ECOTOXICOLOGY: A Comprehensive Treatment - Chapter 31 pptx

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31
Descriptive Approaches
for Assessing Ecosystem
Responses to
Contaminants
31.1 INTRODUCTION
Now that we have an appreciation of the important processes that characterize ecosystems and
the general approaches used to quantify these processes, we will turn our attention to the primary
objective of this section.Aswithcommunity-levelassessments, ecotoxicologists interested in ecosys-
tem responses to anthropogenic stressors employ descriptive, quasi-experimental, and experimental
approaches. In the following two chapters, we will explore the use of these observational and
experimental studies to link changes in primary and secondary production, nutrient cycling, and
decomposition to contaminants and other anthropogenic stressors. In a separate chapter, we will con-
sider effects of globally distributed and atmospheric stressors (e.g., acidification, NO
x
deposition,
elevated CO
2
, and UV radiation) on these ecosystem processes.
Investigations of ecosystem processes may be conducted across a range of spatial and temporal
scales. Functional measures such as community metabolism or nutrient transport can be measured in
isolated soil microbial systems or in whole forests or watersheds. However, as we move up the hier-
archy of biological organization from individuals → populations → communities → ecosystems,
we generally increase the spatial and temporal scales of our investigations. Because many experi-
mental studies of ecosystem processes are often limited in spatial and temporal scale, descriptive
approaches can provide very compellingand ecologically realistic results. As discussed in Chapter 23
for communities, the typical trade-off is that observational or correlative investigations only provide
a catalog of potential causal explanations. A more powerful case for causation in descriptive stud-
ies can be established by the application of strong inference (Platt 1964), other formal inferential
methods such as stressor identification (Suter et al. 2002), or Bayesian inferential techniques.

The initial definition of ecological integrity proposed by Karr (1991) included both structural
and functional measures, and most ecologists would agree that assessing effects of anthropogenic
stressors on ecosystems requires adequate characterization of both patterns and processes. The effic-
acy of using functional measures to assess ecosystem responses to contaminants has received limited
attention. As a consequence, development of functional criteria as indicators of ecological integrity
has lagged behind more traditional approaches based on community structure (Bunn and Davies
2000, Gessner and Chauvet 2002, Hill et al. 1997). Kersting (1994) provides an excellent review
of literature on the use of functional endpoints in freshwater field tests for hazard assessment of
chemicals. Some assessments of ecological integrity measure patterns of community composition as
a surrogate for ecosystem processes (Bunn and Davies 2000); however, patterns and processes are
not necessarily related in some instances, especially in systems where disturbance is relatively weak.
For example, Bunn and Davies (2000) measured stream metabolism at seven sites in southwestern
Australia and relatedecosystemprocessestocommunitystructure.Althoughchangesingrossprimary
production (GPP) and respiration were related to water quality, there was no relationship between
water quality and macroinvertebrate community structure.
665
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666 Ecotoxicology: A Comprehensive Treatment
The characterization of ecological integrity based exclusively on structural measures is incon-
sistent with how most ecologists view ecosystems (Gessner and Chauvet 2002). We believe that
restricting our analyses to mainly structural measures has provided a somewhat incomplete picture
of how ecosystems respond to and recover from anthropogenic disturbances. Issues such as relative
sensitivity, response variability, and functional redundancy have been considered when comparing
the usefulness of structural and functional measures (Howarth 1991, Schindler 1987, 1988). Despite
concern that changes in some ecosystem processes occur only after compositional changes and are
therefore less useful, alterations in material cycling and energy flow are such fundamental properties
of ecosystems that they should be included in ecological assessments. Leland and Carter (1985)
argue that some functional processes in ecosystems are easier to quantify than relationships between
abundance and environmental variables. These functional measures also integrate general charac-

teristics of diverse communities, thus facilitating comparisons among different ecosystems. Given
the recent interest in making comparisons across relatively broad geographic regions, functional
measures may be more useful than structural measures because they are not dependent on specific
taxa that are often restricted to a specific region. Finally, because causal mechanisms that control
ecosystem processes are generally well understood, restoration strategies may be more obvious when
based on functional measures (Bunn and Davies 2000).
Although papers that report relative sensitivity of community metrics to contaminants are com-
mon in the literature (Carlisle and Clements 1999, Kilgour et al. 2004), surprisingly few studies
have compared responses across levels of biological organization (Adams et al. 2002, Bendell-
Young et al. 2000, Cottingham and Carpenter 1998, Niemi et al. 1993, Sheehan 1984, Sheehan
and Knight 1985). There is also the perception that quantifying ecosystem responses is logistically
challenging compared to structural measures (Crossey and La Point 1988), an idea that has not
been rigorously examined in the literature. Thus, we believe that it is premature to conclude that
ecosystem processes are less sensitive or less reliable indicators of stress. In fact, some studies have
reported that changes in ecosystem processes may occur in the absence of alterations in community
structure (Bunn and Davies 2000). Niemi et al. (1993) reported that functional measures such as
GPP were more sensitive indicators of recovery than structural measures. Similarly, Clements (2004)
observed that community respiration was generally more sensitive to heavy metals than common
structural measures such as abundance and species richness. Because alterations in community struc-
ture are not necessarily related to ecosystem processes, we view these as complementary measures
for assessing ecological integrity. More important, simultaneous assessment of pattern and process
can provide insight into the mechanistic linkages between stressors and responses. Studies by Wal-
lace and colleagues (Wallace et al. 1996) provide some of the best examples demonstrating how
contaminant-induced alterations in structural characteristics (e.g., elimination of macroinvertebrate
shredders) directly influence ecosystem processes (e.g., litter decomposition and export). There is
also some evidence thatfunctional measures may be moredirectlyrelated to specific typesofstressors
(Gessner and Chauvet 2002).
Although many different processes could be used to assess ecosystem integrity, we will focus in
this section on three functional measures: ecosystem metabolism (respiration, primary and second-
ary production), litter decomposition, and nutrient cycling. As described in Chapter 30, a significant

amount of background information characterizing these processes is available, although the level of
development varies among ecosystem types. For example, lake ecologists have historically relied
on functional measures, especially primary production, whereas lotic ecologists have tended to
rely on structural measures (Gessner and Chauvet 2002). These differences have resulted in diver-
gent approaches used to assess effects of contaminants in aquatic ecosystems. Similarly, studies of
biogeochemical processes in terrestrialhabitats, especiallyinagricultural systems, focus primarily on
factors that increase primary production, whereas aquatic ecologists have been more concerned with
understanding factors that limit production as a way to control eutrophication (Grimm et al. 2003).
The methodological approaches used to assess effects of contaminants on ecosystem metabolism are
consequently different in aquatic and terrestrial ecosystems.
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 667
31.2 DESCRIPTIVE APPROACHES IN AQUATIC
ECOSYSTEMS
31.2.1 E
COSYSTEM METABOLISM AND PRIMARY PRODUCTION
Energy flow and metabolism are fundamental properties of ecosystems that are also closely related
to the transport of contaminants. Many of the same physical, chemical, and biological processes
that influence the flow of energy between biotic and abiotic compartments also regulate the fate of
chemicals. Primary production in aquatic ecosystems is particularly sensitive to many anthropogenic
stressors. The effects of nutrient subsidies and input of organic materials on productivity have
received considerable attention in streams, lakes, and marine ecosystems. In their description of
subsidy-stress gradients, Odum et al. (1979) contrast the ecosystem-level effects of “utilizable”
inputs such as nutrients and organic materials with toxic materials (Chapter 25). This analysis
contrasts the role of nutrients such as N and P as both regulators of ecosystem production as well as
stressors when threshold levels are exceeded.
Input of nutrients associated with agricultural, domestic, industrial, and atmospheric sources are
widely regarded as major stressors of aquatic ecosystems (National Research Council 1992). Whole
ecosystem nutrient budgets calculated for several ecosystems reveal that inputs often exceed outputs,

resulting in large amounts of nutrients being stored in a watershed (Bennett et al. 1999, Jowarski et al.
1992, Lowrance et al. 1985). Bennett et al. (1999) used amass-balance approach to estimate P storage
based on inputs and outputs in the Lake Mendota (Wisconsin, USA) watershed. They reported that
approximately 50% of the P entering the watershed was retained and could be readily mobilized by
climatic, geologic, or hydrologic events. These increased nutrient levels in aquatic ecosystems are
often associated with toxic algal blooms, increased plant growth, oxygen depletion, fish kills, and
major shifts in community composition. Land-based inputs of nutrients also increase eutrophication
and have negative effects on primary productionofmacrophytes in coastal areas. Using data compiled
from an extensive literature survey, Valiela and Cole (2002) reported a strong inverse relationship
between N loading and primary production of seagrass meadows in coastal marine areas. The percent
of seagrass cover lost reached 100% as N loading approached 100 Kg N/ha/y. These effects resulted
from reduced light supply associated with increased phytoplankton production. The damaging effects
of N enrichment were significantly reduced in areas protected by salt marshes and mangroves.
As described in Chapter 30, availability of N and P can directly regulate primary production
and biomass accrual in aquatic ecosystems (Biggs 2000). However, the direct effects of nutrients
on primary production are complex and may be mediated by other factors such as hydrologic char-
acteristics and abundance of grazers. Riseng et al. (2004) used covariance structure analysis (CVA)
to examine effects of hydrologic regime and nutrients in 97 midwestern U.S. streams. Increased
nutrients in streams with high hydrologic variability resulted in greater algal abundance because
grazers were reduced. In contrast, in more stable streams where grazers were abundant, algal pro-
duction was limited and the net effect was an increase in herbivore production. Because hydrologic
characteristics of a watershed are dependent on watershed physiography and climate, these factors
may ultimately control responses to nutrient additions (Riseng et al. 2004).
Alterations in primary productivity and respiration have been measured in response to chemical
stressors other than nutrients in aquatic ecosystems. Crossey and La Point (1988) could not detect
differences in GPP between metal-impacted and reference sites, but when data were normalized to
algal biomass (as chlorophyll a), GPP was higher at the reference site. Hill et al. (1997) measured
variance and sensitivity of several functional measures in the Eagle River, a Colorado (USA) stream
impacted by metals. Results showed that measures of community metabolism (GPP, NPP, and res-
piration) were lower at stations located downstream from heavy metal inputs (Figure 31.1). These

functional measures were correlated with mortality of Ceriodaphnia dubia, and inhibition concen-
trations (IC50 values) for respiration were comparable to LC50 values derived from these more
traditional toxicological approaches. These results suggest that functional measures were about
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668 Ecotoxicology: A Comprehensive Treatment
Station
ER-1r ER-3 ER-5 ER-12 ER-12a
0
2
4
6
8
10
12
14
Zn concentration (µg/L)
0
50
100
150
200
250
300
GPP
Zn concentration
GPP (g O
2
/m
2

/day)
FIGURE 31.1 Effects of Zn on community respiration measured at different stations in the Eagle River,
Colorado, United States. (Data from Tables 2 and 3 in Hill et al. (1997).)
as sensitive as acute toxicity for quantifying effects of heavy metals. Clements (2004) compared
structural (species richness and abundance of metal-sensitive mayflies) and functional (community
respiration) responses of benthic macroinvertebrate communities to a mixture of Cd and Zn in stream
microcosms (Figure 31.2). Both structuraland functional measures weresignificantly related to metal
concentration, but effects on community respiration were generally greater than effects on species
richness or abundance of metal-sensitive mayflies.
Unlike studies conducted with diatoms and attached algae, research investigating effects of
contaminants on emergent macrophytes has shown that primary production and photosynthesis of
these groups are relatively insensitive. Bendell-Young et al. (2000) compared the response of several
structural and functional endpoints (mutagenic responses, morphological deformities, mortality,
community structure, and plant productivity) measured in wetlands receiving oil sands effluents.
Photosynthetic rates of cattails (Typha latifolia L.), measured as CO
2
uptake, were actually greater in
wetlands receiving processed water from oil sands, a response that contradicted expectations. These
researchers concluded that structural changes in benthic communities and blood chemistry of fish
were more sensitive indicators of stress than functional measures. Photosynthetic rate of salt marsh
plants (Spartina alterniflora) was measured at reference and contaminated sites in the southeastern
United States (Wall et al. 2001). Although significant negative effects on benthic detritivores were
observed at a site heavily contaminated by mercury and PCBs, photosynthesis of Spartina was not
affected.
31.2.2 SECONDARY PRODUCTION
In addition to direct effects on primary producers, contaminants and other stressors may alter the
amount and rate of energy flow to higher trophic levels. The utilization of available energy in an
ecosystem is thus an important measure of ecological integrity. Perhaps the most common functional
response related to energetics measured in aquaticecosystems is the abundance ofdifferent functional
feeding groups (Rawer-Jost et al. 2000, Wallace et al. 1996). In part, the utility of functional feeding

groups as a metric in ecological assessments is based on the assumption that specialist feeders such as
scrapers and shredders are more sensitive to contaminants than generalist feeders such as collector-
gatherers and filterers (Barbour et al. 1996). Although data on functional feeding groups are generally
presented as abundance or density per unit area, and therefore not strictly a functional measure,
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 669
Total number of species
16
18
20
22
24
CCU
010 20304050
Community respiration
0.0
0.5
1.0
1.5
Total number of heptageniidae
200
250
300
FIGURE 31.2 Relationship between structural (total number of species; abundance of metal-sensitive hepta-
geniid mayflies) and functional (community respiration) endpoints and heavy metal (Cd and Zn) concentration
in stream microcosms. Heavy metal concentration was expressed as the cumulative criterion unit (CCU),defined
as the ratio of the measured metal concentration to the hardness adjusted chronic criterion values for Cd and Zn.
(Data from Clements (2004).)
the assumption isthatcompositionofdifferent feeding groups reflects important ecosystemprocesses.

For example, abundance of grazers, organisms that feed directly on periphyton and algae, is related
to primary productivity in streams. Similarly, abundance of shredders, organisms that process leaf
litter, regulates downstream transport of coarse particulate organic material (Wallace et al. 1982).
Secondary production, which we have defined as the production of heterotrophic organisms,
has been used to document effects of several stressors in aquatic ecosystems, including hydrologic
modification (Raddum and Fjellheim 1993), pesticides (Whiles and Wallace 1995), urbanization
(Shieh et al. 2002), and heavy metals (Carlisle and Clements 2003). Because secondary production
integrates individual growth rates and population dynamics, it captures in a single measure several
important aspects of energy flow through ecosystems.Although integration of these measures across
levels of biological organization is a laudable goal in ecosystem ecotoxicology, measures of second-
ary production are rarely included in biological assessments. Even measuring secondary production
of individual species is highly labor intensive because it requires sampling populations with sufficient
frequency to quantify individual growth rates, mortality, immigration, and emigration. The logistical
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670 Ecotoxicology: A Comprehensive Treatment
challenges associated with measuring secondary production will likely deter some researchers from
using this endpoint in ecological assessments. Indeed, France (1996) argues that because secondary
production is dependent on abundance, little additional information is gained by including these
more labor-intensive approaches. However, because secondary production is a composite of indi-
vidual mortality, growth rate, population abundance, and biomass, it represents a potential holistic
indicator of ecosystem bioenergetics that is not reflected in these individual measures.
Because most studies of secondaryproduction are based on detailed analysis of individual species
or groups of related species, our understanding of energeticsfrom the perspective of entire ecosystems
is somewhat limited (Shieh et al. 2002). To be useful as an ecosystem-level indicator, a measure of
secondary production should include a significant number of dominant species in an ecosystem and
should also be combined with data on trophic interactions. Sheehan and Knight (1985) compared
patterns of community composition and secondary production at several sites along a gradient of
metal contamination. Chironomids dominated benthic communities at metal-polluted sites, a finding
commonly reported in the literature. However, despite large shifts in community composition among

sites, relatively little difference in secondary production of chironomids was observed.
Another challenge associated with using secondary production as an indicator of ecosystem
integrity is that production may either increase or decrease, depending on the nature of the stressor.
The theoretical basis for the difference in responses between subsidizing and toxic stressors was
first described by Odum et al. (1979), but there have been relatively few empirical studies docu-
menting this pattern. Shieh et al. (2002) estimated energy flow based on secondary production and
trophic interactions at polluted and reference sites in a Colorado stream receiving urban discharges.
Secondary production, which was primarily supported by detritus, increased by more than two times
at the most impacted site due to the input of nutrients and organic materials (Figure 31.3a). In contrast
to these patterns, Carlisle and Clements (2003) reported a decline in secondary production along a
gradient of heavy metal contamination (Figure 31.3b). Differences in production among these streams
were primarily a result of lower population abundance of metal-sensitive species, especially grazing
mayflies. The large reduction in secondary production of herbivores likely had important cascading
effects on trophic interactions and energy flow through this ecosystem. Results of these studies are
consistent with predictions of the subsidy-stress hypothesis (Odum et al. 1979) and illustrate the
contrasting effects of subsidizing materials and toxic chemicals on ecosystem processes.
Reduced secondary production of zooplankton in lake ecosystems may result from increased
mortality, lower growth rates, and/or shifts in size composition of dominant species. Hanazato
(2001) reviewed effects of pesticides on zooplankton across levels of organization, from individuals
to ecosystem-level responses. Ageneral trend observed in lakes receiving pesticides was a reduction
in mean body size of zooplankton as a result of differential sensitivity among species. Hanazato
(2001) speculated that one potential ecosystem-level consequence of altered size distributions was a
reduction in the amount of energy transferred from primary producers to higher trophic levels. This
reduced transfer efficiency was associated with a variety of anthropogenic stressors, including heavy
metals, acidification, and nutrient enrichment.
31.2.3 DECOMPOSITION
Litter decomposition is a fundamental ecological process that has been studied extensively, espe-
cially in lotic ecosystems (Chapter 30). It integrates responses of a variety of biota, from bacteria
and fungi to shredding macroinvertebrates (Niyogi et al. 2001, 2003). There is a large database
available reporting decomposition rates of leaves and quantifying biotic and abiotic factors that

influence litter decay under a variety of environmental conditions. Gessner and Chauvet (2002)
provided an excellent and comprehensive review of numerous studies that used litter breakdown to
quantify effects of physical and chemical stressors in streams. They make a compelling argument
for the use of decomposition as an ecosystem indicator and provide specific criteria for assessing
ecological integrity. Breakdown rate coefficients (k), measured by regressing remaining mass of
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 671
Reference
Intermed N
High N
Production (g AFDM/m
2
/year)
0
10
20
30
40
50
60
70
0.0
0.5
1.0
1.5
2° Production
Nitrogen
Station
Reference

Low Zn
Intermed Zn
High Zn
Production (g DW/m
2
/year)
Zn (µg/L) NH
4
(mg/L)
0
1
2
3
4
5
6
0
50
100
150
200
250
300
2° Production
Zinc
(a)
(b)
FIGURE 31.3 Contrasting effects of nutrients (a) and heavy metals (b) on invertebrate secondary production.
(Effects of nutrients from Tables 1 and 4 in Shieh et al. (2002). Effects of Zn from Tables 1 and 2 in Carlisle
and Clements (2003).)

litter against time, are generally reduced in disturbed ecosystems. Effects of contaminants on litter
decay may result either from alterations in microbial processes or reduced abundance of macroin-
vertebrate shredders (Figure 31.4). It is also necessary to distinguish effects of contaminants on
biological processes, such as the elimination of shredders or reduced microbial activity, from effects
due to physical characteristics of the system. Methodological approaches that exclude or include
different groups of organisms can be used to separate the relative importance of these processes, thus
allowing ecologists to isolate underlying mechanisms. Because of the diversity of approaches used
to quantify litter decay and the large number of environmental factors that influence k, development
of standardized techniques for assessing effects of contaminants is essential.
The most comprehensive applications of leaf litter methodologies to investigate effects of con-
taminants have been conducted in metal-polluted and acidified streams. Schulthesis and Hendricks
(1999) and Schulthesis et al. (1999) measured macroinvertebrate community composition and leaf
decomposition at sites upstream and downstream from an abandoned pyrite mine in southwest-
ern Virginia (USA). Shredder abundance was greater and decomposition rates were 1.4–2.7 times
faster at the reference site compared Cu-polluted sites. Remediation activities initiated during the
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672 Ecotoxicology: A Comprehensive Treatment
Physicochemical
characteristics
Microbial
processes
Contaminants and
other stressors
Shredder biomass
and production
Leaf
decomposition
FIGURE 31.4 Conceptual model showing the effects of contaminants and physicochemical characteristics
on microbial processes, shredder biomass, and leaf litter decomposition.

TABLE 31.1
The Influence of Aqueous Zn Concentration, Metal Oxide
Deposition, and Nutrient Concentrations on Structural
and Functional Endpoints Measured at 27 Stream Sites in
the Rocky Mountains, Colorado (USA)
Variable Independent Variables R
2
P-Value
Leaf breakdown rate (k) [Zn], oxide deposition .72 .0001
Shredder biomass [Zn], oxide deposition .64 .0001
Microbial respiration Oxide deposition, nutrients .54 .0001
Data from Table 2 in Niyogi et al. (2001).
study period allowed these researchers to compare recovery of structural and functional responses.
Although community composition and abundance of shredders increased following improvements
in water quality, the rate of leaf processing did not increase as expected, suggesting some resid-
ual effects of Cu on microbial processes. In contrast to these studies, Nelson (2000) reported little
effects of Zn contamination on decomposition rates of aspen (Populus tremuloides) in a Colorado
Rocky Mountain stream, despite significant changes in community composition. These research-
ers speculated that the lack of a response in their study resulted from the relative insensitivity of
microbes, especially fungi, to the moderate levels of Zn contamination.Alternatively, because micro-
bial activity is limited by cold temperatures in Rocky Mountain streams, leaf processing may be more
dependent on invertebrate shredders (Niyogi et al. 2001), which were unaffected by Zn in this study
(Nelson 2000).
Comparative studies in streams across a gradient of heavy metal pollution provide the best oppor-
tunity to quantify effects of stressors relative to other factors that regulate leaf decomposition. Niyogi
et al. (2001) measured decomposition rates at 27 stream sites (8 reference and 19 metal-polluted) in
the Rocky Mountains of Colorado, USA. In addition to its broad spatial scale, this study is unique
because researchers quantified the relative influence of several stressors associated with mining
pollution, including acidification, elevated Zn concentration, and metal oxide deposition. Litter
decay coefficients (k) and shredder biomass decreased with increasing aqueous Zn concentration

and deposition of metal oxides (Table 31.1). In contrast, microbial respiration was more influenced
by metal oxide deposition and nutrients. Because decay coefficients were more closely related to
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 673
shredder biomass than microbial respiration, results of this study suggest that macroinvertebrates
were more important than microbial processes in regulating leaf litter processing in streams (Niyogi
et al. 2001). Carlisle and Clements (2005) related leaf decomposition to shredder secondary produc-
tion and microbial respiration in reference and metal-polluted streams in Colorado, USA. Because
leaf decomposition was measured as a function of shredder secondary production instead of shredder
biomass, this study provided a unique opportunity to quantify effects of stressors on energy flow
through an allochthonous food web. Results showed that shredders disproportionately contributed
to leaf litter decay, and that species-specific differences in sensitivity to metals among shredders
helped explain differences among streams.
Stream acidification by atmospheric deposition or other sources can have direct effects on litter
decomposition (Dangles et al. 2004, Griffith and Perry 1993, Tuchman 1993, Webster and Benfield
1986). Griffith and Perry (1993) attributed slower processing rate of litter in acidic streams to
lower biomass of shredders. In contrast, differences in community composition of shredders were
primarily responsible for differences in processing rates between neutral and more alkaline streams.
Tuchman (1993) reported that declines in invertebrate shredders in acidified lakes were correlated
with decreased litter breakdown rates. As with studies of metal-polluted streams, the most convin-
cing evidence demonstrating a relationship between acidification and leaf decomposition has been
obtained from spatially extensive surveys. Dangles et al. (2004) measured microbial respiration,
litter decay, and shredder abundance and composition in 25 streams along a gradient of acidification
in the Vosges Mountains, France. Breakdown rates varied 20-fold between acidified and neutral
streams, with alkalinity and aluminum concentration explaining 88% of the variation. Reduced leaf
decomposition in acidified streams was related to lower abundance and biomass of the amphipod,
Gammarus fossarum, a functionally important and acid-sensitive species. The greater breakdown
rate observed in coarse mesh bags (5.0 mm), which allowed shredder colonization, compared to fine
mesh bags (0.3 mm), which excluded shredders, supported the hypothesis that microbial processes

were relatively unimportant in this investigation (Dangles and Guerold 2001).
Although pesticides and other organic contaminants are likely to have significant effects on
litter decomposition by altering microbial processes and shredder communities, these stressors have
received considerably less attention than heavy metals and acidification (Gessner and Chauvet 2002).
Delorenzo et al. (2001) provided a comprehensive review of the effects of pesticides on microbial
processes related to decomposition. The best examples showing the effects of organic chemicals
on shredder biomass and subsequent alterations in leaf processing involve long-term experimental
studies (Whiles and Wallace 1992, 1995) and stream mesocosm experiments (Stout and Cooper
1983), which will be described in Chapter 32. Swift et al. (1988) examined effects of dimilin,
an insect growth regulator used for control of gypsy moths, on litter decomposition. Although
laboratory bioassays with shredders showed significant mortality when shredders were fed dimilin-
treated leaves, decomposition rates of treated leaves in the field were actually greater than controls.
The faster processing rate of treated leaves was attributed to the potential carbon source that dimilin
provided for bacteria.
Most investigations of leaf litter decomposition report that decay coefficients are reduced in
contaminated streams. However, stressors that subsidize an ecosystem (e.g., nutrients or organic
materials) may have the opposite effect. Niyogi et al. (2003) measured breakdown of tussock
grass (Chionocloa rigida) in 12 New Zealand streams along a gradient of agricultural develop-
ment. Nutrients (N and P), the predominant stressors in this system, increased along this gradient
and stimulated microbial respiration, invertebrate abundance, and the rate of litter decomposition. In
contrast, the macroinvertebrate community index (MCI), a biotic index of organic pollution, showed
increased stress along this same gradient. Similar findings were reported by Pascoal et al. (2001)
for a stream in Portugal receiving elevated nutrients. Despite reductions in abundance of shredders
at polluted sites, leaf breakdown rates were greater. These results serve to illustrate the importance
of understanding mechanistic linkages among stressors, microbial processes, and macroinvertebrate
community composition when using leaf decomposition to assess ecological integrity.
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674 Ecotoxicology: A Comprehensive Treatment
In addition to the direct effects of contaminants on the rate of litter decay, the concentrations

of toxic chemicals may increase in decomposing of plant material, thus providing a direct link to
detritus-based food chains. Windham et al. (2004) measured reduced decomposition of marsh grass
(Spartina alterniflora) in a metal-contaminated marsh as compared to a reference site. The 10–
100 times increase in metal concentrations in decomposing litter was attributed to adsorption and
microbial processes.
31.2.4 NUTRIENT CYCLING
The majority of studies investigating effects of contaminants on nutrient cycling in aquatic ecosys-
tems have focused on nitrification, denitrification, and other processes associated with N flux (Kemp
and Dodds 2002, Kemp et al. 1990, Royer et al. 2004). Most of this research has been conducted
within the context of understanding effects of nutrient enrichment, especially N and P, on freshwater
and estuarine ecosystems. Eutrophication, caused by the release of excess nutrients, is regarded as
the major threat to freshwater and coastal ecosystems in the United States (U.S. EPA1990). Greater
than 50% of the impaired lake area and river reaches in the United States result from excess nutri-
ents. Most of this impairment is associated with nonpoint source inputs from agricultural and urban
activities (Carpenter et al. 1998), although atmospheric deposition is considered an important source
of N to some areas. In particular, N inputs from the upper Midwest to the Gulf of Mexico have
increased dramatically in the past several decades, and excess nutrients have had severe effects on
water quality and community composition. A significant portion of the N from nonpoint sources
is retained in aquatic ecosystems by biological processes such as microbial uptake as well as lat-
eral exchange with the hyporheic zone. However, despite N retention in some aquatic ecosystems,
a large amount of excess N is transported downstream. Royer et al. (2004) measured denitrification
in headwater stream sediments located in agricultural areas. Because denitrification rates were low
in these streams, there was relatively little influence on instream concentrations and therefore most
of the NO
3
–N was transported downstream. These researchers concluded that previous estimates of
denitrification rates may have overestimated N loss to the sediments.
Because rates of nitrification and denitrification in aquatic ecosystems are dependent on concen-
trations of ammonium (NH
4

) and nitrate (NO
3
), these processes are likely to increase in areas receiv-
ing anthropogenic inputs. Kemp and Dodds (2002) measured effects of anthropogenic N on rates of
nitrification and denitrification in pristine and agriculturally influenced watersheds. Whole stream
nitrification and denitrification rates were greater at agriculturally influenced sites, most likely due
to greater input of N. Despite greater denitrification, the large amount of anthropogenic N exceeded
the natural retentive ability of the stream and a significant amount was transported downstream.
As described in Chapter 30, retention of nutrients and organic materials is dependent on a number
of physical, chemical, and biological characteristics. Headwater streams often represent the largest
portion of the linear dimension of a watershed and are closely connected to surrounding riparian and
terrestrial ecosystems. These systems are generally considered to be highly retentive of nutrients
(Peterson et al. 2001). A similar situation exists in coastal marine ecosystems. Meta-analysis of data
collected in coastal areas demonstrated that denitrification by fringing wetlands (e.g., salt marshes,
mangroves) serves to intercept excess nutrients and protect seagrass meadows from anthropogenic
N (Valiela and Coe 2002). Physical disturbances such as removal of vegetation and reduced habitat
complexity may decouple denitrification and nitrification processes in aquatic ecosystems, thereby
exacerbating the effects of nutrient enrichment (Kemp and Dodds 2002).
31.3 TERRESTRIAL ECOSYSTEMS
31.3.1 R
ESPIRATION AND SOIL MICROBIAL PROCESSES
Functional measures of ecosystem processes have also been used to characterize the impacts of
contaminants in terrestrial ecosystems. In particular, effects of contaminants on microbial and soil
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 675
ecosystem processes, especially soil respiration, have been examined in considerable detail (Aceves
et al. 1999, Dai et al. 2004, Megharaj et al. 1998, 2000, Zimakowska-Gnoinska et al. 2000). There
is also evidence suggesting that ecosystem processes in soils are significantly more sensitive to con-
taminants than the plant communities they support (see review by Giller et al. 1998). The emphasis

on soil processes in terrestrial ecosystems is at least partially a result of taxonomic difficulties asso-
ciated with characterizing microbial communities. Consequently, most soil ecologists tend to focus
on functional measures instead of structural characteristics such as biodiversity and species compos-
ition. Dai et al. (2004) reported a strong inverse relationship between heavy metal concentrations
in soils and respiration rates. An increase in the accumulation of organic C, total N, and the C:N
biomass ratio was attributed either to reduced microbial activity in soils or to changes in microbial
community composition. Similar results were reported by Edvantoro et al. (2003) where microbial
respiration was greatly reduced in soils contaminated by DDT and arsenic. Interestingly, although
microbial biomass was significantly lower in polluted soils, bacterial populations measured using
plate counts showed little difference between reference and contaminated sites.
Shifts in community structure, such asanincrease in biomass offungalpopulations and a decrease
in bacterial populations, can also result in changes in soil ecosystem function. Megharaj et al. (1998)
reported that changes in community composition of soil microalgae were associated with large
(>90%) reductions in microbial activity (dehydrogenase, nitratereductase)atfieldsitescontaminated
by pentachlorophenol (PCP). Megharaj et al. (2000) measured changes in soil microbial function at
sites contaminated by DDT and its metabolites. Because sensitive bacteria were replaced by DDT-
tolerant microorganisms, total microbial biomass was found to be less sensitive than dehydrogenase
activity, a measure oftotalmicrobial activity that correlates withrespiration(Brookes 1995). Changes
in respiration can result in carbon accumulation in ecosystems and therefore specific respiration rate,
measured as the ratio of CO
2
production to biomass C, may be a better indicator of stress than
either measure alone (Brookes 1995). For example, Megharaj et al. (2000) reported that specific
activity, defined as the ratio of dehydrogenase activity to microbial biomass C, decreased with DDT
concentration (Figure 31.5). Aceves et al. (1999) also determined that biomass-specific respiration
rate was a better indicator of heavy metal contamination in soils than either microbial biomass or
respiration.
Level of soil contamination
Control Low Medium High
Specific activity (dehydrogenase/microbial C)

0.00
0.02
0.04
0.06
0.08
0.10
0.12
Total DDT residue (mg/kg)
0
10
20
30
40
Specific activity
Total [DDT]
FIGURE 31.5 Relationship between DDT concentration in soil and specific microbial activity. Specific activ-
ity was calculated as the ratio of dehydrogenase activity (mg 2,3,5, triphenyltetrazoluim formazan/kg) to soil
microbial biomass (mg/kg). (Data from Tables 2 and 3 in Megharaj et al. (2000).)
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676 Ecotoxicology: A Comprehensive Treatment
Background Zn (log
10
mg/L)
HONEC (log
10
mg/L)
0.0 0.5 1.0 1.5 2.0 2.5
0.0
0.5

1.0
1.5
2.0
2.5
3.0
3.5
4.0
R
2
= 0.31
FIGURE 31.6 Relationship between background Zn concentration and the highest observed no effect con-
centration (HONEC) for a variety of soil microbial processes. (Data from Table 1 in McLaughlin and Smolders
(2001).)
Toxic effects of heavy metals on respiration rate has been reported for terrestrial ecosystems (Dai
et al. 2004, Laskowski et al. 1994, Niklinska et al. 1998); however, unlike our understanding of
aquatic ecosystems, we lack a clear understanding of biotic and abiotic factors that influence tox-
icity. For example, toxic metal concentrations that affect respiration can vary by more than 100 times
(Giller et al. 1998). Biological factors such as acclimation or adaptation of soil microorganisms to
contaminants may also explain some of the variation observed in these studies. McLaughlin and
Smolders (2001) used literature values to examine the influence of background Zn concentration
on the responses of soil microbial processes (e.g., respiration, nitrification, and ammonification)
to Zn. Background Zn concentrations in soil accounted for a significant amount of the variation in
sensitivity of several processes to Zn (Figure 31.6). Although considerable unexplained variation
resulted from methodological differences among studies and from comparing different endpoints,
the results clearly indicate that background concentration affected sensitivity to Zn. It is uncer-
tain if this increased tolerance to Zn affects the sensitivity of soil processes to other classes of
contaminants.
As in aquatic ecosystems, much of the variation among studies is a result of differences in
physicochemical factors that determine contaminant bioavailability. For example, metal toxicity
in soils is influenced by soil organic matter, clay content, pH, and other factors that regulate the

amount of free metals in solution. In addition, different responses are likely to be observed between
short-term experiments conducted in the laboratory and long-term monitoring studies conducted
in the field. The difficulty extrapolating from the laboratory to the field greatly complicates our
ability to define safe concentrations of contaminants necessary to protect soil processes. Giller
et al. (1998) advocate the use of long-term field experiments to explain effects of heavy metals
on microbial respiration and other soil processes. Because litter respiration has been proposed as
a potential endpoint in ecological risk assessment of forests ecosystems (Niklinska et al. 1998), a
better understanding of factors that influence effects of metals and other stressors on respiration is
necessary.
31.3.2 L
ITTER DECOMPOSITION
Litter decomposition is animportantprocess in terrestrial ecosystems thatisclosely related to primary
productivity, energy flow, and nutrient cycling. Large portions (>90%) of the net production in tem-
perate forests may be deposited on the forest floor as leaf litter. Thus, reduced rates of decomposition
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 677
Contamination
Arthropod
abundance
Microbial
activity
Litter
processing
Litter
accumulation
Nutrient cycling
Productivity
FIGURE 31.7 Conceptual model showing the potential mechanisms of contaminant effects on leaf lit-
ter processing, nutrient cycling, and productivity. Dashed lines indicate potential feedback between certain

processes.
in polluted terrestrial ecosystems may result in the accumulation of leaf litter and reduced supply of
nutrients through microbial mineralization (Figure 31.7). These changes in nutrient supply can affect
primary productivity and subsequent litterproduction. Because litter decompositionis relatively easy
to measure and sensitive to a variety of stressors, it is a useful endpoint for ecological risk assessment
in terrestrial ecosystems. Litter-bag studies measuring decomposition rates are frequently included in
assessments of contaminant effects (De Jong 1998, Johnson and Hale 2004, McEnroe and Helmisaari
2001, Strojan 1978). Slower rates of decomposition are often observed at contaminated sites, result-
ing in accumulation of organic material (Freedman and Hutchinson 1980, Strojan 1978). Reduced
litter decomposition can also affect nutrient cycling and growth of vegetation, thereby reducing soil
organic content and increasing contaminant bioavailability (Derome and Nieminen 1998, Johnson
and Hale 2004). Because contaminants may affect processing of detritus and accumulation of organic
material in soils, the ability of an ecosystem to assimilate contaminants may be reduced because of
lower organic content and subsequently greater contaminant bioavailability (Derome and Nieminen
1998). Alterations in litter decomposition resulting from contaminant deposition may also affect soil
processes and subsequent movement of contaminants. For example, changes in heavy metal con-
centrations in forest soils were attributed to altered decomposition and depletion of organic matter
at sites affected by a nearby aluminum industry (Egli et al. 1999). Derome and Nieminen (1998)
reported much greater flux of heavy metals through soils at disturbed sites adjacent to a smelter as
compared to distant sites. The increased flux of metals was associated with reduced interception of
precipitation by the disturbed canopy and a subsequent greater movement of water through denuded
soils.
Approaches employed in terrestrial ecosystems to measure litter decomposition are similar to
those in streams and generally involve measurement of weight loss of litter placed at reference and
contaminated field sites. Because decomposition rates are generally lower in terrestrial systems,
litter bags must be deployed for longer periods of time (e.g., 12–24 months) to obtain reliable
decay coefficients (k). Terrestrial insects and other macroinvertebrate decomposers are likely to play
a key role in regulating rates of decomposition. However, unlike research conducted in aquatic
ecosystems, most terrestrial studies employ relatively fine mesh litter bags designed to exclude
invertebrate decomposers. As a consequence, the relative importance of invertebrate decomposers,

microbial processes, and physical processes in regulating litter decomposition at contaminated sites
is uncertain in terrestrial ecosystems.
Much of the research employing litter bags to examine decomposition rates has focused on
soils contaminated with heavy metals (Breymeyer et al. 1997, Coughtrey et al. 1979, Johnson and
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678 Ecotoxicology: A Comprehensive Treatment
Hale 2004, McEnroe and Helmisaari 2001, Post and Beeby 1996). Accumulation of litter in a forest
near a lead-zinc-cadmium smelter was attributed to reduced decomposition associated with heavy
metal contamination (Coughtrey et al. 1979). McEnroe and Helmisaari (2001) measured decomposi-
tion of pine needles along a gradient of metal contamination in soil. Decreased mass loss and reduced
C:N ratios were attributed to metal contamination in soils and litter. Metal concentrations in litter,
which increased significantly over the 30-month study, may have contributed to reduced decompos-
ition rates. Although many of the studies measuring litter decomposition have focused on localized
sources such as smelters and abandoned mines, nonpoint sources may alter litter decay rates. Metals
deposited along roadsides were found to inhibit litter decomposition, but this effect resulted from
contamination of plant litter rather than differences in metal content of soils (Post and Beeby 1996).
Although a majority of studies have reported a negative relationship between contaminant levels
and rates of decomposition in terrestrial ecosystems, decomposition may be enhanced in some
polluted habitats. Breymeyer et al. (1997) measured decomposition rates at 15 pine forest sites along
a gradient of metal contamination. Positive correlations between litter decomposition and metal
concentrations were attributed to fertilization effects in the nutrient-poor soils typical of pine forest
ecosystems. Similar results were reported by Post and Beeby (1996) where increased microbial
respiration along urban roadways was attributed to elevated concentrations of resources in the form
of hydrocarbons. Kauppi et al. (1992) attributed an observed increase in forest production in Europe
between 1960 and 1990 to the fertilization effects of some pollutants, which apparently outweighed
adverse effects in these systems.
31.3.2.1 Mechanisms of Terrestrial Litter Decomposition
Despite the large number of studies over the past several decades that have examined the relationship
between contaminant levels in soilsand reduced decomposition rates, the underlying mechanisms are

not well understood (Coughtreyet al. 1979, Johnson andHale 2004, Kohleret al. 1995). Becausecon-
taminants may affect microbial processesin the soil, litter palatability, and/or invertebrateabundance,
additional studies that quantify the relative importance of these processes are necessary. Freedman
and Hutchinson (1980) conducted acomprehensiveanalysis of the effects of smelter emissions onleaf
litter decomposition and the underlying mechanisms responsible for litter accumulation at field sites
(Figure 31.8). Reduced abundance of soil arthropods and lower microbial activity (measured as soil
respiration and acid phosphatase activity) resulted in lower rates of litter decomposition at contamin-
ated sites compared with reference sites. The lower rates of litter processing resulted in a significant
increase in litter accumulation at sites adjacent to the smelter. Interactions between soil arthropods
and microbial communities may also influence leaf litter processing. Kohler et al. (1995) reported
that direct stimulation of microbial activity by arthropods in soils enhanced litter decomposition.
Although it is generally recognized that microbial processes controlling decomposition are par-
ticularly sensitive to contaminants (Derome and Lindroos 1998), few studies have quantified the
relative importance of contaminants in soils versus those deposited directly on litter. Contaminants
deposited on the surface of leaves before litterfall could have direct effects on both macroinvertebrate
decomposers and microbial processes. Distinguishing the relative importance of these mechanisms
has important implications for remediation of contaminated sites. Clean-up of polluted soils will
likely have little effect on litter decomposition rates if microbial processes are inhibited by contam-
inants deposited on leaf surfaces. Breymeyer et al. (1997) reported that atmospheric deposition of
metals to litter was more important in controlling decomposition than metal concentrations in soils.
Reciprocal transplant studies, in which decomposition of clean and contaminated litter is measured at
reference and impacted sites, is an effective approach for distinguishing the effects of contaminants
in soils from those deposited on leaf litter. Johnson and Hale (2004) measured metal accumulation
and decomposition of white birch leaves at metal-polluted sites in Canada. Decomposition rates were
reduced at a metal-polluted site, but concentrations of metals in litter, which accumulated over time,
did not influence dry mass loss.
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 679
2.8 3.5 6.5 10.9 22.1 43

Respiration (µL CO
2
/min)
0
2
4
6
8
10
12
14
16
18
20
Distance from smelter (km)
3.5 5.5 29.9 34.3
Number of soil arthropods
0
2
4
6
8
10
12
Time (d)
0 200 400 600 800 1000
Percent loss
20
30
40

50
60
70
80
90
Distance from smelter (km)
0 10203040
Litter accumulation (g x 10
2
/m
2
)
0
5
10
15
20
25
Contaminated site
Reference site
(a)
(c)
(b)
(d)
FIGURE 31.8 Effects of heavy metals on microbial respiration (a), abundance of soil arthropods (b), decom-
position rate of birch leaves at reference (open symbols) and impacted (closed symbols) sites (c), and litter
accumulation (d) in forest soils in the vicinity of a nickel-copper smelter at Sudbury, Ontario. (Data from
Figure 1 and Tables 2, 5, and 8 in Freedman and Hutchinson (1980).)
In addition to studying decomposition of leaves, some researchers have used standardized mater-
ials, such as cellulose strips buried in soils, to quantify effects of contaminants on decomposition

(De Jong 1998, Post and Beeby 1996). Because these materials degrade rapidly, decomposition
studies can be completed in a much shorter period of time when compared to natural litter. As with
any attempt to standardize approaches used in ecological assessments, the advantages of greater
control must be weighed against the loss of ecological realism. Post and Beeby (1996) concluded
that natural litter bags were more appropriate for measuring effects of contaminants on activity of
microbial decomposers because of the relatively unrealistic resources provided by cellulose strips.
De Jong (1998) conducted a series of experiments designed to assess the effects of pesticides on
litter decomposition. Because of considerable variability and susceptibility of decomposition studies
to external confounding factors, this researcher concluded that “the chance of developing a stand-
ardized field trial with litterbags is deemed too slim.” If litter decomposition is to become a widely
used endpoint in risk assessments of terrestrial ecosystems, a better appreciation for the underlying
mechanisms and complexities that control this process is essential.
31.3.3 NUTRIENT CYCLING
The majority of studies investigating the influence of contaminants on nutrient dynamics in ter-
restrial ecosystems have focused on nitrification, which is defined as the conversion of ammonium
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680 Ecotoxicology: A Comprehensive Treatment
to nitrite (NO
2
) and nitrate (NO
3
). Nitrification is a critical component of the nitrogen cycle that
ultimately determines the availability of soil N to plants and other organisms. The sensitivity of this
two-step process to contaminants is dependent on the relative sensitivity of the two groups of bacteria
responsible (Nitrosomonas and Nitrobacter). Several studies have examined effects of heavy metals
in soils on nitrification rate (Sauve et al. 1999, Smolders et al. 2001, 2003). Potential nitrification
rate (PNR) is measured by amending soils with known amounts of ammonium and then measur-
ing changes in nitrate concentration over time (Sauve et al. 1999). Rates of nitrification are also
dependent on physicochemical characteristics of soil such as pH, the amount of organic material,

and soil composition. Because these soil characteristics may also vary with metal contamination,
field assessments of nitrification must account for natural differences at reference and contaminated
sites. For example, metal deposition from smelters is often associated with soil acidification, which
would increase metal bioavailability. In contrast, soils contaminated by sewage sludge may have
elevated levels of organic material, which would likely reduce contaminant bioavailability. There is
also the potential for positive feedback relationships between the bioavailability of contaminants and
soil properties. For example, the amount of organic material in soils is in part regulated by nitrifica-
tion and microbial decomposition (Figure 31.7). Because nitrification and microbial decomposition
are sensitive to contaminants, bioavailability would likely be increased in contaminated soils where
these processes are inhibited.
Although reduced PNR has been observed in the laboratory using metal-spiked sediments, some
field studies have shown relatively little effects of metals on PNR. Differences in bioavailability of
Zn in field-contaminated and laboratory-spiked soils likely account for these differences. Sauve et al.
(1999) reported both positive and negative relationships between heavy metal concentration in soils
and nitrification rate. Differences in the effects of metals were attributed to the confounding effects
of soil pH and organic matter content, which often accounted for a greater amount of variation
than soil metal concentration. Although EC50 concentrations for Zn-spiked soils were between
150 and 350 mg/kg dry weight (Smolders et al. 2001), there was little relationship between Zn
concentration and nitrification in field-contaminated soils. These researchers questioned the utility
of PNR in assessments of contaminated soils because of high background variability and inconsistent
responses in uncontaminated ecosystems.
Smolders et al. (2003) compared PNR in laboratory-spiked and natural sediments contamin-
ated with Zn. Nitrification rate in the field increased significantly with soil pH, but Zn had little
effect at concentrations shown to be toxic in the laboratory. These differences between natural and
laboratory-spiked sediments were attributed to either microbial acclimation to metals or lower metal
bioavailability in the field. Rusk et al. (2004) reported that adaptation of soil microbes to metals
significantly reduced effects on nitrification. Sauve et al. (1999) questioned the utility of nitrification
as an indicator of contamination because of its sensitivity to other factors such as pH and organic
carbon. Because of potential differences in effects of metals and other contaminants in field-collected
and laboratory-spiked soils, studies that validate laboratory responses are necessary. Furthermore,

because soil pH and organic matter may affect nitrification rates and contaminant bioavailability,
experimental approaches are necessary to determine if the relationship between nitrification and soil
characteristics is a result of direct or indirect effects.
31.3.4 AN INTEGRATION OF TERRESTRIAL AND AQUATIC
PROCESSES
Biogeochemists generally recognize that aholistic understanding of the factors that regulate transport
and transformation of nutrients in ecosystems requires integration of aquatic andterrestrial processes.
Increased alteration of local, regional, and globalcyclesofC,N,andPrequiresthatecologistsdevelop
a broader perspective of factors that influence nutrient movements between ecosystems. Regardless
of the ecosystem, the primary focus should be on factors that limit primary productivity, affect nutri-
ent retention, and alter ratesof nutrient transformation. However, theories describing biogeochemical
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 681
processes in aquatic and terrestrial ecosystems have largely been developed in isolation, with relat-
ively fewer attempts to integrate these disciplines (Grimm et al. 2003). As a result, very different
approaches have been employed in aquatic and terrestrial ecosystems to quantify effects of natural
and anthropogenic stressors on nutrient processing. Grimm et al. (2003) discuss underlying physical,
chemical, and biological characteristics of aquatic and terrestrial ecosystems that led to the applic-
ation of disparate approaches used in these systems. An important distinction between terrestrial
and aquatic ecosystems is that N often limits productivity in terrestrial ecosystems whereas P is
frequently the primary limiting nutrient in aquatic ecosystems. Although this generalization is some-
what overstated, it is at least partially responsible for the different research questions addressed in
aquatic and terrestrial systems (Grimm et al. 2003). Research in terrestrial ecosystems, particularly
in agricultural systems, has largely focused on understanding factors to increase plant productivity.
In contrast, much of the emphasis in aquatic systems has been devoted to understanding factors that
limit productivity and to assessing the negative effects of eutrophication. Not surprisingly, many of
the differences in key structural features (geomorphology, hydrology, nutrient pools, lifespan, and
size of dominant primary producers) and flux of elements between aquatic and terrestrial ecosystems
are closely linked to the quantity and movement of water. A better understanding of the effects of

contaminants on nutrient dynamics in terrestrial and aquatic ecosystems will require development
of methodological approaches that transcend boundaries and allow direct comparison of processes
between systems.
31.4 SUMMARY
Descriptive studies have greatly improved our understanding of contaminant effects on ecosystem
processes such as production, decomposition, and nutrient cycling. The potential for conducting
surveys of contaminant effects on ecosystem processes over a relativelylarge spatiotemporal scale is a
major strength ofthese comparative approaches. Whileattempts to predict ecosystem-levelresponses
to contaminants based on changes in community composition have been reasonably successful,
we feel that direct measurement of ecosystem processes is considerably more informative. We
will explore the ecotoxicological implications of the relationship between structural and functional
characteristics in Chapter 33. One of the more significant developments in ecosystem ecology is
the recognition that traditional boundaries between certain ecosystems are often arbitrary (Fausch
et al. 2002). Although these boundaries have historically defined the spatial extent of individual
ecosystems, the exchange of contaminants and other materials between ecosystems requires a more
holistic perspective. We consider the integration of methodological approaches for aquatic and
terrestrial ecosystems, which have been largely developed in isolation (Grimm et al. 2003), to be a
major research need in ecosystem ecotoxicology.
Because the ecosystemprocessesdescribed in this chapterare strongly interrelated andinfluenced
by natural biotic and abiotic factors (Figure 31.7), quantifying the direct effects of contaminants is
challenging. A more comprehensive understanding of underlying mechanisms would facilitate our
ability to quantify the importance of natural changes in ecosystem processes relative to contaminant-
induced alterations. In the following chapter, we will examine the application of experimental
approaches such as the use of microcosms, mesocosms, and whole ecosystem manipulations to
assess effects of contaminants on ecosystem processes.
31.4.1 SUMMARY OF FOUNDATION CONCEPTS AND PARADIGMS
• Ecotoxicologists interested in ecosystem responses to anthropogenic stressors employ
descriptive, quasi-experimental, and experimental approaches.
• Because experimental studies of ecosystem processes are often limited by spatiotemporal
scale, descriptive approaches can provide compelling and ecologically realistic results.

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682 Ecotoxicology: A Comprehensive Treatment
• Restricting our analyses to mainly structural measures has provided a somewhat incom-
plete picture of how ecosystems respond to and recover from anthropogenic disturbances.
• Lake ecologists have historically relied on functional measures, especially primary
production, whereas lotic ecologists have tended to focus on structural measures.
• Primary production in aquatic ecosystems is particularly sensitive to anthropogenic
stressors.
• Input of nutrients associated with agricultural, domestic, industrial, and atmospheric
sources are widely regarded as major stressors of aquatic ecosystems.
• Direct effects of nutrients on primary production are complex and may be mediated by
other factors such as hydrologic characteristics and abundance of grazers.
• The most common functional response related to energetics measured in aquatic
ecosystems is the abundance of different functional feeding groups.
• Secondary production integrates estimates of individual growth rates and population
dynamics and therefore captures in a single measure several important aspects of energy
flow through ecosystems.
• One significant challenge associated with using secondary production as an indicator of
ecosystem integrity is that production may increase or decrease, depending on the nature
of the stressor.
• Effects of contaminants on litter decay may result either from alterations in microbial
processes or reduced abundance of macroinvertebrate shredders.
• The most convincing evidence demonstrating a relationship between acidification and leaf
decomposition has been obtained from spatially extensive surveys.
• Most investigations of leaf litter decomposition report that decay coefficients are reduced
in contaminated streams; however, stressors that subsidize an ecosystem (e.g., nutrients
or organic materials) may have the opposite effect.
• Most studies investigating effects of contaminants on nutrient dynamics in aquatic eco-
systems have focused on nitrification, denitrification, and other processes associated with

N flux.
• Effects of contaminants on microbial and soil ecosystem processes, especially soil
respiration, have been examined in considerable detail.
• There is evidence suggesting that ecosystem processes in soils are more sensitive to
contaminants than the plant communities they support.
• Contaminant-induced shifts in community structure can also result in changes in soil
ecosystem function.
• Litter decomposition isan important process interrestrial ecosystems that isclosely related
to primary productivity, energy flow, and nutrient cycling.
• Because decomposition rates are generally lower in terrestrial systems, litter bags must
be deployed for longer periods of time (e.g., 12–24 months) to obtain reliable decay
coefficients (k).
• Despite the large number of studies that have examined the relationship between con-
taminant levels in soils and decomposition rates, the underlying mechanisms are not well
understood.
• The majority of studies investigating the influence of contaminants on nutrient dynamics
in terrestrial ecosystems have focused on nitrification.
• Because nitrification and microbial decomposition are sensitive to contaminants and influ-
ence the amount of organic materials in soils, contaminant bioavailability would likely be
increased in ecosystems where these processes are inhibited.
• A holistic understanding of factors that regulate transport and transformation of nutrients
in ecosystems requires integration of aquatic and terrestrial processes.
• Theories describing biogeochemical processes in aquatic and terrestrial ecosystems have
largely been developed in isolation.
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Descriptive Approaches for Assessing Ecosystem Responses to Contaminants 683
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