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MARKERS OF AIR POLLUTION IN FORESTS: NUTRIENT CYCLING
Dale W. Johnson Helga Van Miegroet Wayne T. Swank
Environ. Sci. Divison Environ. Sci. Divison Coweeta Hydrologic Lab.
Oak Ridge National Lab. Oak Ridge National Lab. 999 Coweeta Lab Road
P.O. Box 2008 P.O. Box 2008 Otto, NC 28763
Oak Ridge, TN 37831-6038 Oak Ridge, TN 37831-6038
ABSTRACT
Air pollution may affect forest nutrient cycles in a number of
ways, but many of these effects are difficult to evaluate because
control sites unaffected by air pollution are seldom available for
comparison. Air pollution may alter either the nutrient content
or the flux rates in forest ecosystems, but, with a few exceptions
(e.g., SOD in foliage, Ala+ or Pb in tree rings), flux measurements
provide more sensitive measures of air pollution effects. The fact
that nutrient pools are typically quite large relative to the flux
rates into and out of them accounts for the sensitivity of flux
measurements. However, nutrient pools may be significantly
affected by air pollution over long periods of time, and the
importance of sample archiving as a means of detecting these
long-term changes in nutrient pools (e.g., soils) cannot be
overemphasized. Acid deposition and/or foliar damage by ozone
may cause increased foliar leaching, but the degree to which air
pollution affects foliar leaching is very difficult to assess in the
absence of control sites for comparison. Decomposition may also
be altered by increased acid input, trace metal deposition, or
gaseous pollution; but, once again, many of these potential changes
are very difficult to evaluate for lack of control sites. Acid
deposition nearly always influences soil leaching rates to some
degree. In the case of SOD, reasonable assumptions about
background concentrations often allow an approximate estimate of
the degree to which leaching is affected by sulfate deposition.
For NOR, there is considerably less certainty about exactly what
background levels should be. It is therefore more difficult to
evaluate the degree to which current NO3 leaching rates are
affected by atmospheric N deposition without a full analysis of the
N cycle of the ecosystem in question. The same basic
considerations apply to stream concentrations. Stream
concentrations may be less sensitive to air pollution effects than
soil solution concentrations, but the advantages of ease of
sampling and integration of whole watershed response make stream
concentration monitoring an attractive possibility as a marker of
air pollution in forest ecosystems.
INTRODUCTION
In general, nutrient fluxes (particularly solution fluxes such as foliar and soil
leaching) respond most readily to air pollution, whereas nutrient pools (e.g., soil,
133
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vegetation, and forest floor contents) are often sufficiently large to remain rather
insensitive to all but very long-term effects of air pollution inputs. Even in cases where
long-term changes in nutrient pools have occurred, they may be difficult to identify and
demonstrate if no control sites unaffected (or less affected) by air pollution are available
for comparison. Thus, with the exception of certain key nutrients or ions in certain
ecosystem components, nutrient fluxes usually represent more sensitive markers of the
presence of air pollution than do nutrient pools. The presence of air pollution by no
means implies that this pollution is creating any damage to the ecosystem, however.
From a nutrient cycling perspective, damage may be defined as either the depletion of a
critical nutrient or the elevation of any element to toxicity levels. Thus, while nutrient
or element pools are relatively insensitive markers of the presence of air pollution, they
may be the most sensitive markers of pollutant damage.
In the following discussion, we review very briefly the literature regarding effects
of air pollution on various nutrient fluxes and contents in forest ecosystems, attempt to
extract from that literature useful markers of air pollution on forest nutrient cycles, and
make some recommendations as to general sampling protocols that may be useful in
monitoring pollutant presence and effects on forest nutrient cycles.
DEPOSITION AND CANOPY INTERACTIONS
There has long been a concern over the effects of acid deposition on foliar
leaching. Studies involving artificial acid irrigation have variously found either no effect
on leaching (e.g., Haines et al. 1985) or an increase in foliar leaching (at very low pH)
(e.g., Wood and Bormann 1974~. Amthor (1986) argues convincingly that the energy costs
of increased uptake to compensate for accelerated foliar leaching are inconsequential.
However, there may be significant nutritional consequences if uptake does not compensate
for these losses. Rehfuess ( 1987) and Prinz ( 1985) advance a foliar leaching hypothesis
to explain the decline of high-elevation forests in Germany. They hypothesize that
foliage predamaged by exposure to photo-oxidants like ozone is especially susceptible to
increased Ca and Mg leaching. In soils low in exchangeable Ca and Mg, uptake
apparently does not compensate for this additional foliar leaching, and foliar Ca and Mg
concentrations decline, eventually to deficiency levels. These deficiencies, in turn, lead
to reduced frost hardiness, reduced chlorophyll content and photosynthesis rates, and
reduced root growth and uptake.
Unfortunately, the lack of control sites presents serious problems for testing the
foliar leaching hypothesis and for using foliar leaching as a marker of air pollution
stress. As noted by Lovett et al. ( 1986), the disappearance or H+ in the forest canopy
does not necessarily mean that foliar leaching has increased: the H+ may simply have
gone into a neutralization reaction with organic anions that naturally leach from the
canopy. This does not imply, however, that acid deposition has no effect on the
ecosystem: H+ in combination with organic acid would still result in throughfall with
greater total acidity than that in throughfall in an unpolluted environment. Low foliar
Mg could not serve as a reliable marker either because Mg deficiencies could clearly
occur in pristine as well as in polluted environments.
Tissue concentration of certain pollutants has obvious potential use as a marker of
air pollution. For instance, Bieberdorf et al. ( 1958) found that SOW in Pinus taeda
foliage was a very sensitive index of proximity to an SON source, presumably because of
SO2 absorption by the foliage. Such accumulation of SO6~ in the foliage will occur only
to the extent that trees are not S deficient. Kelly and Lambert ( 1972) indeed found
foliar SOW to be a reliable indicator of S excess (i.e., SOW accumulates only when the S
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supply is greater than tree S requirements). Foliar SOW may not be a sensitive index
of greatly excessive S deposition, either: foliar SOW may reach as high as 50% of total
foliar S. but apparently does not exceed this value even in areas of very high S
deposition (Johnson 1985~.
DECOMPOSITION, NUTRIENT MINERALIZATION, AND NITRIFICATION
Most studies of the effects of air pollution on decomposition have been conducted
under conditions of artificial irrigation, often with extremely high inputs over short time
periods, raising serious questions about the applicability of these results to actual
conditions. For example, would soil organisms be able to adapt to more acidic conditions
over longer periods with lower inputs? Moreover, results have been inconsistent between
sites and from one study to another. For instance, Like and Klein (1985) found that acid
treatments had no effect upon Vitrification but stimulated nitrogen mineralization in soil
columns from Camel's Hump, Vermont, whereas Strayer et al. (1981) found that acid
irrigation had a negative effect on Vitrification and either no effect or a stimulative
effect upon nitrogen mineralization of soils from Panther Lake in the Adirondacks in New
York. Klein et al. (1983) reported that the same treatment enhanced Vitrification in one
soil from the Adirondacks and inhibited it in two other soils from the same region.
Stroo and Alexander (1986) brought forward the suggestion that decreased N
mineralization does not necessarily represent a decline in N availability because part of
the N supply could be entering with acid precipitation.
Baath et al. ( 1980) found reduced decomposition rates (measured by the litterbag
technique) following acidification of forest soils in Sweden, whereas Hovland et al. (1980)
found that acid irrigation caused either increases or decreases in decomposition rates
(litterbag technique) depending upon the time of measurement or the amount of acid
applied. Similarly, Chang and Alexander (1984) showed that simulated acid rain applied to
three northeastern soils enhanced or inhibited decomposition depending on the amount of
acid applied and the amount of organic acid leached from soil samples. In another study,
sulfuric acid applications to hardwood leaf packs generally stimulated the loss of litter,
nutrients, and trace metals from the soil surface (Lee and Weber 1983~. Hay et al.
(1985) suggested that acidification of the mineral horizons of a po~zol soil inhibited the
transport of the major classes of soil organic components but increased the transport of
nitrogenous substances. Hovland ( 1981), Kelly and Strickland ( 1984), and Johnson and
Todd (1984) observed little or no effect of artificial acid irrigation upon CO2 evolution.
On the other hand, in one of the few studies that did not involve artificial irrigation,
Prescott and Parkinson (1985) found that decomposition was greatly reduced in sites near
a sour gas plant that had been emitting sulfur for over 20 years, compared to that in
sites farther from the plant.
In vitro studies by Moloney et al. ( 1983) on the degradation of spruce-fir litter
from New England showed that microbial CO2 evolution was reduced by acidic conditions,
and that it was further repressed by the presence of Pb and Zn but not by Al or Cu.
From these observations it was suggested that the acidity and heavy metals introduced by
polluted rain may adversely affect the metabolism of decomposing microflora.
The effect of trace metals on soil microorganism activity has been examined in a
number of laboratory and field studies. Generally, extremely high trace metal
concentrations, such as those found in the vicinity of smelters, decrease soil
microorganism activity (Zottl 1985~. Friedland et al. ( 1986) reported that levels of Pb,
Zn, and Ni below 103 mg/kg, levels of Cu below 1 o2 mg/kg, and levels of Cd below 120
mg/kg have no measurable effects on decomposition rates. In a long-term field study at
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Camels Hump, Vermont, Friedland et al. (1986) found substantial increases in trace metal
and organic matter concentrations of the forest floor over time. However, they
suggested that current metal concentrations are not capable of significantly reducing
decomposition of organic matter.
Hovland et al. (1980) summarize their results by stating: "The effect of the
artificial acid rain appears to be more pronounced on the leaching of metal elements than
on the biological activity and the dynamics of N and P." This statement seems
appropriate not only for their particular studies but also for the majority of the artificial
irrigation studies taken as a whole. It would seem that decomposition and nutrient
mineralization must certainly be affected by soil acidification at some point, but this may
come after the many decades of acid input usually necessary to change soil acidity. In
the initial stages of acid input, however, biological processes in the soil generally do not
appear to be as sensitive to acid inputs as chemical processes, namely, soil leaching.
SOIL LEACHING AND NUTRIENT EXPORT
The effects of acid deposition on soil leaching and stream water export have been
studied extensively, are reasonably well understood, and have been modeled (e.g., Reuss
and Johnson 1986~. These effects can be estimated rather easily, even in the absence of
control sites, given certain reasonable assumptions about background solution
concentrations of SOT and NO3. After the assumed background SOT concentrations are
subtracted, a comparison of the total cation or total anion concentration accounted for
by SON with that accounted for by HCO3 and organic acids usually provides an
approximation of the net effect of atmospheric S deposition on nutrient leaching from
soils. Exceptions to this principle, such as sites where significant S-bearing minerals are
present or sites where sulfur-bearing fertilizers have been previously applied, are usually
clearly defined. The situation with respect to atmospheric N deposition and soil solution
NOT, concentrations is more complex, because background NO3 can be quite high in
certain circumstances (Van Miegroet and Cole 1984, Foster 1985), and many different
types of ecosystem disturbances other than air pollution can cause an increase in soil
solution NO3 concentration. The effect of N input on soil leaching also depends on
whether NOT or NH$ enters the soil and on whether the latter form is subsequently
vitrified. Further complications and uncertainties arise in cases where organic acids are
the dominant natural leaching agents because organic acids are commonly estimated
through the calculation of an anion deficit (Johnson et al. 1977, Cronan et al. 1978~.
The effect of forest soil acidification on element leaching has been shown in soils
of southern Sweden (Tyler et al. 1987~. Measurements taken over decades show a decline
in soil pH. This decline has increased the solubility of elements such as Mg2+, Ala+,
Cd+, and Zn+, with consequent higher metal concentrations in soil water and increased
output of these metals.
The most reliable way to ascertain the effects of atmospheric S or N deposition on
soil solution SON and NO3 concentrations is through an analysis of the S and N cycles
of the ecosystems in question. Atmospheric S and N deposition in excess of biological
demands almost invariably results in an increase in SON and NOT above background or
natural levels. The SON and NOT can, in turn, leach from the soil if inorganic reactions
permit it. Inorganic interactions between NOT and soils are normally minimal, and, as a
general rule, atmospheric N input in excess of biological demands leads to increased NO3
and associated cation leaching. Fortunately, biological demands for N are often quite
high, and many forests are, in fact, N-limited (Cole and Rapp 1981~. There are
exceptions, however, in which very high rates of NO3 leaching occur as a direct
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137
consequence of high rates of atmospheric N deposition (Van Breemen et al. 1982) or as a
result of excessive N fixation (Van Miegroet and Cole 1984~.
On the other hand, biological demands for S are less than 5% of those for N (on a
molar basis) and are typically exceeded by atmospheric S inputs in polluted regions
(Johnson 1985~. Unlike NO3, however, SOW can adsorb to soils (such as many in the
southeastern United States) rich in Fe and Al oxides and poor in organic matter. This
adsorption can reduce the potential for leaching considerably up to the point where soil
adsorption sites equilibrate with incoming SOi~(Johnson 1985~.
Because solutions have to be electrically neutral, increases in SON or NO3 soil
solution concentrations lead to concurrent increases in total cation concentration. The
cation composition of the solution in equilibrium with a soil is controlled by both the
total anion solution concentration ant! the composition of the cation exchange complex.
These relationships are described fairly accurately by well-known selectivity equations
developed more than 50 years ago (e.g., Gapon 1933~. They predict, in essence, that 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 strength of the soil solution.
Reuss (1983) points out one very interesting aspect of these equations regarding the
effects of solution concentration changes: as total ionic concentration increases, the
ratios of trivalent to divalent to monovalent cations in solution also increase. In other
words, as the solution becomes more concentrated, cation concentrations increase in the
following descending order Ala+, Ca2+, Mg2+, K+, Na+, and H+. Thus, Ala+
concentrations in soil solution increase not only as the soil acidifies (i.e., as the
proportion of Ala+ on the exchange complex increases) but also as the total ionic
concentration of soil solution increases. These equations also imply that, compared with
leaching of the other major cation nutrients, K+ leaching from soils will be the least
affected by acid deposition.
These exchange equations further suggest that the H+ concentration increases to
some extent (though not as much as A13+, Ca2+, or Mg2+) with any increase in total
ionic concentration, even without direct HE input from the atmosphere. This increase in
H+ (decrease in pH) causes HCO3 and organic anions to protonate (becoming H2CO3 and
uncharged organic acids), leading to a decrease in the concentrations of these natural
anions. This replacement of natural, weak acid anions with atmospherically introduced,
strong acid anions (SOW and NO3) is referred to as an "anion shifty It results in a
lesser acceleration of total nutrient leaching than would occur with the simple addition
of SOW and NOD to the natural leaching of HCO3 and organic anions. Krug and Frink
(1983) argued that this anion shift can be such that no net increase in total cation
leaching occurs, even though the anion composition of soil solution changes dramatically.
This is an extremely unlikely, if not impossible, scenario in that it would require a HE
concentration increase equivalent to the addition of SOW and NO3. However, the
mitigative effect of the anion shift can be significant in acid soils, where H+ is a
significant cationic solution component.
CHANGES IN SOILS
Since soil nutrient pools are usually large relative to inputs and outputs of
nutrients, changes in soil content often occur very slowly and are very difficult to
measure. Nonetheless, changes in soils have been noted in certain cases. We found
evidence of modern SOW accumulation in soils near Oak Ridge, Tennessee, by sampling
beneath and adjacent to old houses (Johnson, et al., 1981~. Berden et al. ( 1987) provide
an excellent review of the literature documenting changes in soil acidity over time as
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138
well as supply a reasonable evaluation of the potential role of acid deposition in causing
these changes. Soil acidity per se is most emphatically not a good marker of air
pollution effects; soils can become extremely acid due to natural leaching and plant
uptake in unpolluted forests, under either intensive management (Turner and Kelly 1977)
or pristine conditions (e.g., Johnson et al. 1977, Ugolini et al. 1977~.
Although soils are rather insensitive to changes, we feel strongly that both long-
term monitoring of soil chemical properties and rigorous sampling and archival protocols
are highly desirable features of any program to monitor markers of air pollution. Had
such a program been implemented in the past, much of the current heated debate and
often unsupported speculation concerning the effects of air pollution on soils would not
exist. Furthermore, the existence of archived soil samples, with proper tests for sample
storage effects, would allow measurement and remeasurement of currently unmonitored
soil parameters that could be of importance in the future.
CHANGES IN STREAMS
Perhaps the most reliable integrates! marker of air pollution effects on forest
nutrient cycles is the long-term trend or pattern in stream chemistry. Watershed studies
have shown that alteration of solute concentrations in streams and net nutrient losses
provide a sensitive indicator of ecosystem stress (O'Neill et al. 1977, Likens et al. 1977,
Swank and Douglass 1977~. For example, long-term studies in streams draining hardwood
forests of the southern Appalachian Mountains show the beginning of delayed response to
atmospheric deposition as evidenced by increased SOi~ concentrations, reduced HCO3
concentrations, and changes in levels of other solutes (Swank and Wide 1988~. Shifts in
stream chemistry have also been related to forest insect infestations and forest
harvesting practices (Swank 1986~. The use of stream chemistry monitoring as an
indicator of air pollution in forests has several disadvantages: (1) a potential lag in
response (compared to other ecosystem compartments more directly exposed to
pollutants), (2) a lack of long-term records needed to establish trends, and (3) an
inability to distinguish cause-and-effect relationships (in the absence of process studies).
However, stream chemistry does integrate the spatial and temporal variability of forest
nutrient cycles and may provide the most reliable index of cumulative effects of air
pollution.
CONCLUSIONS
1. The most sensitive markers of the presence of air pollution are probably the
concentrations of certain ions such as SOT and NOR in natural waters. The
sensitivity of a given natural water to air pollution declines with the degree to
which the water is affected by biological and chemical cycling processes within the
ecosystem; thus, the following order of sensitivity is generally found: deposition
throughfall > soil solution > streamflow. With respect to the ease of monitoring,
the following order is found: streamflow > wet deposition > soil solution >
throughfall > dry deposition.
Certain key ion or element concentrations in particular ecosystem components (e.g.,
SOT in foliage, Ala+ in tree rings, lead in forest floor) may also be useful markers
of the presence of air pollution.
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139
3. The mere presence of air pollution in a forest ecosystem can by no means be taken
as evidence that air pollution is causing any significant damage to the ecosystem.
4. Nutrient pools are most likely to be affected by air pollution. However, due to
their relatively large size, they are rather insensitive markers.
5.
Sample archiving from nutrient pools of ecosystems with well-documented long-term
element budgets is essential for determining long-term changes in ecosystem
nutrient pools caused by air pollution as well as by other factors.
ACKNOWLEDGMENT
Research supported by the Electric Power Research Institute under Contract No. RP-2621
with Martin Marietta Energy Systems, Inc., under Contract No. DE-AC05-840R21400 with
the U.S. Department of Energy. Publication No. 3133, Environmental Sciences Division.
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Representative terms from entire chapter:
acid rain
