nitrogen to aquatic systems, and the
complexities of nitrogen transformations in water and watersheds, have the effect of
decoupling nitrogen deposition from nitrogen effects, and reduce our ability to attribute
known aquatic effects to known rates of nitrogen deposition Although it is not currently
possible to trace the pathway of nitrogen from deposition through any given watershed and
into drainage waters, we can, in areas of the United States where nonatmosphenc sources of
nitrogen are small, begin to infer cases where nitrogen deposition is having an impact on
aquatic ecosystems
Any discussion of the aquatic effects of NOX must focus on the concept of nitrogen
saturation Nitrogen saturation can be defined as a situation where the supply of nitrogenous
compounds from the atmosphere exceeds the demand for these compounds on the part of
watershed plants and microbes (Skeffington and Wilson, 1988, Aber et al, 1989) Under
conditions of nitrogen saturation, forested watersheds that previously retained nearly all of
nitrogen inputs, due to a high demand for nitrogen by plants and microbes, begin to supply
more nitrogen to the surface waters that dram them Our conceptual understanding of
nitrogen saturation suggests that, in aquatic systems, the earliest stages of nitrogen saturation
will be observable as increases in the seventy and duration of springtime pulses of nitrate
The aquatic effects of NOX can be divided into three general categories
(1) acidification, both chronic and episodic, (2) eutrophication of both fresh waters and
estuaries, and (3) directly toxic effects
10.9.9.1 Acidification
Acidification effects are traditionally divided into chronic (long-term) and episodic
(short-term effects usually observable only during seasons of high runoff) effects Nitrate,
10-249
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the dominant form of inorganic nitrogen in almost all aquatic systems, is commonly present
in measurable concentrations only during winter and early spring, when terrestrial demand
for nitrogen is low because plants in the watershed are dormant Nitrogen will, therefore,
only be a problem in chronic acidification in rare cases where the process of nitrogen
saturation is very much progressed Chronic acidification by nitrogen can be conclusively
demonstrated only hi parts of Europe (e g , Hauhs, 1989, Hauhs et al , 1989, Van Breemen
and Van Dijk, 1988)
Episodic acidification by nitrate is far more common than chronic acidification, and is
well documented for streams (Dnscoll et al , 1987b) and lakes (Galloway et al, 1980,
Dnscoll et al, 1991, Schaefer et al, 1990) in the Adirondack Mountains, for streams in the
Catskill Mountains (Stoddard and Murdoch, 1991, Murdoch and Stoddard, in press b), and
in a small proportion of lakes in Vermont (Stoddard and Kellogg, in press), as well as in
many parts of Canada (Jeffries, 1990) and Europe (e g , Hauhs et al, 1989)
Based on intensive monitoring data, it is possible to divide lakes and streams into three
groups, based on their seasonal NO3" behavior In many parts of the country, nitrogen
demand on the part of the terrestrial ecosystem is sufficiently high that no leakage of NO3"
from watersheds occurs, even when nitrogen deposition rates are relatively high, and cold
temperatures should limit the biological demand for nitrogen Lakes and streams in these
areas show no evidence that nitrogen deposition is producing adverse aquatic effects
In a second group of lakes and streams, NO3" concentrations show strong seasonably,
with peak concentrations during snowmelt or following large ram events In many cases,
these episodic increases in NO3", along with already low baseline ANC are sufficient to
cause short-term acidification and potential adverse biological effects It is important to note
that seasonal increases in NO3" concentrations can be produced by normal watershed
processes; lowered terrestrial demand for nitrogen during the dormant season, for example,
creates a strong likelihood that springtime drainage waters will show NO3" concentrations
that are elevated over summer and fall concentrations Mineralization of organic matter
during the cold months of winter, coupled with low biological demand for nitrogen, can
produce high winter concentrations of NO3" in soil water that is subsequently flushed into
drainage waters during spring snowmelt or during large rain storms Although the seasonal
pattern of elevated NO3" concentrations in this group of lakes and streams can be considered
10-250
-------
normal, the seventy of the NO3" episodes that these systems experience can be strongly
influenced by the amount of nitrogen stored in the snowpack over the course of the winter
If biological demand for nitrogen is still low at the onset of snowmelt, the entire store of
snowpack NO3" can be flushed into drainage waters in the very early stages of snowmelt
(e g , Johannessen and Hennksen, 1978, Jeffries, 1990)
The third group of lakes and streams exhibits both the strong seasonally in NO3"
concentration described in the previous paragraph, and increasing trends in NO3"
concentrations Because the early stages of nitrogen saturation are expected to produce
increases in NO3" concentrations, especially during episodes, long-term increases in NO3"
may represent the strongest evidence that nitrogen deposition is responsible for aquatic
effects In all cases where increasing trends in NO3" have been documented in the
United States (Smith et al , 1987b, Stoddard and Murdoch, 1991, Murdoch and Stoddard, in
press b, Dnscoll and Van Dreason, in press), they have occurred at a tune when nitrogen
deposition is relatively constant (e g , Simpson and Olsen, 1990) Increased leakage of NO3"
from watersheds in these areas, therefore, represents a long-term decrease in the ability of
watersheds to retain nitrogen A likely cause of such long-term changes is a lowering in the
demand for nitrogen as a nutnent on the part of the terrestrial ecosystem, which may result
from long-term high rates of nitrogen deposition to affected watersheds (e g , Aber et al,
1989), forest maturation (Elwood et al , 1991), or, more likely, a combination of both
factors
The locations of lake and stream sites in each of the three NO3" groups are shown on
maps of the Northeast (Figure 10-36), the Southeast (Figure 10-37), and the West
(Figure 10-38) In order to assess which lake and stream sites fall into each group, it was
necessary to have data collected over several years (at least 3 years) and on a relatively
intensive sampling schedule (at least four tunes per year, to illustrate seasonal patterns)
These criteria exclude many sources of data, most notable are those from the NSWS
(Linthurst et al , 1986, Landers et al , 1987, Kaufmann et al, 1988), and limit the
conclusions that can be drawn concerning the spatial extent of aquatic effects attributable to
nitrogen deposition Nonetheless, the maps illustrate the existence of severe problems in the
Northeast (especially the Adirondack and Catskill Mountains) and the Southeast (in the
10-251
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o Data Indicate no Influence of NO"
• Data indicate strong influence of NO"
* Data Indicate strong Influence of NO*
and Increasing trend In NO3
Figure 10-36. Location of acid-sensitive lakes and streams in the northeastern United
States where the importance of nitrate to seasonal water chemistry can be
determined.
Source: Kahl et al (1991), Wigington et al (1990), Dnscoll et al (1987a), Dnscoll and Van Dreason
(in press), Kramer et al (1986), Murdoch and Stoddard (in press a), Eshleman and Hemond (1985),
Morgan and Good (1988), Baird et al (1987), Likens (1985), Sharpe et al (1984), Stoddard and
Kellogg (in press), DeWalle et al (1988), Barker and Witt (1990), Schofield et al (1985), Phillips and
Stewart (1990)
Mid-Appalachians and Great Smoky Mountains), and the potential for future problems in the
West.
It is also possible to draw correlations between rates of nitrogen deposition and rates of
nitrogen loss from watersheds, although these analyses cannot indicate causal relationships,
they can suggest patterns that merit further attention Two independent attempts have been
made to relate deposition and watershed nitrogen export in the United States, and both
suggest similar conclusions Kaufmann et al (1991) used data from the NSS (Kaufmann
10-252
-------
O Data indicate no influence of NO3
0 Data indicate strong Influence of NO 3
* Data indicate strong influence of NO ~
and increasing trend in NO3
N
Figure 10-37. Location of acid-sensitive lakes and streams in the southeastern United
States where the importance of nitrate ions to seasonal water chemistry
can be determined.
Source Elwood et al (1991), Cosby et al (1991), Elwood and Turner (1989), Buell and Peters (1988), Swank
and Waide (1988), Jones et al (1983), Silsbee and Larson (1982), Katz et al (1985), Weller et al
(1986), Wigington et al (1990), Kramer et al (1986), Edwards and Helvey (1991)
et al, 1988) and interpolated wet deposition values (of NO3" + NH4+) to correlate
deposition and surface water dissolved inorganic nitrogen concentrations (NO3~ + NH4+) in
large physiographic regions of the eastern United States (Figure 10-39) The NSS was a
probability-based sample of streams, sampled at spring base flow in 1987, because it is
probability-based, the results from the relatively small number of streams sampled in the
NSS can be extrapolated to the population of streams within each of the nine regions
sampled The results of the correlation suggest a strong correspondence between median wet
deposition of nitrogen in a region and the median spring base-flow concentration of nitrogen
10-253
-------
Data indicate no influence of NOj
Data indicate strong influence of NOg
Data indicate strong influence of NO3
and increasing trend in NOj
N
Figure 10-38. Location of acid-sensitive lakes and streams in the western United States
where the importance of nitrate ions to seasonal water chemistry can be
determined.
Source Melack and Stoddard (1991), Stoddard (1987a), Loranger et al (1986), Wigington et al (1990),
Kramer et al (1986), Welch et al (1986), Eilers et al (1990), Gilbert et al (1989)
ui a region In addition, the results suggest a threshold rate of wet nitrogen deposition of
approximately 3 kg nitrogen/ha/year, above which significant losses of nitrogen from
watersheds can begin to occur
Driscoll et al. (1989a) collected input/output budget data for a large number of
undisturbed forested watersheds in the United States and Canada, and summarized the
10-254
-------
relationship between nitrogen export (of NO3") and wet nitrogen deposition
(of NO3" + NH4+) These data are supplemented in Figure 10-39 with some published
input/output data that were not mcluded in the original figure Dnscoll et al (1989a) stress
that the data were collected using widely differing methods and over various tune scales
(from 1 year to several decades) Like the data of Kaufmann et al (1991, Figure 10-39),
these budget data suggest a threshold rate of wet nitrogen deposition of approximately 3 kg
nitrogen/ha/year, above which significant export of NO3" from watersheds may occur
10.9.9.2 Eutrophication
Assigning responsibility for the eutrophication of lakes and estuaries to NOX requires a
determination of two key conditions The first is that the productivity of the aquatic system
be limited by the availability of nitrogen, rather than by some other nutrient or physical
factor The second is that nitrogen deposition be a significant source of nitrogen to the
system In many cases of eutrophication, the supply oi nitrogen from deposition is minor
when compared to other anthropogenic sources, such as pollution from either point or
nonpoint sources
It is generally accepted that the productivity of fresh waters is limited by the availability
of phosphorus, rather than the availability of nitrogen (reviewed by Hecky and Kilham,
1988) Conditions of nitrogen limitation do occur in lakes, but are often either transitory, or
the result of high inputs of phosphorus from anthropogenic sources Often when nitrogen
limitation does occur, it is a short-term phenomenon because nitrogen-deficient conditions
favor the growth of nitrogen-fixing blue-green algae (e g , Smith, 1982) Because nitrogen-
fixing species are not limited by the availability of fixed nitrogen (e g , NH4+ or NO3~), they
may thrive under conditions where other species are nitrogen limited, and may effectively
mcrease rates of nitrogen input to the system (by fixation of gaseous nitrogen) beyond the
levels where system productivity can be said to be nitrogen limited It appears that nitrogen
limitation may occur naturally (i e , in the absence of anthropogenic phosphorus inputs) in
lakes with very low concentrations of both nitrogen and phosphorus, as are common in the
western United States and in the Northeast Suttle and Harrison (1988) and Stockner and
Shortreed (1988) suggest that phosphorus concentrations are too low in these systems to
allow blue-green algae to thrive, because they are poor competitors for phosphorus at very
10-255
-------
50-
40-
30-
10-
0-
w
(a)
NSS-I Subrogions
100
200
300 400
500
Wet NO3 + NH, Deposition (eq/ha/year)
400
?
1
o
ŁE
0
it
1
g
(b)
350-
300-
250-
200-
150-
100-
50 .
0 -
« 0
0 0
o «
o °
o
o
o
o
° 0 °0 °
o o
, n n o ^ ° ° o o § °0
0 100 200 300 400 500 600
Rate of Nitrogen Wet Deposition (eq/ha/year)
Figure 10-39. (a) Relationship between median wet deposition of nitrogen (nitrate ions
plus ammonium ions) and median surface water nitrogen (nitrate ions
plus ammonium ions) concentrations for physiographic districts within
the National Stream Survey that have minimal agricultural activity.
[Subregions are Poconos/Catskills (ID), Southern Blue Ridge Province
(2As), Valley and Ridge Province (2Bn), Northern Appalachians (2Cn),
Ozarks/Ouachitas (2D), Southern Appalachians (2X), Piedmont (3A),
Mid-Atlantic Coastal Plain (3B), and Florida (3C)]. From Kaufmann
et al. (1991). (b) Relationship between wet deposition of nitrogen (nitrate
ions plus ammonium ions) and rate of nitrogen export for watershed
studies throughout North America. Sites with significant internal sources
of nitrogen (e.g., from alder trees) have been excluded.
Source Dnscoll et al (1989a), additional data from Barker and Witt (1990), Edwards and Helvey (1991),
Kelly and Meagher (1986), Katz et al (1985), Buell and Peters (1988), Weller et al (1986), Owens
et al (1989), Feller (1987), Stoddard and Murdoch (1991)
10-256
-------
low concentrations Results of the NSWS (Kanciruk et al , 1986, Eilers et al , 1987) suggest
that the largest number of potentially nitrogen-limited lakes in the United States occur in the
West (20 to 30% of the population of lakes sampled by NSWS), and particularly in the
Pacific Northwest, although significant numbers may also occur in the Upper Midwest (15 to
25 % of population) In all cases, because the concentrations of both nitrogen and
phosphorus are low, additional inputs of nitrogen may have a limited potential to cause
eutrophication because their input will quickly lead to a switch in the limiting nutrient,
additions of nitrogen to these systems would soon lead to nitrogen-sufficient and
phosphorus-deficient conditions Increases in nitrogen deposition to some regions would
probably lead to measurable increases in algal biomass in lakes with both low concentrations
of dissolved nitrogen and substantial concentrations of phosphorus, but the number of lakes
that meet these criteria naturally (i e , that do not have large anthropogenic inputs of
phosphorus) is likely to be quite small
Few topics in aquatic biology have received as much attention in the past decade as the
debate over whether estuanne and coastal ecosystems are limited by nitrogen, phosphorus, or
some other factor (reviewed by Hecky and Kilham, 1988) Numerous geochemical and
experimental studies have suggested that nitrogen limitation is much more common in
estuanne and coastal waters than in freshwater systems Experiments to confirm widespread
nitrogen limitation in estuaries have not been conducted, however, and nitrogen limitation
cannot be assumed to be the rule Taken as a whole, the productivity of estuanne waters of
the United States correlates more closely with supply rales of nitrogen than of other nutnents
(Nixon and Pilson, 1983) Specific instances of phosphorus limitation (Smith, 1984) and of
seasonal switching between nitrogen and phosphorus limitation (D'Eka et al , 1986, McComb
et al , 1981) have been observed and stand as exceptions to the general rule of nitrogen
limitation in marine ecosystems Nitrogen-fixing blue-green algae are rarely abundant in
estuanne waters (Howarth et al, 1988a), and so nitrogen-deficient conditions may continue
indefinitely in these systems, unless nitrogen supply exceeds the biological demand for
nitrogen
Estimation of the contnbution of nitrogen deposition to the eutrophication of estuanne
and coastal waters is made difficult by the multiple direct anthropogenic sources (e g , from
agriculture and sewage) of nitrogen against which the importance of atmosphenc sources
10-257
-------
must be weighed Estuaries and coastal areas are natural locations for cities and ports, and
most of the watersheds of major estuaries in the United States have been substantially
developed. The crux of any assessment of the importance of nitrogen deposition to estuarme
eutrophication is establishing the relative importance of direct anthropogenic effects (e g ,
sewage and agricultural runoff) and indirect effects (e g , atmospheric deposition) In the
United States, a large effort has been made to establish the relative importance of sources of
nitrogen to the Chesapeake Bay (e g , D'Eha et al, 1982, Smullen et al, 1982, Fisher
et al, 1988b, Tyler, 1988) Estimates of the contribution of nitrogen to the Chesapeake Bay
from each individual source are very uncertain, estimating the proportion of nitrogen
deposition exported from forested watersheds is especially problematic, but critical to the
analysis because about 80% of the Chesapeake Bay basin is forested Nonetheless, three
attempts at determining the proportion of the total NO3" load to the bay attributable to
nitrogen deposition all produced estimates in the range of 18 to 31 % (Table 10-27) Supplies
of nitrogen from deposition exceed supplies from all other nonpoint sources to the bay (e g ,
agricultural runoff, pastureland runoff, urban runoff), and only point-source inputs represent
a greater input than deposition
10.9.9.3 Direct Toxicity
Toxic effects of nitrogen on aquatic biota result from un-iomzed NH3, which occurs in
equilibrium with ionized NH4+ and OH" Ammonia concentrations approach toxic
concentrations most commonly at high pH and temperature values, which are most typical of
heavily polluted lakes and streams (e g , Effler et al, 1990) In the well-oxygenated
conditions typical of unpolluted lakes and streams (as well as in most watersheds), NH4+ is
rapidly oxidized to NO3", which does not have toxic effects on aquatic organisms Within
the typical range of pH and temperature that unpolluted lakes and streams experience, toxic
concentrations of NH3 resulting from nitrogen deposition would be extremely unusual At a
pH of 7 and a temperature of 15 °C, for example, concentrations of total NEkj. 'would have
to reach over 750 /jmol/L before chronically toxic concentrations of free NH3 would
develop. Currently, no areas of North America are known to experience rates of nitrogen
deposition that are sufficient to produce such high concentrations of total NB^ in surface
waters.
10-258
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TABLE 10-27. THREE NITROGEN BUDGETS FOR THE CHESAPEAKE BAY
Source of Nitrogen
Direct Deposition
Nitrate Ions
Ammonium Ions
Nitrogen Load to Bay (from direct deposition)
Forests
Nitrate Ion Deposition
Ammonium Ion Deposition
Watershed Retention
In-Stream Retention
Atmospheric Nitrate Ion Load to Bay (from forests)
Nitrogen Load to Bay (from forests)
Pasture Land
Nitrate Ion Deposition
Ammonium Ion Deposition
Animal Wastes
Watershed Retention
In-Stream Retention
Atmospheric Nitrate Ion Load to Bay (from
pastures)
Nitrogen Load to Bay (from pastures)
Cropland
Nitrate Ion Deposition
Ammonium Ion Deposition
Fertilizers
Watershed Retention
In-Stream Retention
Atmospheric Nitrate Ion Load to Bay (from
cropland)
Nitrogen Load to Bay (from cropland)
Residential/Urban
Nitrate Ion Deposition
Ammonium Ion Deposition
Watershed Retention
In-Stream Retention
Atmospheric Nitrate Ion Load to Bay (from urban
areas)
Nitrogen Load to Bay (from urban areas)
Point Sources
NITRATE ION LOAD TO BAY (FROM
DEPOSITION)
TOTAL NITROGEN LOAD TO BAY1"
% of Nitrogen from NO3 deposition
EDF Budget Versar Budget
(108 kg/year) (108 kg/year)
08
04
1 3
90
49
08
14
24
1 3
145
07
1 5
25
1 4
158
08
59
04
03
03
04
3 4
35
1394
25%
07
a
07
84
80% -a 95%
50% 0 2 50%
02
17
95%° -a 94-99%
50%c 11 8 50%
001-
006
007-
04
}70% 2fl8
* -a 76-99%
4 1- 50%
270
001-
03
006-
3 6
07
35% -a 62-96%
0% 0 01- 20%
0 14
001-
0 14
20-3 2
094-
1 48
3 03-
826
18-
31%e
Refined Budget
(108 kg/year)
06
03
08
64
35
07
10
1 3
07
195
0 13
08
2 1
1 1
158
007
06
06
03
0 1
03
34
1 53
682
225%
846%
35%
95 %d
35%
95%
35%
50%
35%
aThe Versar Budget (Tyler, 1988) does not calculate loads of ammonium ions
bFor the Environmental Defense Fund (EDF) Budget (Fisher et al , 1988a, Fisher and Oppenheimer, 1991) and
refined budget, total nitrogen load to the bay includes both nitrate ions (NO3") and NH4 The Versar Budget
(Tyler, 1988) includes only NO3"
Watershed and m-stream retention values for pastureland in the EDF Budget apply only to animal wastes For
atmospheric deposition, the cropland retention value (70%) was used
95% retention was used for animal wastes, 85% retention was used for deposition (see text in
Section 10 8 4 3)
eThe range of contributions of NO3~ deposition to the total budget were calculated by comparing maximum-to-
maximum estimates and minimum-to-minimum estimates These combinations are more likely to occur during
extreme (e g , very wet or very dry) years
10-259
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Vose, J M., Swank, W T (1990) Preliminary estimates of foliar absorption of N-labeled nitric acid vapor
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10-308
-------
11. EFFECTS OF NITROGEN OXIDES
ON VISIBILITY
Clear days are an important aesthetic resource for us all They also carry
commercial value for tourism and real estate Tims, the appearance of layers of
smoggy haze over cities and across rural vistas is one of the most widely noticed
effects of air pollution (Sloane and White, 1986)
Emissions of nitrogen oxides (NOX) can contribute significantly to visibility impairment,
or the "layers of smoggy haze" noted by Sloane and White They can have aesthetic impact
because they can cause a yellow-brown discoloration of the atmosphere when present in
plumes or in urban, regional, and layered haze They can also reduce visual range, thereby
diminishing the contrast of distant objects viewed through an atmosphere containing NOX
Only some of the species in the NOX family, however, are optically active and thus able
to affect atmospheric visibility Figure 11-1 illustrates the major categories (including
atmospheric oxidation products) of NOX species and the two species that have an effect on
visibility nitrogen dioxide (NO2), a gas that absorbs light, chiefly at the blue end of the
visible spectrum, and nitrate aerosols, particles that scatter light The other forms of NOX
that occur in ambient air, nitric oxide (NO), nitrous acid (HONO), and nitric acid (HNO^),
are optically inactive gases and therefore do not contribute to visibility impairment
(Peroxyacetyl nitrate [PAN], HONO, and HNO3, however, interfere with chemiluminescence
NO2 measurements and therefore would indirectly affect the estimation of the effects of NO2
on visibility ) Thus, depending on the form in which NOX exists in the atmosphere, NOX
may or may not play a significant role overall in visibility For example, nitrate aerosol may
never form from HNO3 in certain warm climates, in aieas with low ambient atmospheric
concentrations of ammonia (NH3), or in areas with high ambient concentrations of acid
sulfate, since acid sulfate reacts with ammonium nitrate (NH4NO3), thereby releasing nitric
acid
Nitrogen oxides have been found to play a significant role in the aesthetic impact
caused by combustion emission sources such as power plants This impact is dominated by
the yellow-brown coloration caused by NO2 relatively near the source (within 100 km)
Nitrate aerosols have been found to play a significant role in the haze observed in urban
11-1
-------
8
ill
flC
6
«
O
Z
o
'So:
GO
a
o ^
^J **iy
IP
o-
ILJ
O
§
itr
o
CO
UJ
o
X
o
o
a:
11-2
-------
areas in the western United States, particularly during winter and near significant ammonia
sources (such as cattle feedlots) Nitrate aerosols, along with sulfate, may also play a
significant role in the formation of wintertime layered haze that has been observed in the
vicinity of large, isolated power plants
Although NOX has a clearly defined effect on visibility (aesthetic impacts and visual
range reduction), in most areas of the country visibility impairment is usually dominated by
other species, such as sulfate and elemental and organic carbon particles Also, it should be
noted that brownish atmospheric discoloration may be caused by particles such as sulfate and
not solely by NO2 and nitrate
11.1 OVERVIEW OF LIGHT SCATTERING AND ABSORPTION
The visibility effects of the optically active forms of NOX, NO2 and nitrate aerosols,
can best be illustrated by reviewing some of the fundamentals of atmospheric optics The
deterioration of visibility is the result of the absorption and scattering of light by gaseous
molecules and suspended solid or liquid particles (Middleton, 1952) Absorbed light is
transformed into other forms of energy, such as heat, whereas scattered light is reradiated in
all directions
The effect of the intervening atmosphere on the visibility and coloration of a viewed
object, such as the horizon sky, a distant mountain, 01 a cloud, can be calculated by solving
the radiative transfer equation along the hue of sight (see schematic in Figure 11-2) This
equation can be solved if the light extinction properties of the intervening atmosphere are
known
The change in the light intensity of a specific wavelength, or spectral radiance I(X), as a
function of distance along the line of sight can be calculated as follows (Chandrasekhar,
1960, Latimer and Samuelsen, 1975, 1978, Latimer et al, 1978, White et al, 1986)
J(X,e)bscat(X),
11-3
-------
SCATTERING
ANGLE, 0
dr
ELEMENTAL VOLUME
(CONTAINING AIR,
PARTICLES, AND NO2)
_INE OF SIGHT
I + dl
OBJECT
OBSERVER
Figure 11-2. Schematic of an elemental volume of haze along a line of sight.
Source Latimer and Ireson (1980)
where
I(X) = the spectral light intensity of wavelength X,
r = the distance along the hue of sight from the object to the observer (see
Figure 11-2 for definitions),
J(X,0) = source function,
bscat(X) = the light scattering coefficient, and
bext(X) = the light extinction coefficient, the sum of scattering and absorption
An examination of Equation 11-1 indicates that light can be both removed and added to
the line of sight. The first term on the right side of this equation represents the rate at which
light is removed from the line of sight and the second term is the rate at which it is added
If the first term is larger than the second, the net effect is a decrease in light intensity
11-4
-------
(darkening) of an observed object as one moves along the line of sight (see upper curve in
Figure 11-3) If the second term is larger than the first, the net effect is an increase in light
intensity (brightening) of an observed object The darkening effect, the first term, is
dependent on total light extinction (bext), which is the sum of light scattering and absorption
The brightening effect, the second term, is dependent only on light scattering (bscat) Thus,
light absorption can only darken objects viewed through the atmosphere, whereas light
scattering can either brighten or darken viewed objects Since NO2 is a gas that
preferentially absorbs blue hght, it always tends to darken and discolor the sky and objects
viewed through the atmosphere Because nitrate aerosol scatters light, it can either brighten
or darken the sky and objects
BRIGHT OBJECT
LIGHT INTENSITY OF HORIZON
BLACK OBJECT
0 OBJECT-OBSERVER DISTANCE (r0) fv
Figure 11-3. Effect of a homogeneous atmosphere on light intensity of bright and dark
objects as a function of distance along a line of sight.
Source Latimer and Ireson (1980), adapted from Middleton (1952)
11-5
-------
The hght extinction (bext) coefficient is the optical equivalent of ambient pollutant
concentration This parameter (as well as its scattering and absorption components) has units
of inverse distance (e g , m" , km" , Mm") These coefficients can be considered to be the
equivalent hght extinction, scattering, or absorption cross-sectional area (m ) per unit volume
of ambient air (m3) In Equation 11-1, the light extinction coefficient is the sum of its hght
scattering and light absorption components
) + babs(X) = (bsg + bsp) + (bag + bap) (11-2)
The first term, bsg, is the scattering coefficient attributable to gases and is the result
primarily of Rayleigh scattering caused by gases in the atmosphere (chiefly nitrogen and
oxygen). The second term, bsp, is the scattering coefficient from particles suspended in the
atmosphere (aerosols) Nitrate aerosol contributes to this term, along with other aerosols,
including sulfates, organic and elemental carbon, and other particulate matter, both fine
(<2.5 jim in diameter) and coarse (>2 5 /xm in diameter) The third term, bag, is the
absorption coefficient resulting from gases Nitrogen dioxide is the only significant
contributor to this term m the visible spectrum The fourth and last term, bap, is the
absorption coefficient resulting from particles This term is dominated by the effect of
elemental carbon (soot), a combustion product found, for example, in diesel engine exhaust
Except in very clean areas of the western United States, natural bsg is a small fraction
of bext, b<,p usually dominates bext, and fine-particle bsp usually dominates total bsp (White,
1990).
All of these components of total hght extinction, as well as total extinction itself, are
functions of the wavelength of hght As discussed in more detail later, the atmospheric
discoloration caused by NOX (both NO2 and nitrate aerosol) can be explained by the
wavelength-dependent nature of NO2 light absorption and nitrate hght scattering effects
Both scattering and absorption from these NOX species are stronger at the blue end of the
visible spectrum (wavelength X = 0 4 /*m) than at the red end (X = 0 7 /*m)
The scattering or absorption coefficient can be determined from the product of the
concentration of an optically active species and its hght scattering or specific absorption
*y
efficiency OS) This efficiency is commonly stated m units of m /g When the ambient
11-6
-------
concentration (jttg/m3) of a given species is multiplied by its extinction efficiency (m2/g), the
extinction coefficient of that species, in units of inverse megameters (Mm" ), is obtained
The light extinction efficiency for particles is a strong function of particle size (see
Figure 11-4) Fine particles, those with diameters <2 5 jiim, are much more effective per
unit mass in scattering light than are coarse particles, those with diameters > 2 5 pm
Particle scattering efficiency is a maximum for particles having a diameter of approximately
0 5 jum Coarse particles have scattering efficiencies that are approximately an order of
magnitude smaller (see Figure 11-4)
100
10
<
CO.
01
TYPICAL
NONABSORBING
AEROSOL
001
01
1 0
10.0
PARTICLE DIAMETER Gum)
Figure 11-4. Light extinction efficiency at X = 0.55 pun as a function of particle size for
soot and for typical, nonabsorbing atmospheric aerosol.
Source Latimer (1988a) after Bergstrom (1973)
11-7
-------
Nitrate particles can be either coarse or fine Milford and Davidson (1987) reviewed
the sizes of participate sulfate and nitrate in the atmosphere, nitrate mass median diameters
ranged from 0 23 to 4 2 /tm in 16 different measurement sets Wolff (1984) noted that in
continental environments nitrate can exist as either coarse or fine particles, however, in a
number of summertime studies in the eastern United States, nitrate concentrations were quite
low and nitrate occurred primarily in the coarse mode (Wolff, 1984, Mamane and Dzubay,
1986). Wolff explained this qualitatively by the reaction of alkaline soil dust with HNO3,
nitrate aerosol is not formed in the submicron mode if temperatures are high or if NH3 is not
available or is tied up with sulfate It should be noted, however, that the data of Wolff
(1984) were collected using methods later found to have significant artifact problems
In coastal environments, nitrate may also be primarily in the coarse mode because of reaction
with sea salt (Yoshizumi, 1986, Wall et al, 1988, Orel and Seinfeld, 1977, Mamane and
Mehler, 1987). Richards (1983) suggested that coarse-particle nitrate may form from
nighttime oxidation involving nitrogen pentoxide-water reactions on the surfaces of particles
Nitrate is in the submicron fine mode when it reacts directly with NH3 to form NH4NO3
(Orel and Seinfeld, 1977; Wolff, 1984) The submicron nitrate forms when conditions are
favorable (abundant ambient NH3 and moderate temperatures)
Nitrate aerosol in the size range of 0 1 to 2 5 /tm is most effective per unit mass in
scattering light. For particles having a typical density (p) of 2 g/cm3 and a diameter of
0 5 /tm, Figure 11-4 shows that the scattering efficiency at the middle of the visible spectrum
2
(X = 0.55 /an) is approximately 5 m /g By contrast, the average NO2 absorption efficiency
r\
over the wavelengths 0 45 to 0 65 /*m, centered on 0 55 /m, is 0 144 m /g (Latimer and
Ireson, 1988, based on Dixon, 1940) Thus, the extinction efficiency of nitrate aerosol can
be more than an order of magnitude greater than that for NO2 As discussed in the next
section, the extinction efficiencies of both nitrate aerosol and NO2 gas are strong functions of
the wavelength, being larger at the blue end (X = 0 4 /*m) of the visible spectrum
11-8
-------
11.2 ATMOSPHERIC DISCOLORATION CAUSED BY NITROGEN
OXIDES
As Finlan (1981) so aptly stated "Many of the most beautiful sights in nature are
caused by wavelength-dependent light scattering It can be truly exhilarating to see the
beauty of the blue sky or to witness a rainbow, a sunset, or a sunrise Unfortunately, the
physical processes responsible for these beautiful sights also cause much of the color that we
often see in smogs and hazes over cities "
The undesirable yellow or whisky-brown color oi hazes has been an ongoing topic of
discussion in the literature for more than 20 years Hodkuison (1966) described the effects
that NO2 could produce on the color of the atmosphere Charlson and Ahlquist (1969),
however, argued that wavelength-dependent scattering was the primary cause of atmospheric
discoloration in most situations Horvath (1971) countered with the argument that any color
caused by wavelength-dependent light scattering that removed light from the line of sight
would be offset by the additional light scattered into the line of sight by the same
wavelength-dependent scattering Thus, he thought that any color would be the result of the
absorption of blue light by NO2 He did conclude, however, that if extremely bright objects
were viewed through an aerosol, a discoloration could result Charlson et al (1972)
measured NO2 concentrations and the wavelength dependence of the light-scattering
coefficient in Pasadena, CA, during August and September 1970 and concluded that NO2 had
a significant effect on atmospheric color 20% of the tune Sloane (1987) applied Mie theory
to calculate the effects of urban haze mixtures of NO2 and elemental carbon (soot) She
found that soot can offset the coloration caused by NO2, even though both species absorb
preferentially at the blue end of the spectrum Husar and White (1976) performed careful
atmospheric optics calculations using Mie scattering theory (Kerker, 1969) to assess the
relative roles of wavelength-dependent light scattering by particles and wavelength-dependent
light absorption caused by NO2 They found that particles typical of Los Angeles haze could
cause yellow-brown discoloration when the sun was behind the observer (scattering angle
0 > 90°), and typical NO2 concentrations could perceptibly add to this color More
detailed analysis by Finlan (1981) confirmed the importance of scattering angle and the size
distribution and refractive index of the aerosol in determining atmospheric color
11-9
-------
Atmospheric color can be studied theoretically by solving Equation 11-1 for the spectral
radiance or light intensity of an object observed at distance r as follows (Middleton, 1952,
Latimer and Samuelsen, 1975, 1978, Latuner et al, 1978, Husar and White, 1976, White
et al., 1986)-
Ir = Io exp(-r) + J [1 - exp(-T)], (11-3)
where
Ir, IQ = spectral light intensities at distance r from an object and at the object itself,
T = optical depth between the object and the observer (= J bext dr),
J = the source function (the second term in Equation 11-1, divided by bext)
Equation 11-3 can be used to evaluate the effect of a uniform concentration of NO2 on
atmospheric coloration The ratio of the intensity of the horizon sky (h) with and without a
given concentration of NO2 can be calculated from Equation 11-3 as follows (Hodkinson,
1966, Robinson, 1968; White, 1982)
The light absorption coefficient for NO2, bag, is a strong function of wavelength
Figure 11-5 shows the wavelength dependence of the NO2 light absorption efficiency over
the ultraviolet and visible spectrum (Davidson et al , 1988) The light efficiency, a, is the
ratio of the light absorption coefficient to the NO2 concentration The value at the blue end
of the visible spectrum, X = 0 4 /jm, is 5 9 x 10~19 cm2 molecule" or 1 45 km" ppm" , is
nearly six tunes larger than the value at the center of the visible spectrum at a green
wavelength X = 0.55 /tm, which is 1 0 x IO"19 cm2 molecule"1 (or 0 24 km"1 ppm"1) This
value at X = 0 55 ftm of 0 24 km" ppm" is considerably less than the value of 0 33 km"
ppm" derived from earlier measurements (Dixon, 1940) When Equation 11-4 is evaluated
as a function of wavelength (X), and the X-dependence of bscat is neglected, the curves shown
11-10
-------
I
UJ
o
6 -
2 -
1 -
250
350 450 550
WAVELENGTH, A. (nm)
650
Figure 11-5. Light absorption efficiency of nitrogen dioxide estimated for —30.2 °C
(thin line) and 124 °C (dark line). (To obtain units of ppm"1 km"1,
multiply cm2 molecule by 2.46 x 1C18.)
Source Davidson et al (1988)
m Figure 11-6 are obtained for the horizon-sky light-intensity ratio (Hodkinson, 1966, White,
1982) Nitrogen dioxide causes a darkening effect, especially at the blue end of the visible
spectrum For example, with an NO2-visual range product of 0 3 ppm-km, the horizon sky
light intensity at X = 0 4 /on is about 14% less than it would be without NO2 and would
thus be quite noticeably discolored (yellow or brown) This concentration-visual range
product could be caused by 0 03 ppm (60 jwg/m3) NO2 associated with a visual range of
10 km, which is typical of urban haze (Note 0 03 ppm x 10 km = 0 3 ppm-km)
Atmospheric aerosols, including particulate nitrates, can also cause atmospheric
discoloration (Ahlquist and Charlson, 1969, Husar and White, 1976) The scattering
coefficient of particles smaller than 1 5 /-cm in diameter can be strongly dependent on the
11-11
-------
CM
o
1
CO
o
A
31
CD
1 0
09
08
07
06
05
04
03
02
01
00
3 ppm-km
I
04
BLUE
I
05 06
WAVELENGTH (urn)
07
RED
Figure 11-6. Effect of nitrogen dioxide on horizon sky brightness as a function of the
wavelength of light; relative horizon brightness, bscat/(bscat + bag) for
selected values of the product of nitrogen dioxide concentration and visual
range assuming that b^^ = 3/(visual range).
Source White (1982) adapted from Hodkinson (1966)
wavelength of light, as shown in Table 11-1 (Latimer and Ireson, 1980) For example, an
aerosol with a mass median diameter of 0 5 /*m has a light scattering coefficient bscat that is
inversely proportional to wavelength X Thus, light scattering at the blue end (X = 0 4
of the visible spectrum would be 75 % greater (7/4 = 1 75) than at the red end
(X = 0.7 fjtm). Because the light-scattering coefficient caused by aerosols and the
light-absorption coefficient caused by NO2 are both wavelength-dependent, both can cause
atmospheric discoloration
11-12
-------
TABLE 11-1. WAVELENGTH DEPENDENCE OF LIGHT SCATTERING
COEFFICIENT AS A FUNCTION OF PARTICLE
LOGNORMAL SIZE DISTRIBUTION
Mass Median
Diameter (DG)a
01
02
03
04
05
06
08
10
>5
b
Oi
28
2 1
16
1 2
1 0
07
05
02
0
aGeometnc standard deviation ag = 2
a is defined as follows
= bscat(X2)
(appropriate for 0 4 < X < 0 7 /urn)
Source Latimer and Ireson (1980)
Husar and White (1976) formulated the problem of atmospheric coloration rigorously in
terms of radiative transfer theory A solution was derived from theory and from aerosol size
distributions measured in Los Angeles They found that aerosol (without NO2) could cause
yellow-brown discoloration, and that this discoloration would increase as NO2 concentrations
increase and as the scattering angle, 9, increases Noticeable discoloration from NO2 was
found to occur at concentrations as low as 0 05 ppm The discoloration effect caused by
particles, unlike that caused by NO2, is dependent on the scattering angle, 6, with most
intense effects occurring in situations in which the sun is behind the observer (9 > 90°)
In addition, when the viewed object has a light intensity greater than the horizon-sky light
intensity (the Ih asymptote in Figure 11-3), light scattered by fine particles would cause a
darkening and discoloring effect because of the wavelength-dependent light scattering
Waggoner et al (1983) used teleradiometer measurements to determine the color of the
winter haze in Denver that is commonly known as the "brown cloud " Although this haze
11-13
-------
appeared to be brown in contrast to the blue sky above, they found that its spectral
light-intensity distribution was gray and was caused primarily by aerosol rather than NO2
These findings were consistent with the conclusions of Horvath (1971) and of Husar and
White (1976) that yellow haze could appear brown if it were darker than the viewing
background. The chromatic adaptation of the human eye-brain system (Cornsweet, 1970)
also explains why a gray haze may appear yellow or brown An observer that has adapted to
the color of the blue sky will visually perceive a gray haze as the complementary color to
that adaptation (i e , yellow or brown)
11.3 VISUAL RANGE REDUCTION CAUSED BY NITROGEN OXIDES
At some distance from a black object, an observer can no longer distinguish between
the intensity of it and the sky This limit of perceptibility is defined by a threshold (hminal)
contrast that is just noticeable to a human observer The distance at which the contrast of a
black object against the horizon sky equals this threshold is called the visual range or,
commonly, visibility Although a range of values for the threshold contrast from about 1 to
20% is supported by the literature (Middleton, 1952, U S Environmental Protection Agency,
1979, Latimer, 1988b, Gnffing, 1980, Dzubay et al , 1982), the threshold human visual
perception threshold is commonly assumed to be a contrast of 2 %
Koschmieder (1924) developed a formula for visual range, which is based on the
assumptions that the threshold contrast is 2 % , that the atmosphere is uniform and cloud-free,
and that the curvature of the Earth can be ignored when evaluating horizon light intensity
The Koschmieder equation is simply
rv = -In (Cmm)/bext, (11-5)
where
rv = the visual range,
Cmin = the contrast perceptibility threshold, and
~ tne hgnt extinction coefficient, as
defined previously
11-14
-------
If the commonly accepted threshold of 2 % is used above, the Koschmieder equation becomes
rv = 3 9/bert, (11-6)
the most common form of the equation If the perceptibility threshold is assumed to be 5 %,
which appears to correlate best with common airport visibility measurements (Samuels, 1973,
Johnson, 1981, Latimer, 1988b), the equation becomes
rv = 3/bext (11-7)
Note that as the light extinction coefficient increases, visual range decreases This
inverse relationship suggests that increases in atmospheric concentrations of light scattering
and absorbing species will cause a decrease in visibility Figure 11-7 illustrates this
2
relationship for fine particles assumed to have a scattering efficiency of 4 m /g (U S
Environmental Protection Agency, 1979) Because both of the optically active NOX species,
NO2 and nitrate aerosol, contribute to the absorption and scattering components of light
extinction (bext), they both tend to reduce visual range
If it is not uniformly distributed m the atmosphere, NO2 may not contnbute to a
reduction in the contrast of a distant object and hence to visual-range reduction This can
happen when NO2 is located relatively close to the observer (e g , in a plume or haze layer)
In such a situation, the light absorbed by NO2 reduces the light intensity of both the sky and
the dark object equally, so that the sky and object are darkened but their contrast remains
unaffected Latuner and Samuelsen (1975, 1978) developed a formula to account for this
effect for atmospheres containing NO2 plumes
11.4 NITRATE PHASE CHANGES AND HYGROSCOPICITY
Assessment of the role played by nitrate particles m urban, regional, and layered haze
and in plumes is more difficult than for sulfates since certain of the nitrate aerosols (e g ,
NH4NO3) can volatilize during sample collection because of their volatile nature Unlike
sulfate, which is always in the particulate phase, nitrate often remains in the gas
11-15
-------
400 _
ADDITION OF 1 //g/m3
OF FINE PARTICLES
300 -1
HI
CD
CO
200 -
100
I I I I I I
V
I I I I
0.0
10 20 30
FINE PARTICLE CONCENTRATION
40
50
Figure 11-7. Effect on visual range of incrementally adding 1 /ig/m of fine particles
having a light extinction efficiency of 4 m /g. (Greater light extinction
efficiencies and visibility reduction than shown here would occur with
sulfate and nitrate aerosols at high relative humidities. See text.)
Source U S Environmental Protection Agency (1979)
phase as HNO3. In order for condensation of particulate nitrate (NEyNC^) to occur, there
must be sufficient atmospheric NH3 to react with HNO3 Furthermore, the vapor pressure of
NH4NO3 is strongly temperature-dependent, so that even if NH3 is present in the atmosphere
nitrate particles may not condense because of moderate or high temperatures The volatility
of particulate NH4NO3 contributes to the difficulty and uncertainties in most measurement
programs carried out to date These difficulties regarding phase changes are complicated
even more by the fact that NH4NO3 is deliquescent, it absorbs water from the atmosphere at
11-16
-------
moderate to high relative humidities Thus, like sulfate, the scattering efficiency of NH4NO3
is enhanced by associated liquid water in the particle droplet
The issue of changes in phase between gas and aerosol is a key uncertainty in
understanding, measuring, and mathematically modeling the impacts of nitrate aerosol
(Sloane and White, 1986)
Just as a cloud produces a dramatic visual effect when only a small fraction of the
water vapor changes phase, a substantial haze results if only a fraction of the gaseous
pollutant mass enters a condensed phase In this regard, visibility is unique among air
pollution effects, it depends not only on the amount of air pollution but in addition on
its phase This peculiarity greatly complicates the prediction of visibility impairment
and aerosol measurement procedures because the equilibrium between the condensed
and gaseous phases can be fragile
Ammonium nitrate particles will form only if (1) sufficient ambient NH3 is present to
neutralize any acidic sulfates and gas-phase HNO3 and (2) temperatures and relative
humidities are such that the thermodynamic equilibrium favors the formation of nitrate
aerosol (Stelson et al, 1979, Stelson and Seinfeld, 1982, Saxena et al, 1986, Sloane and
White, 1986) Until acidic sulfate compounds are fully neutralized as ammonium sulfate
((NH4)2SO4), they react with NH4NO3, releasing HNO3 vapor (Saxena et al, 1986)
If sufficient gas-phase NH3 is left after sulfate neutralization and temperatures are low
enough, NH4NO3 aerosol will condense At relative humidities above 62%, the deliquescent
point for NH4NO3, water vapor is taken up in the nitrate particle (droplet), forming a water
solution (Saxena et al, 1986) At these higher relative humidities, a new equilibrium is
established favoring more nitrate in the particulate phase (Sloane and White, 1986)
The net result of all of the nitrate phase interactions is that particulate NH4NO3 "can
build up only in locations where sufficient ammonia is present to neutralize the sulfunc acid.
This occurs, for example, in Los Angeles and Denver, where sulfate concentrations are
relatively low compared to concentrations of ammonia" (Milford and Davidson, 1987)
White and Macias (1987) attribute the extremely low nitrate aerosol concentrations observed
in the intermountain West to very low ambient HNO3 and NH3 concentrations and to the
warm temperatures during the nonwinter months Thus, the conditions can be summarized
under which fine nitrate particles are most likely to form high ambient concentrations of
NH3 and HNO3 (e g , Los Angeles, Denver), low ambient concentrations of sulfate (e g ,
most of the western United States), low temperatures (eg, winter), and high humidities
11-17
-------
(e.g., winter, coastal sites) Conversely, fine nitrate particles are least likely to form under
the following conditions low ambient concentrations of NH3 and HNO3 (e g , intermountam
West), high ambient concentrations of sulfate (e g , the eastern United States), high
temperatures (e g , summer), and low relative humidities (e g , the Southwest)
Furthermore, if sufficient coarse particles exist that can react with HNO3 (e g , sea salt,
alkaline soil dust), coarse nitrate particle formation is favored As subsequent discussion
bears out, these generalizations based on thermodynarnic equilibrium explain much of
observed nitrate aerosol behavior
The volatility of particulate nitrate makes its measurement difficult and uncertain
(Sloane and White, 1986) Significant positive and negative artifacts can occur with different
measurement techniques using different filter media (see Section 6 1) Thus, in evaluating
empirical studies of the importance of nitrate to total light extinction, it is important to
consider the complications caused by uncertainty in nitrate particle measurements
Further complicating the definition of the role of nitrate is the fact that nitrate particles
will absorb water vapor, becoming water solutions, at high humidities (above 62%) The
water associated with the nitrate results in scattering efficiencies per unit mass of nitrate that
are much larger than dry particle efficiencies The effect on light-scattering efficiencies of
liquid water associated with aerosols has been known for a long tune, but the specific effect
of associated water is difficult to quantify Empirical studies have used a nonlinear relative
humidity term to attempt to account for this effect
Tang and coworkers (Tang et al , 1981, Tang, 1982) developed a computer model for
calculating the optical properties of nitrate particles, both alone and in combination with
sulfate, as a function of particle size and relative humidity This model was based on
multicomponent aerosol thermodynamic theory as a function of particle chemical composition
and relative humidity Light-scattering efficiencies were calculated from resulting particle
sizes using Mie scattering theory Figures 11-8 through 11-12 summarize the light-extinction
o
coefficients for 1 jtg/m of sulfate or nitrate aerosol, or both, as a function of humidity
Figure 11-8 shows that pure (NH^SC^ exhibits a deliquescent point at 80% relative
humidity. At humidities above 80%, water vapor condenses, thereby increasing the aerosol
particle size, volume, and light scattering At humidities below 80%, the extinction
f\
efficiencies range from 1 to 4 m /g of sulfate, whereas above 80% humidity, extinction
11-18
-------
•5?
CVI
E,
O 100
z
LU
O
111
O
ts
10
1 5
2.0
25
1 01
50 60 70 80 90
RELATIVE HUMIDITY (%)
100
Figure 11-8. Light extinction efficiency for ammonium sulfate aerosol as a function of
relative humidity; with ammonium sulfate having lognormal particle size
distributions characterized by Dg = 0.2 /mi and ag = 1.01, 1.5, 2.0, and
2.5. (Multiply values by 1.375 to obtain efficiencies per unit mass of
sulfate anion.)
Source Modified after Tang et al (1981)
11-19
-------
•0°
22
2.0
LLJ
i 1.8
I 16
N
CO
§ "
Ł
Ł 1 2
1 0
- O
I
.__ THEORETICAL
~\ EXPERIMENTAL
0- ------ Q
20 30 40 50 60 70 80
RELATIVE HUMIDITY (%)
90 100
Figure 11-9. Particle size change for ammonium sulfate aerosols in a moist atmosphere
at 25° C.
Source Tang et al (1981)
rj
efficiencies can increase considerably above 10 m /g Figure 11-9 illustrates the hysteresis
effect, that is, the ability of the particle to hold on to liquid water, that can result when
relative humidity is slowly decreased Figure 11-10 shows the increase in light extinction of
pure NH4NO3 aerosol as a function of relative humidity At and above the deliquescent
point at 62% humidity, the scattering efficiency increases by a factor of two or more because
of the condensed water vapor associated with the nitrate particle Figures 11-11 and 11-12
show the effects of humidity on the light extinction efficiencies of different mixtures of
sulfate and nitrate aerosols Externally mixed aerosols, those in which the sulfate and nitrate
exist on different particles, exhibit the separate deliquescent points for (NH4)2SO4 (80 % RH)
and NH4NO3 (62% RH) Internally mixed aerosols, in which the sulfate and nitrate occur
11-20
-------
D)
CM
LU
O
U.
LJL
UJ
O
6
I
O
100
10
1.01
1.5
20
50 60 70 80 90
RELATIVE HUMIDITY (%)
100
Figure 11-10. Light extinction efficiency for ammonium nitrate aerosol as a function of
relative humidity; with ammonium nitrate aerosol having lognormal
particle size distribution characterized by D = 0.6 jim and
-------
100
•5?
UJ
g
u_
LL.
Ill
10
MOLAR RATIO S N = 3 1
EXTERNAL MIXTURE f S(0 2,1
I N(0 6,1 5
INTERNAL MIXTURE - (0 29,1 5)
5)
WHITE & ROBERTS (1977)
50 60 70 80 90
RELATIVE HUMIDITY (%)
100
Figure 11-11. Light scattering coefficient for 1 jig/m of a dry sulfate/nitrate aerosol
mixture as a function of relative humidity; bscat versus relative humidity
for externally and internally mixed sulf ate and nitrate aerosols (S:N =
3:1) for indicated size distributions (Dg, ag).
Source. Modified after Tang et al (1981), corrected by Tang (1982)
11-22
-------
100
•5?
CM
JE,
O
LU
g
u_
LL
HI
O
O
CD
10
MOLAR RATIO S N - 1 2
EXTERNAL MIXTURE f S(0 2,1 5)
1 N(0 6,1 5)
INTERNAL MIXTURE - (0 4,1 5)
*#>
*<&' "
*
0>
^
^
WHffE&BOBERTSj
50
60
70
80
90 100
RELATIVE HUMIDITY
Figure 11-12. Light extinction efficiency for 1 jig/m3 of a dry sulfate/nitrate aerosol
mixture as a function of relative humidity; b^ versus relative humidity
for externally and internally mixed sulfate and nitrate aerosols
(S:N = 1:2) for indicated size distributions (Dg,
-------
mixed within the same particle, do not exhibit distinct dehquescent points and have more
water associated with them at a given humidity, and hence have larger light-extinction
efficiencies The sulfate and nitrate aerosol mixtures may also exhibit hysteresis effects in
situations where humidity is reduced, thereby causing a haze to linger
11.5 MEASUREMENTS OF THE CONTRIBUTION OF NITROGEN
OXIDES TO URBAN AND REGIONAL HAZE
This section presents the various estimates of the contribution of NO2 and NH4NO3
aerosols to light extinction The discussion is broken unto two sections (1) recent state-of-
the-art measurements, and (2) earlier measurements having significant positive or negative
biases. As mentioned earlier in this chapter and also in Section 6 10 of this document,
earlier measurements of nitrate aerosol were plagued by significant positive and negative
artifacts. Glass filters had positive artifacts (i e , overestimated nitrate concentrations),
whereas Teflon® filters had negative artifacts (i e , underestimated nitrate concentrations)
The best measurements of nitrate are made with a denuder and nylon filter combination
There are relatively few studies with the state-of-the-art measurement technology, these
studies are discussed first For historical completeness, additional studies with significant
nitrate measurement artifacts are summarized next
11.5.1 Recent State-of-the-Art Measurements
Appel et al (1983, 1985) studied the chemical composition of aerosol in July and
August 1982 using state-of-the-art denuder difference measurements in three California cities
San Jose, Riverside, and Ix>s Angeles Mean nitrate anion concentrations were 4 4
(17% of the total fine particle mass of 22 3 jig/m3) in San Jose, 17 4 jwg/m3 (37% of the
total fine particle mass of 47 5 /*g/m3) in Riverside, and 10 2 /*g/m3 (17% of the total fine
particle mass of 61 5 jug/m3) in Los Angeles
Solomon et al (1992) have reported the results of a 1-year measurement program
conducted throughout the South Coast Air Basin in the greater ILos Angeles area during
1986, based on state-of-the-art denuder/nylon-filter measurements Most of the HNO3 in the
area was found in the aerosol phase, and a substantial fraction (about 42%) of the nitrate was
11-24
-------
coarse Fine-particle NH4NO3 concentrations ranged from 6 2 to 18 2 jttg/m3 and averaged
o
10 4 fjig/m for seven metropolitan area sites The background site had a fine-particle nitrate
<1
concentration of 1 1 /tg/m This is a substantial fraction of total fine-particulate mass in the
<5
Los Angeles area (23 1 to 42 1 jitg/m ) measured in 1982, reported by Gray et al (1986)
Lewis et al (1986) measured the chemical composition of fine and coarse fractions of
the aerosol for 20 days in January 1982 With their denuder/nylon-filter combination, they
measured a daytime fine-particle NH4NO3 mass of 3 4 jtcg/m , 18% of the daytime fine-
o
particle total mass of 19 0 jttg/m
Watson et al (1988) and Sloane et al (1991) measured the chemical composition of the
fine-particle mass, making 7-h daytime measurements (n = 24) during the winter of
1987-1988 The Micro-Orifice Uniform Deposit Impactor (MOUDI) was used to measure
fine-particle mass in several size ranges Nitrate measurements were considered accurate
based on a pnor comparison of impactor and denuder/nylon-filter measurements
3 ^
Ammonium nitrate averaged 3 4 jDtg/m , 21 % of the total fine-particle mass of 16 4 jtig/m
Particle size distributions were measured for two distinctly different days during this
measurement period a high-relative-humidity day with prolonged northeasterly flow and a
relatively low-humidity, stagnant day During the high-humidity day, the light-extinction
22 2
efficiency for nitrate anion was 7 2 m /g (6 6 m /g foi light scattering and 0 6 m /g for light
absorption) During the stagnant, lower-humidity day, the light-extinction efficiency for
00 0
nitrate anion was 3 6 m /g (3 0 m /g of scattering and 0 6 m /g for absorption)
Watson et al (1991) studied the chemical composition of the haze in Phoenix from
September 25, 1989, through January 22, 1990 The mean NH4NO3 and total fine-particle
mass concentrations over the entire time period and the four measurement sites were 4 4 and
22 3 jug/m3, respectively, for morning measurements and 4 8 and 15 5 jwg/m3, respectively,
for afternoon measurements Thus, nitrate contributed 19% of the fine-particle mass in the
morning and 31 % of the fine-particle mass in the afternoon The light-scattering efficiency
of nitrate anion was fit with the following equation 2 3 + [i 7/(l — /*)] m2/g, where ju. is
relative humidity defined as percentage divided by 100 Thus, for 50% RH, the nitrate anion
o
scattering efficiency is 5 7 m /g
Stevens et al (1988) reported measurements made during the winter of 1986-1987 in
Boise, ID Nitrate aerosol was a significant component of total light extinction, contributing
11-25
-------
13% of the fine-particle mass Less than 10% of the total nitrate was left in the vapor phase
as HNO3. Measurements in this study were made using an annular denuder followed by
»
Teflon and nylon filters
Malm et al (1989) evaluated the contribution of nitrate aerosol, along with larger
contributions from sulfate and carbonaceous aerosols, to wintertime visibility impairment in
the scenic Southwest near Grand Canyon and Canyonlands national parks Nitrate
concentrations during January and February 1987 at Grand Canyon averaged 0 1 to
3
0.3 /tg/m Multiple linear-regression analysis suggested that nitrate particles had an average
scattering efficiency of 4 7 m2/g and contributed 6 to 14% of the fine-particle light extinction
during the wintertime study Nitrate was generally a much smaller contributor, however, to
light extinction than sulfates, which contributed 62 to 72 % of fine-particle extinction, and
organics, which contributed 15 to 16%
Richards et al (1991) measured aerosol composition in and near the Grand Canyon
during January through March 1990 Ammonium nitrate was 6.4 to 10 4% of the fine-
particle mass at three locations in the Grand Canyon
11.5.2 Earlier Measurements
Because these earlier measurements have significant positive and negative nitrate
artifacts, they are less accurate than the previous studies Such biases should be kept in
mind.
White and Roberts (1977) studied the statistical relationships between light-scattering
coefficient and the aerosol constituents of Los Angeles area smog measured during the
summer and early fall of 1973 as part of ACHEX (Aerosol Characterization Experiment)
Using linear-regression techniques, they estimated that nitrate aerosols contributed, on
average, about 27% of the total light-scattering coefficient Nitrates were found to have a
f\ *\
light-scattering efficiency, having units of m /g of nitrate anion, of 2 9 + 6 5 p , where /i, is
the relative humidity as previously defined Thus, at a humidity of 50%, the light scattering
f\
coefficient of nitrates was estimated to be 4 5 m /g Appel et al (1985) have commented
that White and Roberts (1977) may have senously underestimated nitrate scattering
efficiencies because the glass-fiber filters used to collect aerosol samples had a strong
11-26
-------
positive artifact (i e , gaseous HNO3 was deposited on the filter, thereby inflating the nitrate
aerosol measurement)
Cass (1979) used linear and nonlinear regression to study the relationships between
sulfate and nitrate concentrations and visibility in Los Angeles from 1965 through 1974
Sulfates and nitrates were found to be significant contributors to total hght extinction The
best fits to measured visibility were obtained with regression coefficients of the form,
131'(1 - jtt), where /* is the relative humidity as defined previously This is indicative of
hygroscopic or deliquescent properties of sulfate and nitrate The values for /? for sulfate
2
and nitrate amon were 5 3 and 3 3 m /g, respectively At 50% RH, this would yield overall
respective light-extinction efficiencies for sulfate and nitrate anions and associated water of
2
10 7 and 6 6 m /g The nitrate measurements used by Cass were subject to positive
artifacts The nitrate data were 24-h averages, whereas the extinction data were daytime
averages Light extinction was denved from visual-range observations rather than
nephelometer or transmissometer measurements A Koschmieder constant of 3 9 rather than
3 0 as recommended in Equation 11-7 was used, thereby biasing extinction values high It is
not clear whether the two positive biases would cancel each other out
Tnjoms et al (1982) investigated the visibility-aerosol relationship m California using
data from 34 locations They found that NO2 contributed a rather uniform 7 to 11 % of total
hght extinction (bext) throughout California Although they were not of adequate quality to
make definitive statements, the data suggest that nitrates are more important contributors to
bext in northern California, where they may contribute 10 to 40% of bext
Outside of California, the most significant urban hazes that have been shown to be
associated with NOX occurred in the winter in Denver and Phoenix Nitrogen oxides, both
NO2 and nitrate aerosol, were found to be significant contributors to the winter haze in
Denver (Groblicki et al , 1981), even with the significant negative artifacts of the
measurement techniques used Multivanate statistical analysis (regression) was used to
analyze the relationships between light scattering and absorption and concentrations of
particles and gases measured on 41 consecutive days in November and December 1978
Most of the hght extinction was found to be caused by particles < 2 5 jwm in diameter
Elemental carbon (soot) was found to be the most significant contributor, accounting for 37%
of light extinction above natural Rayleigh background Sulfate (and associated water) was
11-27
-------
found to contribute 20%, nitrate (and associated water), 17%, and organic carbon, 13%, the
remaining fine-particle matter contributed 7%, and NO2 contributed 6% All measurements
were based on a wavelength of light of 0 475 /*m (Hasan and Dzubay, 1983) If the
contribution of nitrate and NO2 are combined, the total NOX contribution to Denver winter
haze is 23 %, second only to the contribution of elemental carbon, however, this is probably
an underestimate of the NOX contribution because of the negative nitrate artifact
Still, data from the Grobhcki et al (1981) study may be better than some of the other
data for Denver because of the cold temperatures and high NH3 concentrations found in
Denver during the study
Wolff et al (1981) determined the emission source contributions to the Denver winter
haze Of the total NOX contribution to the winter haze of 23 % (also an underestimate
because of the negative nitrate artifact), combustion of natural gas, oil, and coal (in power
plants and boilers) accounted for more than half (14%), and automotive contributions were
the largest part of the remainder (9%) Hasan and Dzubay (1983) developed estimates of
light extinction efficiency of various aerosol components of the 1978 Denver winter haze
using both regression analysis and Mie scattering theory based on measured particle size
distributions For nitrate amon, regression gave a scattering efficiency of 3 1 to 3 2 m2/g,
^
whereas theoretical calculations yielded a scattering efficiency of 4 8 to 4 9 m /g
Solomon and Moyers (1984) studied the contributors to light extinction m Phoenix
during January 1983, when winter hazes were observed Elemental carbon was estimated to
be the largest contributor to light extinction, at 41 % of bext, on average Approximately
equal contributions resulted from nitrate (15%), organic carbon (15%), and sulfate (13%)
However, these estimates are biased because they used the Groblicki et al (1981) regression
equations. The contribution from NO2 averaged 32% Solomon and Moyers (1986)
reported that the fine nitrate aerosol measured m Phoenix in January 1983 was 13 4% of the
total fine-particle mass, comparable to the 12 2% contribution of nitrate found in Denver
during November and December 1978 and much higher than the contribution reported m
other major metropolitan and rural areas However, they concluded that their nitrate
measurements were significantly positively biased They concluded that motor vehicle
emissions accounted for most of the nitrate and other fine-particle mass that caused the
observed haze
11-28
-------
Few studies of the role of nitrate aerosol in visibility impairment have been conducted
outside of the western United States Nitrate aerosol contributions appear to be lower in the
eastern United States than in California and other western U S areas, perhaps because of
higher sulfate concentrations competing for the available atmospheric NH3
Using multiple linear-regression techniques, Tnjonis and Yuan (1978a) found that
nitrate did not account for any of the observed light extinction in most of the cities in the
northeastern and north central United States Nitrates accounted for 8 % of total light
extinction in Columbus, OH There the light extinction efficiency of nitrate was estimated
2
from regression analysis to be m the range of 6 to 9 m /g
Wolff et al (1982) found that nitrate contributed minimally to light extinction in Detroit
o
during July 1981 Fine-particle nitrate averaged 0 2 jttg/m , coarse-particle nitrate was
3
higher, at 1 /tg/m This was consistent with other measurements made in the eastern United
States (Ferman et al, 1981), where little nitrate was found in the fine fraction Nitrogen
dioxide contributed 4% of bext in the Wolff et al study (1982)
Dzubay et al (1982) studied the relationships between visibility and aerosol
composition during summer in Houston, TX Nitrate was found mainly on coarse particles
and was determined to be an insignificant (05%) contributor to the total light extinction
It was conjectured that fine nitrate aerosol did not condense because the sulfate was not fully
neutralized (i e , there was insufficient NH3 to react with HNO3), and that HNO3 condensed
on the alkaline coarse particles, which were a significant sink for nitrate Nitrate particle
measurement artifacts may also have been a major factor m this study Nitrogen dioxide
contributed 4 7% of bext
Colbeck and Harrison (1984) found significant quantities of nitrate aerosol in northwest
England Visibility there was strongly correlated with both nitrate and sulfate concentrations
Diederen et al (1985) investigated the nature of the haze m western Netherlands during the
period 1979 to 1981 Ammonium nitrate aerosol was found to contribute 35% of total bext,
and NO2 to contribute 2%
Bravo et al (1988) found high concentrations of nitrate aerosol and NO2 in Mexico
3
City (6 4 jwg/m and 0 07 ppm, respectively), however, the relative contributions of these
species to the total light extinction budget were small (5 and 25%, respectively) because of
11-29
-------
the much higher concentrations of other aerosol species Total light extinction was
dominated by soot (31%), sulfate (30%), orgamcs (15%), and other species (16%)
The effects of NO2 and nitrate on regional haze outside of urban areas appear to be less
significant than their effects on urban hazes Nitrogen oxides may not be significant in these
nonurban regional hazes because of low concentrations of HNO3 and NH3, high ambient
temperatures, and low humidities in the West, and because of high sulfate concentrations in
the East that compete for available NH3
Macias et al (1981) found that nitrate made small or negligible contributions to
regional haze at one site in Arizona on several monitoring days in the summer and winter of
1979, although on one day NH4NO3 was about 8% of the fine-particle mass However,
these measurements were negatively biased
White and Macias (1987) found very low concentrations of nitrate aerosol in the
nonurban, intermountain West Measurements of nitrate aerosol concentrations averaged
o
0.09 jtig/m . Nitrate was very episodic, however, with major contributions to this average
arising from a small number of episodes Higher concentrations were observed in the North
and at all sites during the winter White and Macias (1987) commented that during the
winter the measurements may have underestimated nitrate aerosol concentrations by as much
as a factor of three because of nitrate volatilization from the filters
Tnjoms et al (1988) analyzed data collected in the Mohave Desert of California over a
2-year period, 1983 to 1985, to determine the species contributing to light extinction They
found that for both average and worst-case conditions the sum of particulate nitrate and NO2
contributed 13 ± 5 % of non-Rayleigh bext, however, nitrate measurements were subject to
artifacts
Mathai and Tombach (1987), in their review of visibility and aerosol measurements in
the eastern United States, concluded that fine nitrate concentrations averaged 1
In the studies they summarized, fine-particle nitrate had been measured for very short (week
q
and month) periods and concentrations had ranged from 0 2 to 0 9 /wg/m
Wolff and Korsog (1989) found that NO2 (averaging 4 ppb) accounted for less than
1 % of total light extinction in the Berkshire Mountains of Massachusetts in the summer of
1984. Sulfate and associated water caused most (77%) of the light extinction Nitrate
aerosol was not found The measurements of Vossler et al (1989) at Deep Creek Lake in
11-30
-------
Maryland and of Pierson et al (1987) in the Allegheny Mountains were consistent with the
Berkshire Mountains study, NO2 averaged 4 ppb, and nitrate aerosol concentrations were
very small relative to sulfate The latter two studies, unlike the Berkshire study, used the
more accurate denuder-nylon filter samples
Dzubay and Clubb (1981) found that for summer conditions in Research Triangle Park,
NC (nonurban but near urban areas), the sum of the scattering and absorption coefficients by
species accounted for about 90% of the measured bext Particle scattering caused most of the
light extinction (75%), followed by Rayleigh scattering from air (7%) and particle light
absorption (7%), NO2 light absorption accounted for only 2% of total light extinction
11.6 MODELING REGIONAL AND URBAN HAZE EFFECTS
Latimer et al (1985a) used a Lagrangian regional visibility model and emission
inventories for the southwestern United States to estimate the effects of manmade emission
sources on regional visibility in 1980 and 1995 In this assessment, nitrate aerosol was
found to be a potentially significant contributor to the manmade portion of nonurban regional
haze While manmade sulfate sources were found to be the largest contributors to haze,
contnbutmg over half (50 to 60%) of the manmade fraction, nitrate was estimated to be the
next largest contributor (10 to 20%) Although manmade organic and elemental carbon
contributions to regional haze were found to be small (less than 10 % of the manmade
fraction), biogemc organic aerosol was estimated to be a large contributor to total light
extinction (the sum of natural and manmade fractions)
In this modeling study, it was cautioned that the estimates of the contribution of nitrate
to the manmade total were uncertain because of uncertainties in the relative distribution of
the nitrate anion (NO;;) between optically inactive HNO3 and light-scattering NH4NO3
aerosol This uncertainty resulted largely from the uncertainty regarding background
concentrations of NH3, which is essential to the formation of NH4NO3 aerosol On the basis
of thermodynamic equilibrium considerations, the stud} showed that nitrate aerosol would be
most likely to condense in winter and least likely in summer Nitrate aerosol was found to
be a significant portion of increases in regional haze projected for the period 1980 to 1995
Latimer et al (1985b, 1986) evaluated the performance of this regional visibility model by
11-31
-------
comparing model calculations with participate, visibility, and wet deposition measurements
performed by the U S Environmental Protection Agency (EPA), the National Park Service,
and the Electric Power Research Institute This comparison showed that model predictions
of sulfate and nitrate concentrations and light extinction were only slightly biased and were
highly correlated with actual measurements The average nitrate aerosol concentration
<2
predicted by the model was 0 22 /*g/m , approximately 2 4 tunes the average measured
during the Western Regional Air Quality Study in 1981 of <0 1 /xg/m3 that was reported in
Tombach et al (1987) and the value of 0 09 /xg/m3 reported by White and Macias (1987),
however, these latter studies had negative artifacts
Latimer et al (1986) and Latimer (1988c) applied this regional visibility model to the
case of whiter layered haze observed near the national parks in Utah and Arizona
An average nitrate aerosol concentration of 0 35 /*g/m was predicted This value compares
3
reasonably well with the average of 0 16 jwg/m measured during a special study in 1986
o
(Latimer, 1988c) and the average of 0 38 /*g/m measured during the WEDLTEX experiment
in 1987 (Malm and Iyer, 1988) However, the model underpredicted the observed sulfate
concentrations by a factor of two to four Although considerable uncertainty exists over the
accuracy of nitrate measurements (Malm and Gebhard, 1988), nitrate may be a significant
contributor to winter layered haze (approximately 15 to 25 % of extinction from manmade
sources, according to Malm et al, 1989), even though sulfate appears to be the dominant
contributor
Latimer (1988a) developed a spreadsheet template for calculating the effect of changes
in aerosol species concentration on total light extinction and visibility As part of that effort,
available measurements of chemical composition and concentration of particles and of
visibility or light extinction were compiled Using an assumed nitrate light-scattering
2.
efficiency of 8 m /g, Latimer (1988a) estimated the relative contribution of nitrate to total
light extinction in numerous locations where both aerosol and visibility data were available
Nitrate generally contributed less than 10% to total extinction, except in Portland, OR, where
it was 11 to 14%, Denver, CO, 16%, Los Angeles, CA, 20%, and Riverside, CA, 40%
Latimer (1988a) found that measured visual ranges agreed well with visual ranges denved
from the measured aerosol constituents and their respective light-extinction efficiencies
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Russell and Cass (1986) developed a Lagrangian trajectory model that incorporates
gaseous and aerosol chemistry and aerosol equilibrium This model was applied to a smog
episode in Southern California Predictions from the model compared well with
measurements of O3, NO2, HNO3, NH3, PAN, and pardculate nitrate When the model was
used to investigate alternative control techniques for nitrate aerosol, NOX emission control
was found to produce a nearly proportional (linear) reduction in total nitrate (HNO3 vapor
plus particulate nitrate) and slightly greater than proportional reductions in particulate nitrate
Paniculate nitrate concentrations were found to be effectively reduced by reducing NH3
emissions, especially from farm-related activities
Russell et al (1988a,b) developed and applied a grid-based Eulenan airshed model that
incorporates a chemical reaction mechanism for gaseous and aerosol species The model was
compared with measurements and the model calculations of aerosol nitrate concentrations
were found to be in good agreement with measurements
Pilims and Seinfeld (1987) developed the SEQUTLEB model, which consists of
thermodynamic equilibrium relationships that describe the behavior of the HNO3, NH4NO3,
NH3, NH4+, SO4=, Cl", and H2O chemical system (Stelson and Seinfeld, 1982a,b,c, Bassett
and Seinfeld, 1983, 1984, Saxena et al , 1986, Pilmis et al, 1987) This model calculates
the equilibrium concentrations of these species in the gas and aerosol phases A model of
this type is essential for calculating the amount of aerosol nitrate formed and the water
content of hygroscopic aerosols This model was applied in the Phoenix winter haze study
(Watson et al, 1991) to assess the degree of nitrate and NH3 control required to reduce
NH4NO3 aerosol concentrations
Reactive plume models have been developed (Joos et al, 1987, Hudischewskyj and
Seigneur, 1989) that incorporate such equilibrium models and aerosol coagulation models to
calculate aerosol size distributions of nitrate and other aerosols Zhang (1991) has developed
mathematical models to calculate light-extinction efficiency from aerosol composition
11.7 ROLE OF NITROGEN OXIDES IN PLUME VISUAL IMPACT
Much of the regulatory attention that has been given to visibility during the past decade
has focused on the issue of the visibility impacts of plumes from individual emission sources
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This plume visual impact is commonly called "plume blight" (U S Environmental Protection
Agency, 1979) Particularly in areas of pristine background visibility, such as the
intermountain West, the visual impact of plumes such as those from power plants can be
quite significant as far as 100 km from sources (U S Environmental Protection Agency,
1979; Latimer, 1979, 1980) Considerable work has been earned out during the past decade
to develop and evaluate computer models of plume visual impact and to develop technical
guidance for plume visual impact evaluation as part of the implementation of EPA's visibility
regulations under the visibility protection provisions of the Clean Air Act Nitrogen dioxide
has been found to be a significant contributor to plume visual impact from modern,
well-controlled power plants
The contrast of a plume against an optically thick horizon-sky background can be
calculated by solving Equation 11-1 (Latimer et al , 1978, White et al , 1986)
Cplume = [Jplume/Jback ~ 1] U ~ exp(-Tplume)] [exp(~bext rp)], (11-8)
where
cplume = contrast of the plume against the horizon sky (Pplume - I sky)/Isky],
J = source function defined previously,
Tpiume = optical thickness of the plume ( $ bext dr),
t>ext = extinction coefficient of the intervening background atmosphere between
the plume and the observer, and
rp = distance between the plume and the observer
For a pure NO2 plume, the first term (in the first pair of square brackets) equals -1,
and therefore Cpjume is always negative, signifying a dark plume If one also assumes either
that the plume is very close to the observer (rp « 0) or that the intervening atmosphere is
optically thin (bext « 0), then the last term in this equation equals 1, and the following
equation for an NO2 plume is obtained
Cplume = "[I - expC-Tplume)] = ~[1 ~ exP( $ plume bag dr)] C11'9)
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If one assumes that Cplume must equal at least -0 02 for a plume to be visible, then the
plume optical thickness (Tplume) must be at least 0 02 For a plume that is 1 km wide, this
optical depth can be caused by 0 065 ppm (122 /ig/m ) of NO2 at X = 0 55 /*m or by
0 012 ppm (22 /xg/m ) at X = 04 /*m For a plume 10 km wide, the same effect could be
caused by NO2 concentrations one-tenth as large Melo and Stevens (1981) found that under
typical conditions a plume NO2 optical thickness corresponding to 90 ppm-m (or 0 090 ppm
in a 1 km wide plume) was required to make a plume just visible against a blue horizon-sky
background Using a predecessor of the PLUVUE models (Johnson et al , 1980, Seigneur
et al , 1984), Latimer (1980) investigated the relationship between NOX emission rates from
power plants and plume contrast and other optical parameters He found that the yellow-
brown coloration of the power plant plume was dominated by NO2 for the modeled cases
Melo and Stevens (1981) confirmed the dominant importance of NO2 to coloration in an
actual power plant plume Latimer (1979, 1980) modeled the visual impacts of power plants
of various sizes and NOX emission rates and concluded that yellow-brown plumes could be
observed as far as 100 to 150 km away from a power plant, but only on a few days per year
White and Patterson (1981) developed nomographs that allow one to determine the
optical properties and relative importance of emitted particles and NO2 as a function of the
scattering angle and the particle size distribution Vanderpol and Humbert (1981) identified
NO2 as the primary plume colorant when particle size was greater than 0 5 pm Haas and
Fabrick (1981) performed a sensitivity analysis to investigate the effects of NO2 and particles
m plumes on various indicators of color and contrast
In studies of the Navajo Generating Station plume in the southwestern United States as
part of the VISTTA project, Richards et al (1981) never found paniculate nitrate, even
though HNO3 vapor was formed at rates 3 to 10 tunes the rate at which sulfate aerosol was
formed They concluded that nitrate aerosol did not condense because of inadequate
background concentrations of NH3 Hegg and Hobbs 1 1983) measured the constituents of
another power-plant plume in the Southwest and found rapid formation of both HNO3 and
nitrate aerosol Nitrate aerosol constituted 15 to 75 % of the nitrate in the plume Measured
plume aerosol size was primarily in the 0 25-/*m range Approximately equal contributions
to plume light extinction were made by particles and NO2 The reason the Hegg and Hobbs
(1983) findings were quite different from those of Richards et al (1981) is not clear, but the
11-35
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findings may have differed because background NH3 concentrations differed at the respective
sites
Also as part of the VISTTA study, Blumenthal et al (1981) measured the dispersion,
chemistry, and optical properties of the Navajo Generating Station On the basis of this
measurement program, they concluded that NO2 was the primary plume colorant, that
secondary aerosol formation could be neglected within 100 km of the source, and that the
PLUVUE model adequately characterized observed effects Bergstrom et al (1981)
evaluated the PLUVUE model using VISTTA data and found that the model performed
reasonably well, but that it slightly overpredicted observed plume visual impacts Sensitivity
analyses performed indicated that NO2 was the principal plume colorant
The most detailed evaluation of plume visibility models was earned out as part of the
VISTTA study (White et al, 1985, 1986) Four plume visibility models, including the two
versions of PLUVUE (Latimer and Samuelsen, 1975, 1978, Latimer et al, 1978, Johnson
et al., 1980, Seigneur et al, 1984), the ERT visibility model (Dnvas et al, 1980),
PHOENIX (Eltgroth, 1982), and the Los Alamos visibility model (Williams et al 1980,
1981), were evaluated by comparison with field measurements of plume concentrations,
optical parameters, and observed plume color and contrast made at the Navajo Generating
Station, well-controlled for particulate, at less well-controlled power plants in the Midwest,
and at an uncontrolled smelter in the Southwest Of the four, the first two, the PLUVUE
and ERT models, were found to be most accurate in predicting the plume visual impacts
observed in the field measurement programs The plume contrast for the power plant with
modern particulate controls could be adequately explained accounting just for the plume NO2
concentrations; particulates did not play a significant role In the study of strong particulate
emission sources (White et al, 1986), the performance of PLUVUE n and the ERT models
was less satisfactory than for the NO2-dommated plumes However, the relatively poor
performance of these two models may have resulted in large part from the imprecise
specification of model inputs (particle size and background sky radiance) Model
performance was found to depend strongly on model input specification
11-36
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11.8 SUMMARY OF EFFECTS ON VISIBILITY
Emissions of NOX can contribute significantly to visibility impairment in the form of
plumes and hazes Nitrogen dioxide and NH4NO3 are the optically active species of NOX
Other species, including NO and HNO3, are gases with insignificant optical effects
Nitrogen dioxide is a gas that preferentially absorbs blue light, thus tending to cause yellow-
brown atmospheric discoloration There is agreement among many studies that NO2 is a
strong and consistent colorant Aerosols, however, including nitrate, can cause atmospheric
discoloration, particularly when bright objects are observed or the sun is behind the observer
Nitrogen dioxide has been shown to be the most significant plume colorant for the
yellow-brown power plant plumes that have been observed, primarily in the western United
States, and that are of current regulatory concern to EPA and the States
Nitrogen dioxide and nitrate aerosol are significant contributors to urban haze,
especially in California and the western United States Their combined share of total
extinction can be 20 to 40% of total light extinction in such urban areas In nonurban areas,
NOX appears to be a relatively small contributor to light extinction because NO2, nitrate
aerosol, and NH3 concentrations tend to be lower or because moderate or high temperatures
tend to prevent nitrate aerosol from condensing Nitrate aerosol does not appear in high
concentration in areas of high concentrations of acid sulfate, such as the eastern United
States, mainly because acidic sulfate compounds consume the available atmospheric NH3 that
is needed to condense nitrate aerosol from HNO3 vapor
Theoretical models have been developed for describing the chemical reactions that
result in the formation of optically active NOX species, aerosol dynamics of nitrate aerosol,
chemical equilibrium of nitrate-water aerosols, the light scattering and absorption properties
as a function of the wavelength of light, and effects on visual range, haze contrasts, and
atmospheric color The available comparison of plume visibility models suggests that the
effects of plume NO2 can be accurately predicted but that model predictions of the effects of
aerosol particles are less adequate Limited work has been done to develop and test models
for urban, layered, and regional haze, but much more work is clearly needed
Measurement of nitrate aerosol is complicated by its volatility However, newer
measurement techniques based on the use of denuders have provided reliable measurements
11-37
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Because older techniques (such as Teflon filters) can seriously underestimate nitrate aerosol
concentrations, care must be taken when interpreting data obtained by those techniques
Work is needed to understand the apparently nonlinear effects of NOX emission controls
on nitrate aerosol concentrations and resulting visibility effects Also, work is needed to
understand the effects of sulfur dioxide emission controls on nitrate aerosol production,
because the large-scale reduction of sulfate, which competes with nitrate for available NH3,
may result in increases in nitrate aerosol
11.9 ECONOMIC VALUATION OF EFFECTS ON VISIBILITY FROM
NITROGEN OXIDES
Hie primary effects of NOX on visibility were descnbed in previous sections of this
chapter and are believed to be (1) discoloration, producing a brownish color seen in plumes,
layered hazes, and uniform hazes, and (2) reductions in visual range (increases in light
extinction), especially in urban areas in the western United States This section discusses the
available economic evidence concerning the value of preventing or reducing these types of
effects on visibility Economic studies have not focused specifically on NOX- associated
changes in visibility for the most part, but some studies have considered the types of
visibility effects that are associated with NOX The following summary of economic
estimation methods and available results is brief For more detail see Chestnut and Rowe
(1990a), Mitchell and Carson (1989), Fischhoff and Furby (1988), Cummings et al (1986),
and Rowe and Chestnut (1982)
11.9.1 Basic Concepts of Economic Valuation
Visibility has value to individual economic agents primarily through its impact upon
activities of consumers and producers Studies of the economic impact of visibility
degradation by air pollution have focused on consumer activities Most economic studies of
the effects of air pollution on visibility have focused specifically on the aesthetic effects to
the individual Some commercial activities, such as airport operations, may be affected by
visibility degradation by air pollution, but available evidence suggests that the economic
magnitude of NOX effects on commercial operations probably is very small In a 1985
11-38
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report, EPA concluded that some percentage of the visibility impairment incidents sufficient
to affect air traffic activity might be attributable, at least in part, to manmade air pollutants
(possibly 2% to 12% in summer in the eastern United States), but according to the
information presented previously in this chapter, NOX would not be expected to be a
significant contributor to these incidents
It is well established that people notice those changes in visibility conditions that are
significant enough to be perceptible to the human observer, and that visibility conditions
affect the well-being of individuals This has been verified in scenic and visual air quality
rating studies (Middleton et al , 1983, Latimer et al , 1981, Daniel and Hill, 1987), through
the observation that individuals spend less tune at scenic vistas on days with lower visibility
(MacFarland et al, 1983), and through use of attitudmal surveys (Ross et al, 1987) The
intent of visibility-related economic studies has been to put a dollar value on changes in well-
being associated with visibility degradation
Welfare economics defines a dollar measure of the change in individual well-being
(referred to as utility) that results from a change in the quality of any public good, such as
visibility, as the change in income that would cause the same change in well-being as that
caused by the change in the quality of the public good One way of defining this measure of
value is to determine the maximum amount the individual would be willing to pay to obtain
improvements or prevent degradation m the public good (see Freeman [1979] for more
detail) For most goods and services traded in markets, this measure can be derived from
analysis of market transactions For non-market goods, such as visibility, this economic
measure of value must be derived some other way
For purposes of this discussion, consumer values for changes m visibility can be
divided into use and non-use values (there are slight variations in the way these are defined
by different economists) Use values are related to the direct influence of visibility on the
current and expected future activities of an individual at a site Non-use values are the
values an individual places on protecting visibility foi use by others (bequest value) and on
knowing that it is being protected regardless of current or future use (existence value) Total
value, combining use and non-use, is sometimes called preservation value
11-39
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11.9.2 Economic Valuation Methods for Visibility
Two main economic valuation methods have been used to estimate dollar values for
changes in visibility conditions in various settings (1) the contingent valuation method
(CVM), and (2) the hedomc property value method Both methods have important
limitations, and uncertainties surround the accuracy of available results for visibility
Ongoing research continues to address important methodological issues, but at this tune some
fundamental questions remain unresolved (Chestnut and Rowe, 1990a, Mitchell and Carson,
1989, Fischhoff and Furby, 1988, Cummings et al, 1986) Recognizing these uncertainties
is important, but the body of evidence as a whole suggests that economic values for changes
in visibility conditions are probably substantial in many cases and that a sense of the likely
magnitude of these values can be derived in some instances from the available results
(Chestnut and Rowe, 1990a)
11.9.2.1 Contingent Valuation Method
The CVM involves the use of surveys to elicit values that respondents place on changes
in visibility conditions (see Rowe and Chestnut [1982], Mitchell and Carson [1989], and
Cummings et al [1986] for more details on this method) The most common variation of the
CVM relies on questions that directly ask respondents to estimate their maximum willingness
to pay (WTP) to obtain or prevent various changes in visibility conditions The potential
changes in visibility conditions are usually presented to the respondents by means of
photographs and verbal descriptions, and some hypothetical payment mechanism, such as a
general price increase or a utility bill increase, is posed
The CVM offers economists the greatest flexibility and potential for estimating use and
non-use values for visibility There are many types of changes in visibility for which total
values cannot be derived from market data As a result, most recent visibility value
applications use the CVM This approach continues to be controversial, however, and there
are those who question whether the results are useful for policy analysis (Fischhoff and
Furby, 1988, Kahneman and Knetsch, 1992) Smith (1992) has responded to some of the
questions raised about the CVM, but a consensus on its usefulness and reliability has not
been reached in the economics community Cummings et al (1986) and Mitchell and Carson
(1989) have conducted the most comprehensive reviews of the CVM approach to date and
11-40
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have concluded that there is sufficient evidence to support the careful use of results from
well-designed CVM studies in certain applications
Among the fundamental issues concerning the application of CVM for estimating
visibility values are (1) the ability of researchers to present visibility conditions in a manner
relevant to respondents and to design instruments that can elicit unbiased values, and (2) the
ability of respondents to formulate and report values with acceptable accuracy As with any
survey instrument, it is important that the presentation be credible, realistic, and as simple as
possible The optimal level of detail and the most critical pieces of information necessary in
the presentation to respondents to obtain useful CVM responses continue to be topics of
research and discussion Another important issue in CVM visibility research concerns the
ability of respondents to isolate values related to visibility aesthetics from other potential
benefits of air pollution control such as protection of human health Preliminary results
(Irwin et al, 1990, Carson et al, 1990) suggest that simply telling respondents before asking
the WTP questions to include only visibility is not adequate and may cause some upward bias
in the responses
11.9.2.2 Hedonic Property Value Method
The hedomc property value method uses relationships between property values and air
quality conditions to infer values for differences in air quality (see Rowe and Chestnut [1982]
and Tnjoms et al [1984] for more detail on this method) The approach is used to
determine the implicit, or "hedomc," pnce for air quality in a residential housing market,
based on the theoretical expectation that differences in property values that are associated
with differences in air quality will reveal how much households are willing to pay for
different levels of air quality in the areas where they live The major strength of this
approach is that it uses real market data that reflect what people actually pay to obtain
improvements in air quality in association with the purchase of their homes The method can
provide estimates of use value, but non-use values cannot be estimated with this method
There are many theoretical and empirical difficulties in applying the hedomc property
value method for estimating values for changes in visibility, but the most important limitation
is the difficulty in isolating values for visibility from other effects of air pollution at the
residence Hedonic property value studies to date provide estimates of total value for all
11-41
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perceived impacts resulting from air pollution at the residence, including health, visibility,
soiling, and damage to materials and vegetation The potential for estimating separate values
for visibility with this method is limited for two reasons First, the actual effects of air
pollution often are highly correlated, making it difficult to separate them statistically usmg
objective measures Second, individuals are likely to perceive a correlation between these
effects and to act accordingly in their housing decisions, even if the effects are actually
separable using objective measures
11.9.3 Studies of Economic Valuation of Visibility
Economic studies have estimated values for two types of visibility effects potentially
related to NOX: (1) use and non-use values for preventing the types of plumes caused by
power plant emissions, visible from recreation areas in the southwestern United States, and
(2) use values of local residents for reducing or preventing increases in urban hazes in
several different locations
11.9.3.1 Economic Valuation Studies for Air Pollution Plumes
Three CVM studies have estimated on-site use values for preventing an air pollution
plume visible from recreation areas in the southwestern United States (Table 11-2) One of
these studies (Schulze et al, 1983) also estimated total preservation (use and non-use) values
held by visitors and non-visitors for preventing a plume at the Grand Canyon A fourth
study concerning a plume at Mesa Verde National Park (Rae, 1983) was not included
because of methodological problems with the contingent ranking approach used (Ruud,
1987) The plumes in all three studies were illustrated with actual or simulated photographs
showing a dark, thin plume across the sky above scenic landscape features, but specific
measures such as contrast and thickness of the plume were not reported Respondents were
told that the source of the plume was a power plant or an unspecified air pollution source
In one study (Brookshrre et al, 1976), a power plant was visible in the photographs
The estimated on-site use values for the prevention or elimination of the plume ranged
from about $3 to $6 (1989 dollars) per day per visitor-party at the park These value
estimates are comparable to values obtained in these and other studies for preventing fairly
significant reductions in visual range caused by haze at parks and recreation areas in the
11-42
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Southwest A potential problem common to all of these studies is the use of daily entrance
fees as a payment vehicle Respondents may have anchored on the then-typical $2 per day
fee and stated an acceptable proportional increase in entrance fees rather than reporting a
maximum willingness to pay This may have caused some downward bias in the responses,
but empirical exploration of this question is needed An alternative payment vehicle to
consider might be total expenditures for the top to the park
The results of the Schulze et al (1983) study suggest that on-site use values may be
easily dwarfed by total preservation values held by the entire population For example, with
average annual visitation at the Grand Canyon of about 1 3 million visitor-parties (about
three people per party), annual on-site use values for preventing a visible plume every day
would be about $8 miUion based on the Schulze et al results, whereas the implied
preservation value for preventing a visible plume most days (the exact frequency was not
specified) at the Grand Canyon would be about $5 7 billion each year when applied to the
total United States population There is, however, considerable uncertainty in the
preservation value estimates from this study Chestnut and Rowe (1990b) found that the
Schulze et al. (1983) preservation value estimates for haze at national parks in the Southwest
are probably overstated by a factor of two or three and the same probably applies to the
preservation value estimates for plumes
11.9.3.2 Economic Valuation Studies for Urban Haze
Six economic studies concerning urban haze caused by air pollution are summarized in
Table 11-3 Five of these are CVM studies and one is a hedonic property value study
Although many other hedonic property value studies concerning air quality have been
conducted (see Tnjonis et al [1984] and Rowe and Chestnut [1982] for reviews), the study
by Trijoms et al (1984) is the only one that has used visibility as the measure of air quality
The magnitudes of the changes in visual range considered in each study vary, making
direct comparisons of the results difficult In Table 11-3 implicit values obtained for a 10%
change in visual range are reported to allow a comparison of results across the studies
Values for a 10% change are shown to illustrate the range of results across the different
studies These estimates are based on a model developed for comparison purposes that
11-44
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assumes economic values are proportional to the percentage change in visual range Values
for a 20% change, for example, would be about twice as large as those shown for a 10%
change, given the underlying comparison model Each of these studies relied on a
reasonably representative sample of residents in the study area, such that a range of
socioeconomic characteristics and of neighborhood pollution levels was included in each
sample
The first five studies in Table 11-3 all focused on changes in urban hazes with fairly
uniform features that can be described as changes in visual range The sixth study (Irwin
et al , 1990) focused on visual air quality in Denver, where a distinct edge to the haze is
often noticeable, making visual range a less useful descriptive measure because it would vary
depending on the viewpoint of the individual and whether the target was in or above the haze
layer The studies conducted in Denver and in the California cities are the most relevant
because hazes in these cities are likely to have a higher NOX component than in the eastern
cities, but none of these studies focused specifically on NOX
Both of the CVM studies in California asked respondents to consider health and visual
effects but used different techniques to have respondents partition the total values They
found that, on average, respondents attributed about one-third to one-half of their total values
to aesthetic visual effects In spite of many similarities in the approaches used, the CVM
results for San Francisco are notably higher than for Los Angeles when adjusted to a
comparable percentage change in visual range One potentially important difference in the
presentations was that Loehman et al (1981) defined the change in visibility as a change in a
frequency distribution rather than simply a change in average conditions This type of
presentation is more realistic but more complex, and it is unclear how it may affect responses
relative to presentation of a change in the average It is possible that the distribution
presentation might elicit higher WTP responses because it may focus respondents' attention
on the reduction in the number of relatively bad days (and on the increase in the number of
relatively good days), whereas the associated change in the average may not appear as
significant The implied change in average conditions m the Loehman et al (1981)
San Francisco study was considerably smaller than that presented in the Brookshire et al
(1982) Los Angeles study, which may have also resulted in a higher value when adjusted to a
11-47
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comparable size change in average visual range because of diminishing marginal utility (i e ,
the first incremental improvement is expected to be worth more than the second)
The California studies in Los Angeles and San Francisco provide some interesting
comparisons because two different estimation techniques were applied for the same locations
Property value results for changes in air quality for both cities were found to be higher than
comparable values (for changes in total air quality) obtained in the CVM studies This is as
expected given the theoretical underpinnings of each estimation method, although Graves
et al (1988) have reported that subsequent analysis of the property value data revealed that
the estimates are more variable than the original results suggest These property value
results are not reported here because they are for changes in air pollution indices that are not
tied to visual air quality
The property value study results reported in Table 11-3 from Tnjonis et al (1984) were
estimated using light extinction as the measure of air quality However, as discussed in the
previous section on the hedonic property value method, these estimates are still likely to
include perceived benefits to human health for reductions in air pollution as well as values
for visual aesthetics Consistent with this expectation, the results for a 10% change in light
extinction are higher than the CVM results for visual range changes for the same cities
Respondents in several CVM studies have reported that, on average, they would attribute to
visibility aesthetics about one-fourth to one-half of their total WTP for improvements in air
quality This would imply that the Tnjonis et al results may reflect $25 to $100 for a
change in visibility alone
The results for the uniform urban haze studies in cities in the eastern United States fall
between the respective CVM results for the California cities The changes in visual range
presented in these studies were similar to those presented in the Los Angeles study In all of
the eastern studies respondents were simply asked to consider only the visual effects when
answering the WTP questions This approach is now considered to be inadequate (Irwm
et al, 1990; Carson et al, 1990)
A recent study that has not as yet completed the peer-review process has applied the
approach recommended in recent methodological explorations to estimate values for changes
in visibility McClelland et al (1991) conducted a mail survey in 1990 in Chicago and
Atlanta Residents were asked what they would be willing to pay to have an improvement in
11-48
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air quality, which amounted to about a 14% improvement in annual average visual range
Respondents were then asked to say what percentage of their response was attributable to
concern about health effects, soiling, visibility, or other air quality impact Respondents, on
average, attributed about 20% of their total WTP to visibility The authors conducted two
analyses and adjustments on the responses One was to estimate and eliminate the potential
selection bias resulting from non-response to the WTP questions (including what has been
called protest responses) The other was to account for the potential skewed distribution of
errors caused by the skewed distribution of responses (the long tail at the high end) Both of
these adjustments caused the mean value to decrease The annual average household WTP
for the designated visibility improvement was $39 before the adjustments and $18 after the
adjustments This adjusted mean value implies about $13 per household for a 10%
improvement in visual range This is at the low end of the range of estimates shown in
Table 11-3 If peer-review of this research effort confirms the appropriateness of the study
design and analysis, the results suggest that greater confidence should be placed in the lower
end of the range of results shown in Table 11-3 because this study represents an
improvement in approach over the other eastern-cities studies
Irwin et al (1990) have reported preliminary results for the Denver study (Part n,
Table 11-3) Comparison of these preliminary results with results from other studies is
difficult because the photographs used to illustrate different levels of air quality were not tied
to visual range levels Instead, they were rated on a seven-point air quality scale by the
respondents, who were then asked their maximum WTP for a one-step improvement in the
scale This study reports some important methodological findings One of these is
confirmation that simply asking respondents to think only about visibility results in higher
WTP responses for visibility changes than when respondents are asked to give WTP for the
change in air quality and then to say what portion of that total they would attribute to
visibility only The latter approach produced a mean WTP estimate for a one-step change in
visibility that was about one-half the size of the mean WTP estimate given when respondents
were simply asked to think only about visibility This may result from the effect of budget
constraints on marginal values (the respondent has less to spend on visibility when he also is
buying health), however, the authors express the concern that some, but not all, of the value
for health may be included in the response when respondents are told to think only about
11-49
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visibility. They recommend that respondents be asked to give total values for changes in
urban air quality and then be asked to say what portion is for visibility
11.9.4 Summary of Economic Valuation
Visibility has value to individual economic agents primarily through its impact upon
activities of consumers and producers Most economic studies of the effects of air pollution
on visibility have focused on the aesthetic effects to the individual, which are, at this tune,
believed to be the most significant economic impacts of visibility degradation caused by air
pollution in the United States It is well established that people notice those changes in
visibility conditions that are significant enough to be perceptible to the human observer, and
that visibility conditions affect the well-being of individuals
Welfare economics defines a dollar measure of the change in individual well-being
(referred to as utility) that results from the change in the quality of any public good, such as
visibility, as the change in income that would cause the same change in well-being as that
caused by the change in the quality of the public good One way of defining this measure of
value is to determine the maximum amount the individual would be willing to pay to obtain
improvements or prevent degradation in the public good Two economic valuation
techniques have been used to estimate willingness to pay for changes in visibility (1) the
contingent valuation method, and (2) the hedomc property value method Both methods have
important limitations, and uncertainties exist in the available results Recognizing these
uncertainties is important, but the body of evidence as a whole suggests that economic values
for changes in visibility conditions are probably substantial in some cases, and that a sense of
the likely magnitude of these values can be derived from available results in some instances
Economic studies have estimated values for two types of visibility effects potentially related
to NOX: (1) use and non-use values for preventing the types of plumes caused by power
plant emissions, visible from recreation areas in the southwestern United States, and (2) use
values of local residents for reducing or preventing increases in urban hazes in several
different locations
Available evidence suggests that visitors to major recreation areas in the southwestern
United States value the prevention of manmade plumes visible from the recreation area The
results of two studies suggest values per visitor-party per day in the range of $3 to $6 (1989
11-50
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dollars) in additional park entrance fees to ensure that a thin, dark plume is not visible from
a popular observation point attJrand Canyon National Park A similar study at Lake Powell
found somewhat smaller values, in the range of $2 to $3 per day Schulze et al (1983)
found that total preservation values held by visitors and non-visitors for preventing a plume
visible from the Grand Canyon may substantially overwhelm on-site use values based on a
few dollars per day at the site, however, considerable uncertainty exists in the quantitative
results of this study, given the pioneering nature of the effort
The best economic information available for visibility effects associated with NOX is for
on-site use values related to changes in visual range in urban areas caused by uniform haze
These values fall roughly between $10 and $100 per year pei local household for a 10%
change in visual range in major urban areas in California and throughout the eastern
United States Reasonable extrapolations of on-site use values (with an order-of-magnitude
range of uncertainty) could be made from these studies for estimates of changes in visual
range that are attributable to changes in NOX levels in these and other major urban areas,
where NOX contributes to uniform haze that can be characterized by changes in visual range
Available results with regard to visual range in urban areas appear to be sufficient to
determine the importance of visibility values (on-site use) related to NOx-caused uniform
haze in urban areas relative to other potential benefits of NOX controls, and to provide order-
of-magmtude estimates of such visibility values To do so, however, would require estimates
of the changes in visual range that might be expected as a result of NOX controls
Extrapolations to less urbanized areas or to other visibility changes, or both, would
require additional assumptions and might introduce additional uncertainty Because each of
the studies completed to date has some important weaknesses and limitations, it would be
desirable to continue to enhance the geographic extent and the technical breadth of issues
addressed in these studies to arrive at a broader and more defensible set of estimates
Very little work has been done regarding layered hazes in recreation or residential
settings Preliminary results from Irwin et al (1990) suggest annual residential household
values of about $30 for a noticeable improvement in visibility conditions in the Denver area,
where layered hazes are common More information is needed about the specific visual
characteristics of such hazes that are most important to viewers, as well as about the value
people may place on reducing or preventing them
11-51
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11-62
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12. EFFECTS OF NITROGEN OXIDES
ON MATERIALS
12.1 INTRODUCTION
Materials exposed to the atmosphere in both indoor and outdoor environments may
suffer undesirable physical and chemical changes Although many of these changes occur
whether or not pollutants are present, the rate at which these changes occur can be influenced
by pollutant concentrations Nitrogen oxides (NOX), including nitric oxide (NO), nitrogen
dioxide (NO2), and nitric acid (HNO3), are known to affect the fading of dyes, the strength
of fabrics, plastics, and rubber products, the corrosion of metals, and the use-life of
electronic components, paints, and masonry Although the materials damage potential of
sulfur oxides (SOX) has been extensively studied, much less research has been reported for
NOX Graedel and McGill (1986) have pointed out, however, that sulfur dioxide (SO2)
concentrations are generally decreasing across the country Levels of NOX increased through
1985 but declined from 1985 through 1991 (U S Environmental Protection Agency, 1992)
The amount of materials damage attributable to NOX can therefore be expected not to
increase This chapter discusses the impact of NOX on a number of categories of materials
Emphasis is placed on those experiments and materials in which degradation was observed
To understand the results of materials exposure to NOX, it is important to appreciate the
influence of several factors on the materials damage pi ocess
1 The environment in which materials are exposed,
2 The mechanisms that cause damage in different exposures,
3 The wet and dry deposition processes that influence damage rates, and
4 The chemical interactions of NOX species with materials and with other
components of the environment, for example, other airborne pollutants and
moisture
It is also necessary to understand the experimental techniques used to study damage processes
and the limitations of these study techniques, as well as the results of the studies Finally, if
12-1
-------
estimates of the costs of materials damage are desired, an understanding of the economic
estimation procedures is needed A useful survey of the topic of air pollution damage to
materials is contained in Jorg et al (1985)
12.1.1 Environmental Exposures of Materials
The materials affected by NOX occur in both indoor and outdoor environments
Outdoor materials will be exposed to NOX concentrations such as those discussed in
Chapter 7 plus stresses caused by a wide range of temperatures and humidities, sunlight, and
precipitation Identical materials exposed in nearby locations may be damaged at very
different rates depending on their microenvironments (e g , building stone sheltered by an
overhang will be damaged at a different rate than stone openly exposed on the face of the
same structure). Most materials exposed for extended penods to the outdoor environment
are selected or designed to withstand these exposures and, therefore, they degrade at a slow
rate. Materials that may be subject to NOX damage and that are widely used outdoors
include paints, cement and concrete, stone, architectural and statuary metals, plastics, and
elastomers.
Indoor concentrations of NOX are discussed in Chapter 7 Although indoor
environments are free of many of the extreme environmental stresses present outdoors, NOX
concentrations may be significantly higher in some indoor environments (e g , where
unvented gas appliances are in use) and the matenals exposed indoors may be more sensitive
Virtually all the matenals found outside are also found indoors to some extent, however,
additional matenals such as paper, fine textiles, and electronic components are more common
in indoor than outdoor environments In addition, paint formulations intended for indoor
applications are different from those formulations mtended for outdoor use
12.1.2 Mechanisms of Materials Damage
Damage to exposed matenals results from attack through both physical and chemical
processes, and damage is induced both by pollution and by other agents Physical processes
include erosion by windborne particles, differential heating, and frost attack Chemical
processes include corrosion, biological attack (e g , mildew), direct attack by acid mists, and
gaseous and particle deposition and subsequent reactions (Tombach, 1982, Yocom and Baer,
12-2
-------
1983) It is difficult to distinguish a single causative agent for observed damage to exposed
materials because many agents, together with a numbei of environmental stresses, act on a
surface throughout its life Even some extensively studied systems (such as the effect of SO^
pollution on metals) are not thoroughly understood, and there is work still needed to
understand the interaction of NOX with the variety of materials in use today
12.1.3 Deposition Processes
For them to cause damage to a material, atmospheric pollutants such as NOX must come
in contact with the material Oxides of nitrogen are deposited on material surfaces through
both wet and dry deposition processes (Tombach, 1982) Dry deposition processes for
gaseous NOX include Browman or molecular diffusion 1o the surface, Stefan flow toward
surfaces where moisture is condensing, thermophoresis toward cold surfaces, and
diffusiophoresis toward evaporating surfaces In addition, particles containing NOX can be
transported to a material surface through gravitational settling or inertial impaction of the
particles on the surface Wet deposition (e g , acid ram) processes include the scavenging of
gaseous NOX or particles containing absorbed NOX into precipitation or fog droplets that
impact the surface The rate at which deposition processes transport NOX to the surface is
dependent on the NOX concentrations in the environment, the chemistry and geometry of the
surface, the concentrations of other atmospheric constituents, and the turbulent transfer
properties of the air (Lipfert, 1989)
The transfer of pollutants from the atmosphere to a surface is often visualized in terms
of the "multiple resistance analogy" (Sherwood et al, 1990) In this analogy, the rate of
mass transfer of pollutants is modeled as a series of resistances to the mass transfer
RT = Ra + Rb + Rc (12-1)
The total resistance, RT, is made up of the sum of "free air" turbulent transfer
resistance, R^ the near-surface, quasi-laminar boundary layer resistance, Rj,, and the surface
uptake resistance, R^ The aerodynamic resistance, Ra, is dominated by atmospheric
turbulence The boundary layer resistance, Rb, depends on the aerodynamics of flow
immediately adjacent to the surface and the molecular diffusivity of the pollutant The
12-3
-------
surface resistance, R^ depends on the physical and chemical interactions of the surface and
the pollutant Depending on the aerodynamic conditions, and the physical and chemical state
of the surface, any of these terms can be the rate-limiting step for the transfer
The inverse of the total resistance is the deposition velocity, Vd (in units of cm/s) The
r\
deposition velocity is the ratio of flux of mass to the surface (g/cm s) to the free air
concentration of the pollutant (g/cm3)
In a laboratory study, Edney et al (1986) measured the deposition of NO2 and various
other compounds to both wet and dry galvanized steel A large "smog chamber"
(an environmental chamber designed to simulate photochemical processes) was used for the
study; NO2, propylene (C3H6), and SO2 were introduced in various combinations to study
deposition processes Galvanized steel was exposed both dry and wet with artificial dew
cycles caused by cooling the samples An experiment with a dry surface and NO2 alone
yielded a deposition velocity for NO2-to-galvanized steel of 0 05 cm/s A similar test with
SO2 yielded an SO2-to-galvanized steel deposition velocity of 0 8 cm/s, or deposition about
16 tunes greater for SO2 than for NO2 Dry deposition of NO2 on galvanized steel is thus
significantly slower than the dry deposition of SO2 These researchers suggest that, for the
purposes of developing a damage function representative of typical polluted atmospheres,
NO2 dry deposition on galvanized steel can be ignored
In a test with an NO2 and C3H6 mixture, Edney et al (1986) simulated smog conditions
that might be similar to Southern California conditions (i e , smog with very low SO2
concentrations) This experiment was allowed to proceed in the smog chamber for
336 h (2 weeks) with a total tune of induced dew of 196 h in 7-h periods At the end of the
experiment, concentrations in the gas phase and in dew on the surface of the galvanized steel
were measured Results are shown in Table 12-1 Fairly small amounts of nitrite ions
(NO2~) and nitrate ions (NO3~) were found on the surface and relatively little zinc was freed
(corroded). Clearly, however, the NO2 and other reactants had reacted to form a number of
species.
A test with NO2, C3H6, and SO2 was also run for comparison After 25 h, with a total
time of wetness of 14 h for the galvanized steel, the gas and surface-dew concentrations
shown in Table 12-2 were measured The gaseous species concentrations were similar to
those found in the previous test, except for SO2 Again, little nitrate or nitrite was found in
12-4
-------
TABLE 12-1. SMOG CHAMBER REACTIONS OF NITROGEN DIOXIDE
AND PROPYLENE AND DEPOSITION OF REACTION PRODUCTS
ON GALVANIZED STEEL
Chemical
Species
°3
CH3CHO
HCHO
PAN
NOX-PAN
HNO3
NO2"
NO3'
S04=
Zn
Gas-Phase
Concentration
(ppb)
134
254
621
57
359
7
—
—
—
—
Surface-Dew
Concentration
(nmol/cm )
—
—
133
—
—
—
11
77
1
77
Source Edney et al (1986)
TABLE 12-2. SMOG CHAMBER REACTIONS OF NITROGEN DIOXIDE,
PROPYLENE, AND SULFUR DIOXIDE AND DEPOSITION OF
REACTION PRODUCTS ON GALVANIZED STEEL
Chemical
Species
03
HCHO
PAN
NOX-PAN
HNO3
SO2
NO2'
S03=
NO3~
S04=
Zn
Gas-Phase
Concentration
(ppb)
240
1,150
114
159
9
1,190
—
—
—
—
—
Surface-Dew
Concentration
(nmol/cm2)
—
560
—
—
—
—
4
595
19
91
441
Source Edney et al (1986)
12-5
-------
the dew on the surface of the galvanized steel, especially when compared to the SOX
deposition. Furthermore, far more zinc was found in solution (i e , corroded) when SO2 was
added to the NO2-C3H6 mixture
The above laboratory studies illustrate both the complex nature of the NOX chemistry
and the relatively low deposition rate of NOX on galvanized steel In a subsequent field
experiment, Edney et al (1987) measured the ion concentrations for dry deposition and in
rainwater runoff from galvanized steel samples exposed outdoors in Research Triangle Park,
NC. The dry deposition ratio of sutfate ions (SO4=) to NO3" was 3 4, again illustrating the
relatively low deposition velocity of NOX compared to SOX for galvanized steel, this time
under outdoor exposure conditions This ratio might change as ambient concentrations of
SOX and NOX change These researchers speculated that the NO3" resulted from dry
deposition of HNO3 and particulate nitrate The ratio of dry to total nitrate deposition was
0.46, suggesting that wet and dry deposition appeared to play about equal roles in nitrate
deposition. Regression analysis of the ion concentration showed that the NO3" did not
significantly relate to the zinc in solution concentrations, however, SO4= concentrations were
in a one-to-one relationship with dissolved zinc Edney et al (1987) concluded that NOX is
not effectively deposited on galvanized steel surfaces and that sulfates dominate galvanized
steel corrosion
Although NOX deposition to galvanized steel may be insignificant, Spicer et al (1987)
found that there is a significant range of removal rates of NO2 by common indoor materials
2 1
Samples of 35 materials (surface area 33m) were exposed in chambers to 282 /*g/m
(0.15 ppm) NO2 (initial condition) at 50% relative humidity (RH) for 12 h and the rate of
NO2 removal was measured The results of these experiments are shown in Figure 12-1
Galvanized metal ducts were near the low end of removal rates measured in the Spicer et al
(1987) experiments Many common indoor materials (wallboard, wool carpet) were found to
have very high removal rates Nitric oxide gaseous concentrations were also monitored
during these experiments and were often found to increase as NO2 levels decreased The
author suggested that judicious selection of indoor materials might be considered as a means
of indoor NO2 control However, it was not possible from these experiments to determine
the amount of NOX accumulating on the surfaces of these materials, nor could conclusions be
drawn on any damage to indoor materials that might result from exposure to NO2
12-6
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01234567
WALLBOARD
CEMENT BLOCK
WOOL CARPET
BRICK (USED)
MASONITE
COTTON/POLYESTER BEDSPREAD
PAINTED (FLAT LATEX) WALLBOARD
PLYWOOD
ACRYLIC FIBER CARPET
NYLON CARPET
VINYL WALL COVERING (PAPERBACKED)
CEILING TILE
POLYESTER CARPET
ACRYLIC CARPET
FURNACE FILTERS (NEW)
DEHUMIDIFIER
OAK PANELING
VINYL-COATED WALLPAPER
PARTICLE BOARD
FURNACE FILTERS (USED)
CERAMIC TILE
WOOL (80%) POLYESTER (20%) FABRIC
COTTON TERRYCLOTH
SPIDER PLANTS (WITH SOIL COVERED)
WALLTEX COVERING
WAXED ASPHALT TILES
WINDOW GLASS
USED FURNACE HEAT EXCHANGER
FORMICA COUNTER TOP
POLYETHYLENE SHEET
ASPHALT FLOOR TILES
VINYL FLOOR TILE
GALVANIZED METAL DUCT
PLASTIC STORM WINDOWS
01234567 8 9
RATE CONSTANT FOR NOg REMOVAL (1/h)
Figure 12-1. Bar graph of nitrogen dioxide removal rate for various materials evaluated
in a 1.64-m test chamber at 50% relative humidity.
Source Spicer et al (1987)
12-7
-------
Miyazaki (1984) conducted a similar experiment, exposing common interior matenals
in a chamber to initial concentrations of 1,645 mg/m3 (875 ppm) NO2 and 1,124 mg/m3
(914 ppm) NO A summary of these results is shown in Table 12-3 The trend in these data
is similar to that reported by Spicer et al (1987), with wool carpeting and cement showing
relatively high deposition velocities for NO2 Vinyl floor tile, glass, and metals showed
relatively low deposition velocities for NO2 Insulation board and an ester/acrylic carpet,
materials not tested by Spicer et al (1987), had the highest deposition velocities Miyazaki
(1984) also found that NO2 deposition rates increased if turbulence, humidity, and
temperature were each increased in the chamber Increased turbulence escalates the rate of
delivery of NO2 to the surface Increased humidity probably results in dissolution of NO2
Increased temperature causes faster reaction rates
The deposition rates reported by Miyazaki appear to be low compared to the rates
reported by Edney et al (1986) The reason for the discrepancy is not apparent, however,
the differences may have been caused by different levels of turbulence in the two
experimental chambers Caution should be used in applying data from Miyazaki (1984) for
more than comparative purposes
12.1.4 Chemical Interactions of Nitrogen Oxides Species
Not only is there wide variation in the deposition of NOX to different surfaces but NOX
species themselves are reactive and their interactions with other atmosphenc constituents are
complex. Bassett and Seinfeld (1983) proposed a chemical equilibrium model for the
behavior of NOX, SOX, ammonia (NH3), and water in the atmosphere that is instructive for
understanding the role of NOX in matenals damage Nitrogen species (NO, NO2, HNO3,
etc.) are present as gases and in particulates (liquid and solid) and are deposited on material
surfaces Nitric acid is potentially the NOX species most directly damaging to matenals and
is formed by photochemically initiated reactions involving NOX in the atmosphere Under
dry conditions, HNO3 can deposit on a surface and can cause direct damage If liquid water
is present, HNO3 exists in equihbnum between the liquid phase in water solution and the
gaseous phase in the atmosphere However, Bassett and Seinfeld (1983) showed that in the
presence of atmosphenc NH3 and sulfunc acid (H2SO4), the HNO3 gas-phase versus liquid-
phase equilibrium is shifted toward the gas phase Thus, as nitrates accumulate on the
12-8
-------
TABLE 12-3. DEPOSITION VELOCITIES OF NITROGEN DIOXIDE AND
NITRIC OXIDE FOR INTERIOR MATERIALS
Deposition Velocity
(cm/s)a
Interior Material
Flooring materials
Carpet 1 (Acrylic fiber)
Carpet 2 (Acrylic fiber)
Carpet 3 (Acrylic fiber)
Carpet 4 (Wool)
Carpet 5 (30% Ester, 70% Acrylic fiber)
Tatami facing
Needle punch
Bath mat (100% Cotton)
Floor sheet 1 (Vinyl chloride)
Floor sheet 2 (Vinyl chloride)
Floor sheet 3 (Vinyl chloride)
Plastic tile
Ceramic tile
NO2
003
002
002
006
010
001
001
005
0001
0003
0003
0003
0004
NO
00003
—
—
—
—
0003
00008
—
000
—
—
—
—
Wall materials
Wallpaper 1 0 002 0 00
Wallpaper 2 0 002
Printed plywood 0 001
Ceiling materials
Insulation board Oil 0 00
Faulted insulation board 0 06 0 001
Plaster board 0 02 0 003
Wooden cement board 0 03 0 003
Asbestos cement board 0 04
Fittings
Glass 0 00 0 0008
Painted stainless steel 0 0008 0 001
Painted wood 0 003 0 0003
Curtain 0 0008 0 0003
Fusuma paper 0 003 0 002
Shoji paper 00003 00003
aThese values were averaged from the results of the experiments at 20 to 26 °C, 40 to 60% relative humidify
Source Modified from Miyazaki (1984)
12-9
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surface of a material, much of the accumulated nitrate mass may be evaporated into the
atmosphere as HNO3 Baedecker et al (1990) believe that this mechanism explains why
most post facto microanalytical investigations of damaged surfaces reveal very small amounts
of nitrogen species, whereas sulfates are frequently present It is also possible that, because
of their soluble nature, nitrate compounds have been washed off the damaged surfaces pnor
to analysis Wolff et al (1990) reported the results of a field study during which pollutant
fluxes were analyzed They found that SO4= accounted for 79%, on average, of the total
acidity of the wet deposition, whereas NO3" was responsible for 21 % of the acidity The
findings of Wolff et al (1990) indicate that, in polluted atmospheres containing SO2 and
condensing moisture, it is possible that NOX currently plays a relatively small role compared
to SO2 in causing the observed damage to most materials
12.1.5 Materials Damage Experimental Techniques
Because of the number of possible damaging agents and the complexity of synergistic
interactions, deposition processes, and exposure scenarios, researchers have typically relied
on controlled environmental chambers to quantify the damage rates attributable to specific
agents such as NOX Often materials exposure chamber studies are conducted at high
concentrations or at elevated temperatures and humidities in order to see damage within a
reasonable exposure period In addition, some chamber studies are conducted at low flow
rates that poorly simulate mass transfer properties in the natural environment and lead,
therefore, to underestimation of real-world deposition rates Also, the sequence in which
materials are exposed to different pollutants can affect the formation of protective corrosion
films, and this process is sometimes poorly simulated in chambers Although such studies
are useful, care should be exercised in the extrapolation of data and conclusions based on
chamber studies to effects expected from ambient exposures
The alternative to chamber studies has been ambient exposure studies In these
exposure studies, the materials of interest are usually exposed to ambient conditions at
several locations representing a spectrum of environmental variables (e g , temperature,
sunshine, humidity, pollutant concentrations) Statistical and chemical analyses are then used
to assess the contribution of the measured environmental variables to the materials damage
Again, the number of possible agents and the complexity of synergistic interactions makes it
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difficult to apportion observed damage among all the possible causes Franey and Graedel
(1985) reviewed the pollutant species that induce damage under actual ambient exposure
conditions, and have suggested that for any chamber study to be realistic, moisture,
radiation, carbon dioxide, reduced sulfur, a chlorine-containing gas, and a nitrogen-
containing gas must be included Because of the difficulties involved in apportioning the
causes of materials damage, reliable appraisals of the damage induced by NOX exposure
alone are not yet available
Both chamber studies and ambient exposure studies have come to rely on sophisticated
surface chemistry analytical techniques, as well as traditional bulk chemistry analyses and
measurements of physical properties Additionally, moisture collected from the samples
(runoff) has been analyzed for its chemical constituents The objective of these efforts is to
understand the chemical reactions occurring on the sample surfaces
Generally, little evidence of NOX species has been found in these analyses As noted in
the previous section, much of the NOX will be converted into HNO3 and subsequently will be
evaporated back into the atmosphere Thus, if HNO3 is leading to damage, it may not be
adequately accounted for in either surface chemical or runoff chemical analyses, and its role
in the damage process could be underestimated Better experimental techniques are needed,
both for investigating materials damage on the whole amd for determining the role played by
NOX
12.2 EFFECTS OF NITROGEN OXIDES ON DYES AND TEXTILES
12.2.1 Fading of Dyes by Nitrogen Oxides
Textile and dye manufacturers have recognized the problem of dye fading induced by
NOX for some tune Rowe and Chamberlain (1937) reported that dyes fade because of the
presence of NOX in combustion effluents Carpets, upholstery, and drapes that have been
subjected to elevated NOX levels in buildings using unvented gas heat have been observed to
fade within a year when dyes not resistant to NOX fading have been used Fading is
exacerbated when susceptible fabrics are dried in gas-fired clothes dryers, in which the
concentrations of NO2 can reach 3,760 ^g/m (2 0 ppm) (McLendon and Richardson, 1965)
Moreover, dryer exhaust is sometimes vented to the indoor environment to conserve heat and
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humidity, thus increasing indoor concentrations of NOX Textile and dye manufacturers have
responded to NOx-induced deterioration by seeking out and using NOx-resistant dyes or
inhibitors that forestall fading Fading from NOX has been observed on acetate, cotton,
nylon, rayon, silk, wool, and polyester
Nitrogen oxide-induced ("gas-fume") fading received serious attention when blue
disperse dyes were found to deteriorate significantly on cellulose acetate Salvin and
coworkers (1952) pointed out that NO2 is soluble in cellulose acetate, and that during
laboratory tests significant fading of dyes on the material can be observed within an hour
Hemplull et al (1976) tested a spectrum of dyes on various fabrics and found that NO2
caused significant fading on the cellulose acetate samples Salvin and Walker (1959) and
Salvin (1964) showed that alternative dyeing processes are available to minimize the impact
of NOx-induced fading on cellulose acetate, but that in many cases these substitute processes
and dyes are more expensive to use than the processes and dyes they replaced
Beloin (1973) exposed a variety of fabrics and dyes to 120 /ig/m3 (0 1 ppm) and
1,230 /*g/m3 (1 ppm) of NO, and 90 ^g/m3 (0 05 ppm) and 940 /^g/m3 (0 5 ppm) of NO2
for 12 weeks in an environmental exposure chamber He found that "appreciable" to "very
much" (the most severe category) fading occurred at both concentrations of NO for cottons
with direct, reactive, and vat blue dyes, cellulose acetate with disperse blue dyes, and nylon
with a blue dye. The cellulose acetate samples exposed to NO2 had generally greater
amounts of color change than the samples exposed to NO In addition, NO2 affected cotton
with direct and reactive red dyes, cotton with reactive blue dye, and rayon with direct red
dye. Beloin (1972) conducted tests on 67 dye-fabric combinations at 1 1 urban and rural sites
nationwide for 3-mo exposures The tests were conducted outdoors using chambers designed
to let the ambient air circulate around the samples but to exclude sunlight Using multiple
regression analysis, he sought to determine which pollutants played a significant role in the
observed change of colors on the fabncs He found that SO2 concentrations were significant
for 23 fabncs, ozone (O3) was significant for 8 fabncs, and NO2 was significant for
7 fabrics Fabnc-dye combinations affected by NO2 mcluded cellulose acetate with red and
blue disperse dyes, cotton muslin with reactive red and blue dyes, wool flannel with acid
blue dye, and the NOX gas-fading control nbbon recommended by the American Association
of Textile Chemists and Colonsts (AATCC) for testing NOX fading
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Cotton is the most widely used natural textile fiber and, again, significant gas-fume
fading has been noted Hayme et al (1976) exposed plum-colored cotton drapery fabnc to
NO2 in a chamber for 1,000 fa and found that serious lading occurred. Based on
o
extrapolation, they predicted that the use-life of draperies exposed to 100 /ig/m (0 053 ppm)
NO2 would decrease 19% In Beloin's chamber study described above, dyes on cotton were
found to experience "noticeable" to "much" fading when exposed to NO and "noticeable" to
"very much" fading when exposed to NO2 McLendon and Richardson (1965) found that
blue-dyed cotton fabnc became green after repeated NOX exposures in gas-fired dryers and
that the NOX exposure caused white fabnc to "yellow" Salvm (1969) reported the results of
sheltered, outdoor exposures of dyed cottons for 90 days in Los Angeles Thirty-one colors
of direct, vat, reactive, and sulfur dyes were tested and fifteen faded substantially The
author concluded that NOX and O3 were primarily responsible Hemphill et al (1976) also
demonstrated NOx-induced fading of vat, direct, and reactive dyes on cotton at
concentrations of 940 ^ig/rn3 (0 5 ppm) in a chamber for a 5-h exposure
Imperial Chemical Industries Limited (1973), a supplier of dyes for synthetics, issued a
technical bulletin on the gas-fume fastness of dyes used for nylon (polyamide) Nylons have
high resistance to wear and thus are often used as carpeting In this application, nylons are
exposed to indoor atmospheres for long penods Imperial Chemical Industry's bulletin
showed that several of the commercially available dyes faded noticeably on nylon when
exposed to NOX fumes and advised that these dyes not be used The susceptible dyes fade,
become duller in appearance, or acquire a redder or yellower cast Hemphill et al (1976)
demonstrated that certain blue and red dyes on nylon fade substantially when exposed to
O
940 /xg/m (0 5 ppm) NO2 Beloin's (1973) chamber study found that "appreciable" to "very
much" fading occurred on nylon fabncs exposed to NO or NO2 In outdoor exposures in
Los Angeles, Salvin (1964) found that nylon faded only slightly to very slightly
Other fabncs have been tested for dye gas-fading resistance as well Hemphill et al
(1976) investigated dye fadmg of rayon They found that two of the dyes tested, Direct Blue
86 and Direct Red 79, showed "noticeable" to "significant" fading Beloin (1973) found that
rayon withstood NO exposure with only a trace of fading, but exhibited "very much" fading
when exposed to NO2 In checking orlon, Hemphill et al (1976) found minimal dye fading
Salvin (1964) found that wool did not fade significantly in Los Angeles ambient exposures,
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but Hemphill et al (1976) showed moderate fading of red dye on wool in chamber
exposures Polyester exhibited very good dye-fading resistance in Salvin's Los Angeles
study (1964).
Whitmore and Cass (1989) report the results of a chamber study in which various art
materials were evaluated for color change due to NO2 exposure in the absence of light
The air in the exposure chamber was stirred and maintained at 24 °C and 50% RH for the
12-week exposure penods The NO2 concentration was 940 /tg/m3 (0 5 ppm) and the NO
concentration was 48 /tg/m3 (0 04 ppm) They tested 23 different natural dyes traditionally
used in Japan on silk and found that, in most cases, the changes were small The largest
color change occurred for enju (a dye made from the Japanese pagoda tree) Whitmore and
Cass rated the change as noticeable
The AATCC encourages textile manufacturers and suppliers to test dye and fabric
combinations for NOX fading These tests are routinely performed and NOx-susceptible dye
and fabric combinations rarely are produced in large quantities for the open market (Tew,
1990).
12.2.2 Degradation of Textile Fibers by Nitrogen Oxides
Nitrogen oxides not only affect fabric color, but can also alter the physical
characteristics of the fibers themselves, especially synthetic fibers Jellinek (1970) and
Jellmek et al. (1969) reported significant chain-scissioning of nylon after NO2 exposure
Chain-scissiomng is the breaking of the molecular structure that makes up a polymer and it
results in a loss of strength Vijayakumar et al (1989) found statistically significant amounts
of damage to nylon textiles exposed for 28 days to 0 1 ppm and 0 5 ppm concentrations of
HNO3 Zeroman et al (1971) investigated the impact of NO2 on acrylic, modacrylic, nylon,
and polyester yarn The yarns were continuously exposed in chambers for 1 week to
*a
simulated sunlight and 3,760 /*g/m (2 0 ppm) NO2 The yarn strength and rupture energies
were reduced for all materials The most seriously affected was nylon yarn, which lost
approximately 30% of its strength and 33% of its rupture energy as compared to control
samples exposed without NO2 The least affected was polyester, with about a 13% decrease
in strength The loss of strength of the acrylics was intermediate between the other two
yarns.
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12.3 EFFECTS OF NITROGEN OXIDES ON PLASTICS AND
ELASTOMERS
Plastics are highly polymerized materials, mostly synthetics, combined with other
constituents such as hardeners, fillers, and reinforcing agents (Hawley, 1981) Plastics
include fluorocarbon resins, phenolics, polyimides, polyethylene, acrylic polymers,
polystyrene, polyurethane, and numerous other synthetic compounds Major uses of plastics
include automobile bodies and components, boat hulls, building and construction materials
(pipe, siding, flooring), packaging (bottles, vapor barriers, drum linings), textiles (carpets,
cordage, hosiery), organic coatings such as paint and varnish vehicles, adhesives, electrical
components, and numerous other applications Use of plastics in the United States in 1980
was estimated at approximately 60 billion pounds per year, 01 double the 1970 consumption
Further development of and greater reliance on plastics are expected to increase the demand
for them in the future
Elastomers are synthetic polymers with the ability to stretch to at least twice their
normal length and retract rapidly to near their normal length when released Examples of
elastomers include butyl, mtnle, and polysulfide rubber, and neoprene Elastomers are used
for vibration dampers, wire coatings, fabrics, automobile tires, bumpers, and windshield
wipers, and other applications
Plastics and elastomers are subject to deterioration on exposure to ultraviolet (UV)
radiation, O3, SO2, and NOX Jellinek et al (1969) and Jellinek (1970) reported a series of
experiments in which a variety of polymers and elastomers were exposed to UV radiation
and pollutants in chamber experiments Jellinek et al (1969) reported the following results
for high concentration (nearly pure) NO2 exposures
1 Polyethylene minimal effect except for an increase m viscosity
2 Polypropylene some cross-linking (forming of additional chemical bonds) of the
polymer, although not as much as when exposed to SO2
3 Polystyrene some chain-scissiomng (breaking of chemical bonds)
4 Polymethyl methacrylate some chain-scissiomng (breaking of chemical bonds)
5 Polyvinyl chloride loss of chlorine due to reaction with NO2
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6. Polyacrylomtnle no significant change
7 Nylon chain-scissiomng occurs
8. Butyl rabbet- chain-scissioning
9. Polyisoprene appreciable chain-scissioning
10. Polybutadiene. cross-linking occurs
They concluded that damage to elastomers was generally greater than damage to plastics, but
that O3-induced damage was probably more important than NO2-induced degradation
Jellinek (1970) reported findings for the same series of plastics and elastomers at NO2
3 3
concentrations of 1,880 /ig/m and 9,400 /tg/m (1 and 5 ppm) for 1 h exposures At these
levels polymethyl methacrylate, nylon, and butyl rubber were found to suffer chain-
scissioning. Polyethylene, polypropylene, polyisoprene, and polybutadiene exhibited cross-
linking.
Krause et al. (1989) exposed polyvinyl chloride, polyurethane, glass-fiber-reinforced
polyester, and alkyd resin for 5 years in glass chambers to either 5,000 ^g/m NO2,
3 3
5,000 /*g/m SO2, 2,500 jtg/m O3, or a mixture of the pollutants The exposure cells were
kept at a humidity of 50 to 60% Half of each chamber was exposed to sunlight through
UV-transmitting glass The other half was kept dark The investigators found that most of
the degradation was caused by sunlight, with significantly less degradation occurring from
dark exposures to pollutants
Haynie et al (1976) exposed tire rubber and vinyl house siding to NO2, SO2,
O3, radiation, and humidity in a chamber Two NO2 concentrations, 94 and 940 /xg/m3
(0.05 and 0.5 ppm), were used with exposure tunes of 250, 500, and 1,000 h Various
combinations of other pollutants, radiation, and humidity were used in the exposures
The primary cause of damage to rubber was O3 exposure, and NO2 actually seemed to
inhibit the rate of O3-induced damage No appreciable damage to vinyl siding was observed
The National Research Council (1977) notes that discoloration and deterioration of strength
of foam rubber occurs from NO2 exposure
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12.4 EFFECTS OF NITROGEN OXIDES ON METALS
12.4.1 Role of Nitrogen Oxides in the Corrosion Process
Atmosphenc corrosion of metals is a serious problem and air pollution is known to
accelerate corrosion processes Sulfur oxides and chlorides are the atmospheric contaminants
most frequently implicated in the corrosion of metals Nitrogen oxides are also involved but
have received less attention Moisture enables these contaminants to form aggressive acids
that attack the metal surface and promote electrochemical reactions For this reason, both
pollutant concentrations and the "tune of wetness" (i e , how long liquid water is present on
the surface of the material) for exposed surfaces are important in determining the amount of
damage that will occur
For most metals, NOX alone as an attacking agent is much less aggressive than sulfur or
chlorine compounds Svedung et al. (1983), Kucera (1986), and Johansson (1986), however,
have pointed out the synergistic impact of NOX on atmospheric corrosion mechanisms
Using an exposure chamber, Kucera (1986) showed that carbon steel corrodes rapidly when
exposed to 3,421 /*g/m3 SO2 and 90% RH, but very slowly when exposed to SO2 at the
same concentration and 50% RH At 50% RH, steel corrodes about three tunes more
q
quickly when exposed to NO2 (5,640 /*g/m ) However, when both NO2 and SO2 at the
same concentrations are present at 50 % RH, the corrosion rate is approximately 30 tunes the
rate seen with SO2 alone Kucera noted that the presence of NO2 increases the rate of
deposition of SO2 on the metal surface Johansson (1986), also using an exposure chamber,
showed that NO2 deposition leads to the formation of hygroscopic nitrate-containing
corrosion products on the surface of the metal These corrosion products, in turn, absorb
moisture onto the surface, making the moisture available to mobilize other ions (such as
sulfates and chlorides) and thus leading to active corrosion at much lower relative humidities
than if NO2 were not present Effectively, NO2 acts to increase the time of wetness for the
surfaces Svedung et al (1983) showed similar results for gold-coated brass (a common
electrical contact), with NO2-contamuig atmospheres accelerating degradation at all humidity
levels between 40 and 80%
In the outdoor environment, the deposition of NO2 is limited, for most materials, by the
surface uptake resistance, and NO2 is more slowly adsorbed than SO2 In the experiments
conducted by Svedung et al (1983), Kucera (1986), and Johansson (1986), low flow rates
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were used in the exposure chambers During low-flow conditions, the deposition rate
becomes limited by the surface boundary layer resistance and the effective deposition rates of
NC>2 and SC>2 will become more nearly equal Thus, the conclusion from chamber studies
that NOX is synergistic with SO2 may not be applicable in outdoor environments In indoor
exposures of materials, however, the conclusions of Svedung et al, Kucera, and Johansson
are applicable
12.4.2 Effects of Nitrogen Oxides on Economically Important Metals
Steel
Steel is the most widely used structural metal and is available in a wide variety of types
with varying percentages of alloying elements Basically, steel consists of iron containing
0.02 to 1.5% carbon The corrosion behavior of common construction steels (carbon steels,
containing about 0.2% carbon) is similar, and rusting of exposed surfaces proceeds rapidly
Low-alloy steels, containing chromium, nickel, copper, molybdenum, phosphorus, and
vanadium in the range of a few percent or less for the total inclusion, are substantially
stronger and offer improved resistance to atmospheric corrosion Specialty steels, such as
stainless steels containing over 10% chromium, are designed to be highly corrosion-resistant,
but are also much more costly Bare steel is not usually exposed to the environment, but
rather is painted to prevent rust and premature failure Nevertheless, except where
specifically noted, the following discussion concerns common construction steel that is boldly
exposed with no coatings.
Samples of enameling steel were exposed at 57 of the National Air Surveillance
Network locations (Hayme and Upham, 1974), for 1- and 2-year exposure cycles Sulfur
dioxide and particulate matter concentrations, relative humidity, and paniculate chemistry
were monitored at the sites Corrosion rates for the steel samples, determined from weight
loss measurements, were correlated against the pollution measurements Hayme and Upham
(1974) concluded that either SO2 or particulate sulfate, or both, were significant in causing
steel corrosion. Particulate nitrate (PN) was not statistically significantly related to the
observed corrosion, however, their measurement techniques for PN were unreliable
Measurements of gaseous NOX species were not made
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Johansson (1986) showed in a low-flow chamber study that gaseous NO2 adsorbs on
steel surfaces and reacts with water to form HNO3 and HONO Construction steel was
3 3
exposed continuously for 6 weeks to 376 jttg/m or 5,640 /«g/m (0 2 or 3 0 ppm) NO2 and
different levels of moisture and SO2 He determined that the deposition rate of NO2 was
much lower than the deposition rate for SO2 and that steel exposed to NO2 alone, in the
absence of other pollutants, will slowly acquire a thin oxide layer (rust) that protects the
underlying steel from further damage Unfortunately, the nitrates formed during the
corrosion process are hygroscopic and act to adsorb further moisture from the atmosphere at
around 50% RH and above If it is also present, SO2, which does not form hygroscopic
corrosion products but does have a higher deposition rate than NO2 (Johansson, 1986), reacts
with this moisture to form strong acids that corrode the surface very rapidly In addition to
its hygroscopic effect, Johansson suggested that NO2 might increase the oxidation rate of SO2
to SO4=, and thus enhance corrosion At relative humidities in excess of 90%, the
synergistic effect of NO2 is lost because at these high Jiumidity levels moisture forms on the
surface whether or not NO2 is present In fact, Henriksen and Rode (1986) have suggested
that NO2 may actually inhibit SO2-rnduced steel corrosion at 95 % RH
Haynie (1986) analyzed data from 30-mo exposures of weathering steels at nine sites
around St Louis, MO, as part of the U S Environmental Protection Agency's Regional Air
Pollution Study Weathering steels are architectural steels specifically formulated to rapidly
develop a surface corrosion layer that protects the underlying substrate steel The exposure
samples were co-located with air quality monitoring stations Haynie (1986) statistically
analyzed the observed corrosion in relation to meteorological and air quality variables
He found that the sample weight change was positively correlated with the SO2 levels, but
negatively correlated with NO2 He concluded that NO2 decreases the solubility of the
corrosion layer
Haynie et al (1976) studied weathering steel in an exposure chamber. Although they
concluded that NO2 did not have as significant an impact as SO2 on the indicated corrosion,
a review of the data showed that at low relative humidities the samples showed somewhat
more damage at high NO2 concentrations (940 /tg/m3 [0 5 ppm]) than at low concentrations
(94 jig/m3 [0 05 ppm]).
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Galvanized Steel and Zinc
Because most carbon steels rust readily when exposed to moist air, a layer of zinc is
frequently coated or galvanized onto the surface The zinc acts to protect the substrate steel
electrochemically by preferentially corroding away, leaving the steel rust-free Zinc
galvanized steel is used for many outdoor purposes, including chain-link fences, highway
guard rails and sign posts, roofing, and automobile body panels
Whitbeck and Jones (1987) studied the accumulation of nitrates on galvanized steel in
an exposure chamber They exposed the galvanized steel to 18,800 /tg/m3 (10 ppm) of NO2
(much higher than ambient air levels) and measured the nitrate formation as a function of
time on the sample surface. They found that the formation of nitrates was linear with tune
Haynie et al (1976) included galvanized steel in their chamber study discussed above and
concluded that the effects of SO2 are much more significant than those of NO2
These results are further supported by the field investigations reported by Cramer et al
(1988). They found that SO2 is more readily absorbed on galvanized surfaces than NO and
NO2 and that SO2-induced corrosion probably dominates corrosion by NOX in most
environments In relatively dry environments, Cramer et al (citing Johansson, 1986) pointed
out that NO2 can participate in a reaction to oxidize SO2 and form H2SO4, which is very
aggressive to galvanized surfaces Edney et al (1987) statistically analyzed the results of
exposures of galvanized steel and chemical analyses of the runoff rainwater from the
samples. They found that the amount of deposited SO4= dominated the amount of deposited
NO3", and that SO4= and NO3" deposition rates were strongly correlated at the field exposure
site. From the regression analysis, therefore, SO4= was found to dominate the corrosion of
galvanized steel and NO3" was found not to be a significant contributor to corrosion at this
location Subsequent analysis of data from the same site by Spence et al (1988), using a
more complete regression model, found no statistically significant effects of pollution on
either galvanized steel or weathering steel exposed for 3 years The site used for this
experiment, Research Triangle Park, NC, is relatively rural and SO2 and NO2 concentrations
are fairly low. The analysis of Spence et al suggests that natural weathering processes
dominate over corrosion at this site
Although rarely used alone as a construction material, zinc is used for galvanizing and
as an alloying metal and its corrosion behavior has been investigated Johansson (1986)
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exposed zinc to NO2 and SO2 in a low-flow exposure chamber He showed that NO2 alone
had little impact, but that small amounts, 376 jug/m3 (0 2 ppm), were strongly synergistic
when combined with SO2 As the NO2 concentration m the mixture was increased from
376 fjLglm3 to 5,640 jwg/m3 (0 2 ppm to 3 0 ppm) and the SO2 concentrations were held
constant, there was little change in the rate of corrosion
Kucera (1986) has noted that, in the open air, zinc tends to form a layer of sulfates and
carbonates on the surface that acts to passivate the metal This layer is basic, and if rain
with a pH value of 4 or less washes the surface, the layer is removed, exposing the substrate
metal In this way zinc is sensitive to acid deposition, so that any pollutant, including NOX,
that adds to the acidity of the environment is damaging to zinc
Hermance (1966) and Hermance et al (1971) reported the impact of nitrates on zinc-
containing nickel-brass wire springs used in telephone relays They pointed out that
hygroscopic nitrate salts collected on the springs and moisture formed on the surface at any
relative humidity exceeding 50 % The nitrate deposition resulted in attack on the zinc in the
springs and premature failure of the relays In addition, Graedel and McGill (1986) have
pointed out that NO2 is known to be moderately aggressive towards nickel Ultimately, the
telephone companies were forced to replace zinc-containing nickel-brass springs in areas with
high NOX levels, such as Los Angeles Hennkson and Rode (1986) showed that at 95 % RH
the synergistic effects of NO2 and SO2 were not detectable for zinc corrosion At high
humidities, SO2 appears to dominate zinc corrosion
Aluminum
Aluminum is widely used because of its corrosion resistance and is second only to steel
in the amount of metal in use Aluminum is often exposed without coatings, such as paint,
and is used for architectural trim, aircraft, small buildings, cooking utensils, etc Kucera
(1986) noted that the tune of wetness of aluminum surfaces correlates with NOX
concentrations, but could not conclude that NOX was of any practical importance m the
aluminum corrosion process Johansson (1986) demonstrated in a chamber study that NO2
did not significantly adsorb on aluminum but that at 90% RH NO2 was synergistic with SO2
and caused nearly three tunes the corrosion caused by either pollutant alone Hennksen and
Rode (1986) showed that NO2 inhibits SO2-induced aluminum corrosion at 95 % RH In a
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chamber study, Loskutov et al (1982) demonstrated that the interaction of NO2 and water on
an aluminum surface was a complex process They concluded that adsorbed water acted to
displace NO2 on the surface, and that metal corrosion occurred simultaneously with the
adsorption/displacement process but slowed substantially as water displaced NOX
Vijayakumar et al (1989) exposed aluminum to 940 and 1,880 jtig/m (0 5 and 1 ppm)
NO2 m a chamber for 28 days They found no statistically significant impact of NO2 on
•5
aluminum. They also exposed aluminum to 252 and 1,260 /ig/m (0 1 and 0 5 ppm) HNO3
and determined that there was statistically significant damage and that the rate of the
damaging reaction was relatively rapid
Copper
Copper is used for architectural trim, electrical components, and heat transfer coils in
air conditioners. Chamber studies (Schubert, 1978, Rice et al, 1981) have shown that NO2
o
has little impact on copper at concentrations up to 2,444 jitg/m (1 3 ppm) Rice et al
(1980a) concluded from a multiple-city exposure study that hydrogen sulfide (H2S), SO2, and
O3 all had more impact than NOX on copper Kucera (1986), Johansson (1986), and
Hennksen and Rode (1986), using chamber studies, found that NO2 and SO2 in combination
was synergistic and increased the observed corrosion rate of copper by 10 to 20 times the
rate observed with, single-gas exposures under low-flow-rate conditions
Nickel
Nickel is used as a coating material to protect other metals from corrosion and is
particularly resistant to environments that aggressively attack steels, aluminums, and a
variety of other metals (e.g , marine environments) Rice et al (1980a) investigated the
indoor corrosion of nickel in several urban areas and found that SO2, NO2, and chlorides
played a significant role in accelerating nickel corrosion In a chamber study, Rice et al
(1980b) found that NO2 attacked nickel but that SO2 and chlorine (Cy were more aggressive
than NO2. Graedel and McGill (1986) have listed NO2 as being moderately aggressive
toward nickel
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12.4.3 Effects of Nitrogen Oxides on Electronics
Although the impact of air pollution on architectural and structural metals in the
outdoor environment has been recognized for some tune, the attack of NOX on electronic
components, generally used in indoor environments, is a more recently recognized problem
Telephone companies first reported the problem, noting failures of wire-spring relays in
telephone switching offices located in regions with high NOX levels (Hermance, 1966,
McKmney and Hermance, 1967, Hermance et al, 1971) Nitrogen oxides were depositing
on the springs and eventually leading to stress corrosion failures Here, the cost of the failed
part, the spring, was a minor consideration compared 1o the loss of service Eventually,
technology made the wire-spring relays obsolete, but, meanwhile, inconveniences and costs
were incurred as the result of these failures
Most of the gold used for industrial purposes is used to inhibit corrosion in electrical
contacts Svedung et al (1983) tested the corrosion resistance of gold-plated brass, one of
the most common contact materials, in an atmosphere containing 940 /ig/m3 (0 5 ppm) NO2
They found that NO2-contaimng environments were more aggressive than SO2 environments
at all relative humidities from 40 to 80% As found with common metals, an environment
containing a mixture of NO2 and SO2 was even more damaging Samples of gold contacts
exposed to mixed-gas atmospheres became partly covei ed by visible corrosion after 2 to 3 h
Kucera (1986) reported similar findings for electrolytic copper contacts Buildup of
corrosion layers on electrical contacts causes loss of conductivity and possible failure of the
contact
Voytko and Guilinger (1988) exposed gold, nickel, and palladium samples electroplated
on copper substrates to an atmosphere containing 100 ppb NO2, 100 ppb H2S, and 10 ppb
C12 at 60% RH for 332 h These samples were designed to simulate typical electrical contact
materials They found that all coatings developed "poies" that allowed the substrate copper
to corrode and that the "solderability" of the specimens generally decreased after exposure
Graedel and McGill (1986) reviewed the impact of pollutants on a variety of materials, and
listed NO2 as being moderately aggressive to solder
Abbott (1987) exposed electrical contacts made of cobalt-hardened gold over sulfamate-
mckel to different pollutant mixtures in a laboratory test environment He found that H2S
and SO2, both singly and in combination, were fairly benign to the contact surfaces,
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producing only mild pore corrosion even as concentrations approached 1 ppm The reaction
became more severe when NO2 was added to the mixture A mixture of 0 1 ppm each of
H2S, SO2, and NO2 was more aggressive than 0 5 ppm H2S plus 1 0 ppm SO2 Abbott also
estimated that approximately 30% of indoor electrical and electronic equipment environments
are corrosive enough to result in pore corrosion and film creep that could lead to component
failure.
Freitag et al (1980) investigated the corrosion of magnetic recording heads of the types
used in computers They found that exposure to 0 3 ppm each of NO2 and SO2 led to the
formation of corrosion products on the heads This corrosion would lead to a degradation of
the magnetic properties of the recording head
12.5 EFFECTS OF NITROGEN OXIDES ON PAINTS
Paints are by far the dominant class of manmade materials exposed to the atmosphere in
both indoor and outdoor environments Paint systems are used to protect substrate materials
such as wood, steel, and stucco from damaging environmental agents, including moisture,
sunlight, and pollutants Paints are also applied for aesthetic reasons Paints are broadly
classified as architectural coatings (e g , house paints, stains, varnishes), product coatings
(e.g , furniture finishes, automotive paints, appliance coatings), and special-purpose coatings
(e.g., bridge paints, swimming pool coatings, highway marking paint)
Although paints are designed to erode uniformly and repainting is expected, any
damaging process that exposes the substrate material or discolors the finish more rapidly than
natural weathering results in premature failure of the paint system and leads to the need for
more frequent maintenance and thus to increased costs Major paint manufacturers routinely
conduct proprietary tests of their coatings, and some information is available in the open
literature about the effects of NOX on selected paint systems Because paint formulations
vary widely, however, results obtained for one paint may not be directly applicable to other
paints
Spence et al (1975) investigated the effects of various pollutants on oil-based house
paint, vinyl coil coating, and acrylic coil coating A chamber study approach was used with
1,000 h of exposure to 94 and 940 /*g/m3 (0 05 and 0 5 ppm) NO2 in combination with
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various levels of SO2, O3, and humidity The coil coatings were very resistant to all
pollutants and showed little change over the course of the experiment The oil-based house
paint was found to be most sensitive to SO2 and humidity, but increased concentrations of
NO2 led to increased sample weights This implies that the NO2 was reacting with the paint
in some way, although whether this reaction was significant was not discussed
Hayme and Spence (1984) reported results of exposures of latex and oil extenor house
paints for 30 mo at nine sites around St Louis, MO They reported that NOX became
incorporated into the latex paint film and suggested that it reacted with the polymers that
make up the paint Similar results were reported for oil-based paint and brown staining
Vijayakumar et al (1989) exposed samples of high- and low-carbonate paints to NO2
and HNO3 for 28 days in an exposure chamber They found statistically significant damage
to low-carbonate paints at 940 jwg/m (0 5 ppm) NO2, but not at 1,880 jwg/m (1 ppm) NO2
The amount of damage was slight At 1,260 jwg/m3 (0 5 ppm) HNO3, however, both
carbonate and noncarbonate paints were damaged
12.6 EFFECTS OF NITROGEN OXIDES ON STONE AND
CONCRETE
Air pollution has been known to damage both budding and statuary stone Many
famous edifices, such as the Parthenon in Athens, have been the subject of studies of air
pollution-induced damage to building stone Calcareous stone, such as limestone, marble,
and carbonate cemented sandstone, is subject to air pollution attack Silicate stone, such as
granite, slate, and noncarbonate sandstone, is much less susceptible The effects of SO2
deposition on calcareous stone are well documented because calcium sulfate (gypsum) has
limited solubility and remains on protected stone surfaces as a gypsum coatmg Calcium
nitrate resulting from direct NOX attack is both very soluble and hygroscopic and thus washes
off the stone surface almost as soon as it is exposed to rain Livingston and Baer (1983)
suggest that the solubility of calcium nitrate has caused many researchers to overlook NOX
deposition to stone Thus, although few data are available, NOX may have a significant
effect on certain types of stone
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The interaction of NOX with building stone is complex Not only will nitrogen
compounds interact directly with the stone, but various endolithic bacteria present in the
stone result in biochemical interactions (Baumgaertner et al , 1990) Nitrosomonas species
oxidize ammonium to HONO and Nitrobacter species oxidize HONO to HNO3 Production
of these acids results in direct chemical attack to calcareous stone and concrete
Baumgaertner et al. (1990) have also reported that the surface of construction stone is a
significant source of NO, apparently biologically produced On the other hand, NO2 and
NH3 are absorbed by the stone
Baedecker et al (1990) summarized the work of several researchers for the National
Acid Precipitation Assessment Program (NAPAP) They noted that by far the greatest
chemical erosion of calcareous stone results from the natural constituents of clean rain
Carbon dioxide dissolved in ram forms carbonic acid that reacts with the calcium of the
stone. Baedecker et al (1990) estimated that wet-deposited hydrogen ions from all acid
species account for about 20 % of the chemical weathering of the NAPAP limestone and
marble samples. Dry deposition of SO2 was responsible for approximately 6 to 10% of the
chemical weathering and dry deposition of HNO3 (believed to be the major form of NOy
attack) accounted for 2 to 6% of chemical erosion They noted that an adequate model for
predicting dry deposition of HNO3 to stone is not available, and suggested that this topic
needs further research
Mansfeld (1980) performed a statistical analysis of damage incurred on marble samples
exposed for 30 mo at nine air quality monitoring sites around St Louis, MO He concluded
that NO3" and total suspended paniculate levels best correlated with observed stone
degradation, however, the analytical techniques used may be questionable and could have
resulted in inappropriate conclusions Livingston (1985) reviewed current studies regarding
the impact of NOX on calcareous stone He concluded that sulfates dominate the damage to
stone, but that NOX can play a role Livingston also showed that the reaction of stone with
SO2 is thermodynamicaUy favored over the reaction with NO2, and that if both pollutants are
present more calcium sulfate than calcium nitrate will be formed Amoroso and Fassuna
(1983) have suggested that the primary impact of NOX on stone may be its role in oxidizing
SO2 to form sulfate and eventually H2SO4 Although this is not a direct NOX attack, it does
lead to the degradation of stone
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Johansson et al (1988) exposed limestone, marble, and travertine to SO2 and NOX for
6 weeks at various concentration combinations in the parts-per-million and sub-parts-per-
million range The exposure chamber flow rates were low, with a net "wind speed" over the
samples of only 0 004 m/s The investigators found that significantly more gypsum
formation occurred with the combinations of pollutants than with either pollutant alone The
low flow rates in the chamber, however, make these data questionable for direct application
to outdoor exposures
Concrete is a widely used construction material aind dominates infrastructure
construction (bridges, highways, water and sewer systems) Webster and Kukacka (1985)
surveyed the construction industry and the technical literature for information regarding the
impact of pollutants on concrete and cement They speculate that HONO and HNQ3 are
more damaging than H2SO4 to concrete on brief exposures because they convert calcium
hydroxide to very soluble calcium nitrate They also believe that even highly diluted HNO3
solutions can bring about extensive destruction to concrete
12.7 EFFECTS OF NITROGEN OXIDES ON PAPER AND ARCHIVAL
MATERIALS
Paper is the primary storage medium for permanent records ranging from personal
photographs to the Constitution of the United States The National Research Council (1986)
noted that NO2 and other "acid gases" are expected to promote the failure of the cellulose
fibers that make up paper They recommended that the storage condition standards suggested
by the National Institute of Standards and Technology be followed and that NOX levels in
archives, libraries, and museums not exceed 5 j«g/m
Baer and Banks (1985) have pointed out a particular problem with NOX pollution that
libraries, museums, and archives face In the nineteenth century, cellulose nitrate was
produced in large quantities as the first plastic and was used in a wide variety of products
The common uses included photographic film, "acetate" recording disks, pre-vinyl imitation
leather, adhesives, and finishes As cellulose nitrate ages, it continuously emits NOX
If large quantities of books with artificial leather bindings (or replacement bindings using
pyroxylin-impregnated cloth) or of early photographic film are stored, NOX indoor emissions,
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which can be significant, may cause elevated concentrations unless the storage area is
adequately vented In extreme cases of nitrate film storage in sealed vaults with no
ventilation, the resulting gas pressure "may be enough to force out masonry walls "
If cellulose nitrate film is stored in sealed containers, NOX concentrations can build up to the
point of causing an autocatalytic reaction that can end in spontaneous combustion Several
collections of historic motion picture films have been destroyed in fires resulting from this
process
Salmon et al (1990) measured nitrogen species deposition during two seasons in five
museums in Los Angeles and measured outdoor concentrations of NOX species, as well
They noted that previous studies that attributed the damage to NO2 may have actually been
seeing damage induced by "co-pollutant" species, such as HNO3 Concentrations of HNO3
within the museums were in the range of 1 to 40% of the outdoor concentrations They
measured apparent HNO3 deposition velocities to vertical surfaces inside the museums, and
found values of approximately 0 18 to 2 37 cm/s They suggested that the deposition of total
inorganic nitrate (gas-phase plus aerosol-phase) onto vertical surfaces is dominated by gas-
phase species (probably HNO3 vapor) A further study of HNO3 removal by air-handling
systems was conducted at one museum, and Salmon et al (1990) found that approximately
40% of the HNO3 was removed by deposition within the ventilation system It was
suggested that very low measured values of HNO3 within galleries may be misleading
Deposition of HNO3 on surfaces within the museums, probably including the collection, was
rapid and potentially induced damage
Whitmore and Cass (1989), in the chamber study described in Section 12 2 1, tested a
selection of natural and synthetic artists' colorants applied to paper Nitric acid was carefully
removed from the chamber environment for these studies, and the NO2 concentration was
3
940 [ig/m (0 5 ppm) Seventeen natural organic colorants, 18 synthetic organic colorants,
and 7 inorganic colorants were tested in the absence of light for 12 weeks of exposure
Changes in color were measured with a spectrophotometer The paper itself exhibited slight
yellowing as the result of exposure, and several of the natural colorants showed noticeable
color changes For many of these samples, there was yellowing as measured by decreased
reflectance of blue light Four of the synthetic organic colorants and two of the inorganic
colorants showed measurable changes The authors noted that the cumulative NO2 dose to
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which the samples were exposed was roughly equivalent to 2 years of exposure in an
unprotected museum in downtown Los Angeles They concluded that the damage to a few of
the samples should be regarded as unacceptable
12.8 COSTS OF MATERIALS DAMAGE FROM NITROGEN OXIDES
Cost estimates for materials damage have been based on two distinct approaches
The first technique, the "top-down approach", involves determining the dollar value of a
material produced each year and then estimating the percentage of that value that is lost each
year from pollutant-induced damage The advantage of this approach is its ease of
application However, it is not rigorous and is likely to contain significant errors For
example, using the top-down approach, it is not possible to determine the pollutant exposure
levels of the materials because there is no way to determine the locations in which the
materials are deployed All that can be done is to use gross averages for exposures with this
technique
The second technique is the "bottom-up approach", in which as much detail as possible
is gathered regarding the geographic distribution of materials, the spatially resolved pollutant
concentrations and other variables, and the costs of repairs and replacement. The bottom-up
approach is more rigorous and demanding in terms of data requirements, and may yield a
closer estimate of actual costs than the top-down, production approach The accuracy of
either approach is unknown The methodology of cost estimation for materials damage needs
further research and development
The costs of some types of NOx-induced damage to textiles were estimated by the
National Research Council (1977) The following estimates, in 1977 dollars and based on
1977 production rates and pollutant concentrations, were made
$53 million incurred from dye fading on acetate fibers This
includes costs for more expensive, fade-resistant dyes, inhibitors,
research, quality control, fade losses at the manufacturing and retail
level, and reduced product life at the consumer level as the result of
fading
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2 $22 million incurred from dye fading on cotton fibers This includes
estimates of cotton fabrics exposed in polluted areas, percentages of
dyes known to be susceptible to NOX fading, and yearly loss in use-
life
3. $22 million incurred from dye fading on viscose rayon and rayon
blends with nylon, polyester, or acetate This includes reduced
wear-life for sensitive dye shades
Estimates of the costs of other types of losses caused by adverse NOX impacts on
textiles and fibers are not available Loss of strength and shortened use-life may be a
significant cost for fibers used for industrial purposes According to the National Research
Council (1977), 18 to 20% of all fibers produced are used by industry for items such as
tarpaulins, cords, and rope Loss of strength for fibers used for these purposes shortens use-
life and may present a safety hazard
Estimates of the costs of NOx-induced damage to plastics and elastomers are not
reported in the literature. The damages suffered through cross-linking and chain-scissioning
are loss of strength, increased cracking, and discoloration As the use of these compounds
for construction and automotive applications increases, the amount of exposure to NOX will
increase and the disbenefit costs of this exposure are expected to increase
No overall estimates of the costs of NOx-induced damage to metals and electronics are
available. For metallic corrosion in general, the costs are large The paint and coatings
industry, for example, produces a spectrum of products designed to prevent rust on steel and
these coatings would not be needed if corrosion were not a problem
Damage to paints, concrete, and stone produces potentially one of the largest economic
disbenefits of NOx-induced materials damage because the use of these materials is
widespread In 1987, sales by the paints and coatings industry alone approached $10 billion
The costs of infrastructure replacement because of concrete degradation can be seen as part
of the annual highway budgets Damage to historic stone structures and statues is mostly a
cultural cost and is not readily calculated
All of the foregoing cost estimates are either based upon old information (e g , the
National Research Council data were compiled in 1977) or are not specific for NOx-induced
damage. Also, the materials reported are only a subset of all materials exposed to NOX
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Recent and specific NOx-induced materials damage cost estimates are not available in the
literature This is an area of research that requires attention and updating
12.9 SUMMARY OF THE EFFECTS OF NITROGEN OXIDES ON
MATERIALS
Nitrogen oxides have been shown to cause or accelerate damage to manmade materials
exposed to the atmosphere Nitrogen oxides atmospheric and surface chemistry is complex
and there many compounds, including NO, NO2, and HNO3, that can contribute to this
damage
Strong evidence exists for the negative impact of NOX on dyes and fabrics Many
varieties of dyes are known to fade, become duller, or acquire a different cast, and white
fabrics may "yellow" from exposure to NOX Nitric oxide and NO2 were found to be
significant causes of color change for various fabric and dye combinations exposed in
ambient air Fade-resistant dyes and inhibitors have been developed, but are generally more
costly to employ Nitrogen oxides also attack textile fibers, resulting in a loss of strength
Nylon, in particular, appears to be susceptible to NO2 damage Plastics and elastomers are
subject to NO2 reactions that cause discoloration and changes in physical properties,
including loss of strength
Although NOX attacks metals, attack by SO2 is more aggressive, partly because in
outdoor environments the uptake of NO2 is limited by surface resistance and SO2 deposits
more rapidly There is evidence that HNO3 attacks aluminum, but that NO2 is not directly
damaging to aluminum Damage to metals from NOX can generally be discounted, except
perhaps in indoor exposures, where NO2 may react synergistically with SO2 Also largely
indoors, NOX is deposited on electronic components and magnetic recording equipment and
may lead to failures in these systems Nitrogen dioxide leads to pore corrosion on the gold-
plated surfaces of electrical contacts, leading to component failure
The influence of NOX on paints and stone has not been clearly demonstrated Many
researchers have reported that NOX and NOy (e g , HNO3) play a role in damaging these
materials, but most concede that SO2 and O3 are more directly damaging than NOX and NOy
in typical polluted atmospheres Nitrogen oxides, along with other "acid pollutants", attack
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the cellulose fibers in paper, leading to discoloration and weakened structure Nitrogen
dioxide has been shown to affect art supply colorants adversely and thus can damage works
of fine art
The highest NOX levels are to be found indoors where unvented combustion systems
(e.g , gas stoves) are used and the widest variety of materials are routinely exposed
Therefore, the principal effects of NOx-induced damage to materials are probably seen in the
indoor environment Few data are available regarding materials deterioration indoors
The presence of NOX will shorten the use-life of susceptible materials, and generally the
rate of damage is proportional to the pollutant concentration Adequate NOX damage
functions for a wide variety of materials are not available Consequently, practical
cost/benefit analyses of permissible NOX levels vis-a-vis shortened use-life estimates may be
impossible Cost estimates for NOx-specific damage at existing concentrations are available
only for dye fading ($97 million annually in 1977 dollars), and these estimates are very out
of date
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