United States Office of Water EPA843-R-01-
Environmental Protection Wetlands Division (4502F) Fall 2001
Agency Washington, DC 20460
Indicators for Monitoring
Biological Integrity of Inland,
Freshwater Wetlands
A Survey of North American Technical Literature (1990-2000)
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Indicators for Monitoring Biological Integrity of Inland, Freshwater Wetlands
A Survey of North American Technical Literature (1990-2000)
by
Paul Adamus
Department of Fisheries and Wildlife
Oregon State University
Corvallis, Oregon
Thomas J. Danielson
Maine Department of Environmental Protection
Augusta, Maine
Alex Gonyaw
Department of Fisheries and Wildlife
Oregon State University
Corvallis, Oregon
U.S. Environmental Protection Agency
Office of Water
Office of Wetlands, Oceans, and Watersheds
Washington, DC 20460
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Notice
Part of the information in this report was compiled while one of the authors (T.
Danielson) was at Duke University's Nicholas School of the Environment in Durham,
NC. Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
Acknowledgments
We thank Doreen Vetter and Susan Jackson, USEPA Office of Wetlands and Watersheds,
for recognizing the importance of this effort and for arranging the cooperative agreement
with Oregon State University. We also thank Dr. Erik Fritzell at Oregon State, and
Concepcion Cahanap at USEPA, for administering the project, and members of the
Biological Assessment of Wetlands Working Group (BAWWG) for encouraging this
effort.
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Section 1. Introduction
This document has been written for wetland managers, researchers, and monitoring
specialists. For wetland managers, it serves as a resource for identifying and
understanding biological impacts that could result from regulated and unregulated
activities in wetlands. For researchers and monitoring specialists, it facilitates
interpretation of collected data by providing a context of what we already know.
In preparing this, our sole focus has been to update a literature review on the same topic
sponsored and published previously by the USEPA (Adamus and Brandt 1990), see:
http://www.epa.gov/owow/wetlands/wqual/introweb.html
As such, this document is not intended as stand-alone guidance for persons wishing to
learn how to develop wetland biomonitoring programs, or for persons seeking to
prioritize research. For additional wetland monitoring information, databases,
publications, and guidance, readers should see:
http://www.epa.gov/owow/wetlands/bawwg/
This document grows from the recognition that in-depth knowledge of the most current
research findings is imperative for developing and using scientifically-sound biological
indicators of wetland condition. This document has the following important attributes:
• Literature from the period 1990-2000 is the primary focus.
• Literature from North American wetlands is the primary focus.
• Literature from inland freshwater wetlands is the primary focus.
• Literature on impacts to assemblages of species, rather than single species, is the
primary focus
• Literature in peer-reviewed journals is referenced almost exclusively.
• Not every paper fitting the above descriptions could be reviewed. However, we
believe this document — based on review of over 1500 publications — covers a
majority of the relevant literature. The largest numbers of publications are cited in
the Invertebrates and Vascular Plants sections of this document, but for these two
topics we also excluded the largest numbers of relevant papers, due to limited time
for review relative to the enormous number that were published in the last decade.
The past decade has seen predictable diversification of wetland research into an
enormous array of subdisciplines and subtopics. A multitude of subjects previously
unexplored and some never imagined have emerged in the literature. Our approach in
preparing this document has been to emphasize wide coverage of the wetlands biological
literature, rather than cover any particular topic or subtopic in depth. Because of the
enormous number of studies that have been published, time constraints, and our stated
goals for the effort, we have sought primarily to organize the recent literature in a helpful
way, not to interpret or synthesize it.
The document is structured around 11 categories of human-related disturbances to which
wetlands are commonly exposed (Table 1.1), and the effects of these "stressors" on the
following groups, each the focus of a separate section: microbes, algae, vascular plants,
invertebrates, fish, amphibians, and birds. This document assumes the reader is generally
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familiar with wetland terminology and ecological principles, as well as with terms and
concepts that are associated with indices of biological integrity. For this reason no
glossary is provided. This document does not attempt to summarize our understanding of
each combination of stressor and biological group, but instead simply describes literature
published since 1989 on each pairing of stressor with biological group. Information from
saltwater wetlands, non-wetland systems, and foreign literature is included sparingly,
mainly when it serves to highlight significant gaps in our understanding of the types of
wetlands covered by this document.
Table 1.1. Human-related stressors addressed in this document
Enrichment, Eutrophication, Organic Loading, Reduced Dissolved Oxygen (DO). Interrelated
increases in concentration or availability of nitrogen or phosphorus. Typically associated with
excessive fertilizer application, livestock waste management, wastewater treatment systems, fossil fuel
combustion, unmanaged urban runoff, and other sources. Includes increases in carbon, to the point
where an increased biological oxygen demand reduces dissolved oxygen in the sediments and the
water column and increases toxic gases (e.g., hydrogen sulfide, ammonia).
Contamination Toxicity. Increases in concentration, availability, and/or toxicity of metals and synthetic
organic substances. Typically associated with agriculture (pesticide applications), aquatic weed
control, mining, urban runoff, landfills, hazardous waste sites, fossil fuel combustion, wastewater
treatment systems, and other sources.
Acidification. Increases in acidity (decreases in pH). Typically associated with mining and fossil fuel
combustion.
Salinization. Increases in dissolved salts, particularly chloride, and related to parameters such as
conductivity and alkalinity. Typically associated with road salt used for winter ice control, irrigation
return waters, stormwater, seawater intrusion (e.g., due to land loss or aquifer exploitation), and
domestic/industrial uses.
Sedimentation/Burial. Increases in deposited sediments, resulting in partial or complete burial of
organisms and alteration of substrate. Typically associated with agriculture, disturbance of stream
flow regimes, urban runoff, wastewater treatment plants, deposition of dredged or other fill material,
and erosion from mining and construction sites.
Turbidity/Shade. Reductions in solar penetration of waters as a result of blockage by suspended
sediments and/or overstory vegetation or other physical obstructions. Typically associated with
agriculture, disturbance of stream flow regimes, urban runoff, poorly functioning wastewater treatment
plants, and erosion from mining and construction sites, as well as from natural succession, placement
of bridges and other structures, and resuspension by fish (e.g., common carp) and wind.
Vegetation Removal. Typically associated with aquatic weed control, agricultural and silvicultural
activities, channelization, bank stabilization, urban development, defoliation from airborne
contaminants and other stressors included in this report, grazing/herbivory (e.g., from muskrat, grass
carp, geese, crayfish, insects), disease, and fire.
Thermal Alteration. Long-term changes (especially increases) in temperature of water or sediment.
Typically associated with power plants, other industrial facilities, removal of shading vegetation,
lowering of summertime water levels, and global climate warming.
Dehydration. Reductions in wetland water levels and/or increased frequency, duration, or extent of
desiccation of wetland sediments. Typically associated with ditching, channelization of nearby
streams, invasion of wetlands by highly transpirative plant species, outlet widening, subsurface
drainage, global climate change, and ground or surface water withdrawals for agricultural, industrial,
or residential use.
Inundation. Increases in wetland water levels and/or increase in the frequency, duration, or extent of
saturation of wetland sediments. Typically associated with impoundment (e.g., for cranberry or rice
cultivation, flood control, water supply, waterfowl management) or changes in watershed land use that
result in more runoff being provided to wetlands.
Other Human Influences. Increases in distance between, and reduction in sizes of, patches of suitable
habitat (i.e., fragmentation). Increases in noise, predation from pets, disturbance from visitation,
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invasion by non-native species capable of outcompeting species that normally characterize intact
communities; electromagnetic, ultraviolet (UV-B), and other radiation; and other factors not addressed
above.
In addition to addressing the above for each biological group, this report briefly
summarizes published information most relevant to monitoring the particular group.
Within each group and under a subheading "Wetland Monitoring," recent information is
compiled on spatial and temporal variation, techniques and equipment for monitoring the
group, and biological metrics that have or have not been found to correlate with
individual or cumulative disturbances to wetlands.
Ideally, it would be best to separate the published results by wetland type (e.g., Cowardin
class, hydrogeomorphic class). Due to the lack of information on many groups, however,
it was not practical to do so in this document. Thus, readers should be cautious when
attempting to extrapolate the reported results.
This document was developed in four stages. First, Thomas Danielson identified,
obtained, and prepared written reviews of relevant literature covering the period 1990-
1996, with the exception of literature on wetland microbes and fish. Second, Alex
Gonyaw (a student supervised by Paul Adamus in the Fisheries and Wildlife Department
at Oregon State University) identified, obtained, and prepared written reviews of relevant
literature covering wetland microbes and fish, plus updated the sections on the other
groups, through literature published in 2000. Third, Paul Adamus edited the manuscript
extensively and reviewed hundreds of publications that the co-authors had either not
known about, or had not had time to review. Fourth, after re-readings and comments by
the co-authors as well as external peer reviewers, Paul Adamus prepared the final
document.
At every stage, potentially relevant literature for the years 1990-1999 was identified by
(a) conducting keyword searches of computerized bibliographic databases, especially
CAB Abstracts and Aquatic Sciences and Fisheries Abstracts, (b) reading through the
tables of contents of a few especially relevant journals, (c) searching the internet for
pertinent bibliographies, and (d) to a lesser extent, reviewing articles listed in these
bibliographies and in the literature cited sections of relevant journal articles and books.
Information and references from the parallel review of wetland biological studies
specifically from Florida, prepared for the US EPA by Steve Doherty and others, were
selectively incorporated.
Comments regarding information presented in the document should be addressed to:
Paul R. Adamus
6028 NW Burgundy Dr.
Corvallis, OR 97330
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Section 2. Microbial Assemblages and Processes
2.1 Use as Indicators
This section addresses microbes that are closely associated with naturally-occurring
wetlands. Included in this discussion are bacteria, protozoans, viruses, yeasts, and
microscopic fungi (including mycorrhizae and hyphomycetes). Microscopic algae are
discussed in Section 3. Like the other sections in this document, this section focuses
almost entirely on research published since 1989. For a general discussion of the topic
based on pre-1990 scientific studies, and for broader discussion including advantages and
disadvantages of using microbial assemblages and processes as indicators of wetland
integrity, readers should refer to Adamus and Brandt (1990).
Microbial organisms are omnipresent in wetlands, even living within the individual
submerged roots of some wetland shrubs (Fisher et al. 1991). Through interactions with
wetland plants and hydrology, wetland microbial assemblages can remove inorganic
nutrients, heavy metals, dissolved organic carbon, particulate organic matter, and
suspended solids from the water column and sediments (Mickle 1993), as well as play a
key role in supporting food webs (Schallenberg & Kalff 1993) and influencing global
climate change through their role in methanogenesis (Bartlett & Harriss 1993,
Kumaraswamy et al. 2000). The presence of bacteria in "biofilms" on the enormous plant
and detrital surface area in wetlands is fundamental to wetland ability to degrade complex
organic contaminants (Hamilton et al. 1993, Taylor et al. 1996). Iron-oxidizing bacteria
in roots of wetland plants also influence plant nutrition (Emerson et al. 1999). Production
from indigenous bacteria may surpass production from algae in forested wetlands
(Hudson et al. 1992). Many naturally-occurring bacteria inhibit waterborne pathogens.
For example, bacteria that normally inhibit Clostridium botulinum type C (an organism
that causes extensive waterfowl deaths) were found to be quite widespread in northern
California wetlands (Sandier et al. 1998). A reduction in these inhibiting bacteria might
lead to an increase in botulism poisoning of waterfowl.
Among the microbes, root fungi (mycorrhizae) are particularly important. Along with
aquatic invertebrates, they are responsible for decomposing dead plant material, thus
constantly renewing system nutrients (Tuchman 1993). Along with nitrogen-fixing
bacteria, fungi supply nitrogen and/or phosphorus to wetland plant roots (Allen 1991,
Wigand & Stevenson 1997, Pieher et al. 1998, Turner & Friese 1998, Lovell et al. 2000,
Piceno & Lovell 2000). Indeed, lack of mycorrhizae in some constructed wetlands has
been suggested as a reason for failure of plantings. Mycorrhizae occur widely in
wetlands: an examination of 290 plants from 89 species of Connecticut wetland plants
found that all species of mature plants as well as selected young plants on developing
shorelines were colonized by mycorrhizal fungi (Cooke & Lefor 1998). This, plus
surveys in Midwest wetlands (Turner & Friese 1998) and southeastern bottomland
hardwoods (Jurgensen et al. 1997) and elsewhere suggest that mycorrhizae are common
even among plants associated with fluctuating water, nutrient, and oxygen conditions.
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Fungi also can influence the structure of vascular plant communities in wetlands. For
example, lack of ectomycorrhizal fungi in beaver meadows (probably as a result of
exclusion of fungi-spreading rodents by prolonged flooding) may prohibit the succession
of these meadows into forested wetlands (Terwilliger & Pastor 1999). Reductions in
ectomycorrhizal infection of willows in Alaska wetlands, as caused by herbivore
browsing, caused a shift in plant community composition (Rossow et al. 1997). Southern
boreal bogs and fens contain mostly mycorrhizal fungi that enable characteristic plant
species to proliferate in these nutrient-poor ecosystems by accessing otherwise
unavailable nutrient pools. In contrast, marsh vegetation mainly contains non-
mycorrhizal fungi, possibly due to higher surface-water nutrient concentrations and
fluctuating water levels (Thormann et al. 1999). In general, little is known about the
effects of contaminant additions or other wetland alterations on mycorrhizae (Cairney &
Meharg 1999).
Because microbial assemblages have so many important ecosystem functions, and
because these functions are typically easier to measure than the taxonomic structure of
the responsible microbial community, most literature on impacts describes effects on
these functions rather than on the microbes themselves. To date, no North American
studies have used microbial taxonomic or functional diversity or composition to assess
the ecological conditions of a large series of wetlands, but use of microbes as indicators
of aquatic system integrity has been considered generally (Lynn & Gilron 1992) and with
reference to contaminants in particular (Maier et al. 2000).
2.2 Effects of Enrichment, Eutrophication, Reduced Dissolved Oxygen
Bacterial assemblages, with generation times as little as 15 minutes are well suited for
detecting short-term nutrient pulses (Miyamoto and Seki 1992). In an Ohio marsh,
experimental dosing with phosphate stimulated an increase in bacterial density (Willis
and Heath 1993). Excessive enrichment can quickly allow anaerobic taxa to gain
dominance. Microbial assemblages receiving agricultural nutrient inputs in part of the
Florida Everglades were dominated by methanogens, sulfate reducers, and acetate
producers (Drake et al. 1996). These bacteria flourish where porewater total phosphorus
concentrations and conductivities are high (Drake et al. 1996). Excessive nutrients from
agricultural operations may reduce the normal ability of wetland microbial assemblages
to detoxify particular pesticides (Kazumi & Capone 1995, Chung et al. 1996, Entry &
Emmingham 1996, Entry 1999). Although nitrogen additions to a riparian system briefly
stimulated bacterial and fungal activity, long-term effects were perceived as negative,
thus potentially compromising the ability of the system to remove nitrogen via
denitrification (Ettema et al. 1999). Two constructed wetlands in Massachusetts had
microbial biomass and potential denitrification that were within the range of variability of
natural wetlands, owing at least partly to use of organic sediments in the construction of
the wetlands (Duncan & Groffman 1994). Sediments in the Fox River/Green Bay
watershed of Wisconsin were not toxic ioPhotobacterium phosphor eum despite having
elevated ammonia (1.3-54.4 mg/L) (Ankley et al. 1990). Some bacteria (termed nitrifiers)
play a key role in wetland food webs by converting ammonium and nitrite to nitrate,
which supports algae and vascular plants (Groffman et al. 1992). This nitrifying capacity
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is influenced by the plant species with which the bacteria are associated (the plants may
secrete antibacterial substances) and not with the cation exchange capacity of the plant
litter (Eriksson & Andersson 1999).
In New York, experimental additions of particulate detritus derived from the most
common submerged macrophyte (Vallisneriaamericana) and wetland plant (Typha
angustifolia) to Hudson River water did not result in increases in bacterial productivity.
In contrast, additions of dissolved organic carbon derived from these same plants
consistently yielded large increases in bacterial production (Findlay et al. 1992). Growth
response of bacterial colonies in streams may depend on timing and source of natural
leachates from local plants as well as on sources of dissolved organic carbon from further
upstream. Growth of bacterial assemblages in streams exhibited 'generalist'
characteristics in headwater reaches and 'specialist' characteristics farther downstream
(Koetsier et al. 1997).
Decomposition rates, which generally reflect microbial activity, increased along a
eutrophication gradient from a bog, a poor fen, a wooded moderate-rich fen, a lacustrine
sedge fen, a riverine sedge fen, a riverine marsh, and a lacustrine marsh in southern
Alberta (Thormann etal. 1999). Although ammonium fertilizers have been thought to
potentially increase methane emissions from wetlands (due to effects on particular
bacterial assemblages), evidence to the contrary was reported by Bodeller et al. (2000).
Colonization of wetland plants by mycorrhizal fungi is sometimes greater in less fertile
(low phosphorus) wetlands (Wetzel & Van der Valk 1996, White & Charvat 1999), but
nutrient levels are probably not a major influence. Also, mycorrhizae can be restricted by
low oxygen conditions typical of some eutrophic wetlands (Johnson et al. 1992, Johnson
1993, Cantelmo & Ehrenfeld 1998).
2.3 Effects of Contaminant Toxicity
Bacteria have a variety of responses to contaminants, ranging from direct utilization of
toxins as energy sources, to being being harmed or unaffected. Crude oil, leaded
gasoline, and the herbicide 2,4,5-T reduced taxonomic and/or functional diversity of
microbial communities (Atlas et al. 1991). High concentrations of iron, manganese,
magnesium, calcium and nickel in water from an abandoned coal mine reduced the
diversity of fungi that decompose leaf litter. This, plus a shift in fungal species
composition, reduced the leaf decomposition rates (Bermingham et al. 1996). Several
other contaminants, by inhibiting microbial and invertebrate decomposers, can slow the
rate of decomposition of wetland vegetation at the end of each growing season
(Schultheis and Hendricks 1999), potentially causing anoxia in wetlands as springtime
temperatures rise. However, the species composition of fungal decomposers in streams
exposed to the insecticide methoxychlor did not differ significantly from that in
unexposed streams (Suberkropp & Wallace 1992). A single application of the pesticide
diquat reduced protozoan species richness, net productivity, community respiration, and
enzyme activities. Both photosynthetic and non-photosynthetic taxa were affected by
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diquat (Melendez et al. 1993).
Antibacterial agents (e.g., tetracyclin) in domestic wastewater, such as from leaky septic
systems and feedlot runoff, have the potential for inhibiting microbial assemblages and
thus harming fundamental ecosystem processes. Some instances have been noted of
evolution of resistant strains in nature, thus posing potential threats to human health
(DePaola 1995).
Some microbial assemblages tolerate or even biodegrade particular contaminants. Such
assemblages can be naturally selected and flourish in some wetlands that are only mildly
or moderately contaminated. Contaminants that can be processed when at low-moderate
concentrations by microbes include copper (Farago and Mehra 1993), mercury (Marvin-
Dipasquale & Oremland 1998), selenium (Steinberg & Oremland 1990, Azaizah 1997),
cadmium (Sharma et al. 2000), manganese (Sikora et al. 2000), and petroleum (Nyman
1999, Megharaj et al. 2000). Microbial assemblages attached to plant roots also help
wetland plants take up some metals (e.g., selenium, mercury) contained in runoff (Souza
et al. 1999). Sulfate-reducing bacteria play a major role in detoxifying some
contaminants, and may be most diverse in oxygen-containing environments (Minz et al.
1999), but other functional groups of bacteria can also be important (Webb et al. 1998).
A wetland fed by leachate from an abandoned landfill in Washington supported large
populations of total and fecal coliforms, fecal streptococci, the opportunistic pathogens
Pseudomonas aeruginosa, Staphylococcus aureus, Enterococcus faecalis, and the
anaerobe Clostridiumperfringens (Boening and Vasconcelos 1997). A Georgia tidal
wetland containing elevated levels of mercury and PCB's showed little evidence of
adverse impact to fungal communities; elevated nutrient concentrations associated with
the contaminants may have somewhat offset any toxic effects occurring at the population
level (Newell & Wall 1998).
Compensatory mechanisms within microbial assemblages sometimes cloud the response
to complex mixtures of toxicants. In aquatic microcosms with continuous dosing, the
herbicide atrazine was found to increase microbial richness and biomass, whereas copper
reduced these attributes and chlorpyrifos and nutrient-rich chlorinated effluent had no
detectable effect (Pratt et al. 1993).
2.4 Effects of Acidification
Microbes are highly sensitive to pH (James 1991). Taxa richness and species
composition appear to be more sensitive to acidity than biomass and net oxygen
metabolism. Bacterial abundance and richness of protozoan assemblages has been shown
to decline below a pH of 5.3, and shifts in taxa composition can occur at higher pH's
(Niederlehner & Cairns 1990). Slow rates of plant litter decomposition can indicate
acidic conditions, as noted in a West Virginia wetland (Kittle et al. 1995). Wetlands with
low acidity and high salinity are generally at greatest risk for outbreaks of Clostridium
botulinum, the cause of avian botulism poisoning (Rocke and Samuel 1999).
Acidification of soils can diminish populations of mycorrhizae that are crucial to plant
nutrition (Hutchinson et al. 1999).
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Occasionally, some microbial assemblages respond positively to moderately acidic
conditions (Brenner 1995). Acidity enhances the availability of metals that are a source
of energy for many bacteria. Wetlands that have been acidified, or which naturally have
more acidic conditions, support a characteristic assemblage of iron-oxidizing bacteria
(Emerson 1999). Dissolved carbon levels, which often parallel acidity levels, can
influence microbial population densities even more than can acidity (Fisher et al. 1991).
2.5 Effects of Salinization
Diverse bacterial assemblages are key to cycling of energy in many naturally saline
wetlands (Zahran 1997). Bacterial mineralization (i.e., increased bioavailability) of
dissolved organic nitrogen declines as salinity increases along a freshwater to seawater
gradient (Stepanauskas etal. 1999). Decomposition of cellulose gradually declines along
a salinity gradient, although at high salinity (seawater strength) decomposition can
increase greatly (Mendelssohn etal. 1999). Salinities of greater than about 300 g/L can
inhibit the ability of microbial assemblages to detoxify toxic forms of selenium
(Steinberg & Oremland 1990).
2.6 Effects of Temperature
Methanogenesis by microbes is much greater at warmer temperatures (Boon and Mitchell
1995, Sorrell etal. 1997), as is microbial degradation of some detergents, e.g.,
alkybenzenesulfonates (Inaba 1992). Microbial communities in wetlands containing hot
springs exhibit a tremendous degree of genetic diversity (Hugenholtz et al. 1998).
2.7 Effects of Sedimentation, Burial
Fine sediments generally contain larger populations of microbes than do coarse
sediments, because they provide additional surfaces for attachment and protection from
predators (Davies & Bavor 2000). However, application of unenriched and phosphorus-
enriched fine clay suppressed decomposition of Sparganium eurycarpum decay by about
6-8% over 117 days (Vargo et al. 1998). A single application of coarse sediments to cat-
tail (Typha latifolia) litter inhibited decomposition by 10% over 470 days.
2.8 Effects of Vegetation Removal, Shade
Bacterial biomass production was unaffected by light intensity under the conditions used
in one series of laboratory experiments (Neely and Wetzel 1997). Fungal biomass was
found to be greater than bacterial biomass throughout the year in an Oregon riparian
system (Griffiths et al. 1997). Colonization of wet prairie herbs by fungal mycorrhizae in
Oregon was not significantly correlated with prior type of intensity of land use (Ingham
& Wilson 1999). Widespread removal of host plants by prolonged flooding reduced the
extent of colonization of remaining plants by mycorrhizae (Ellis 1998).
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Decomposition of various wetland macrophytes did not differ significantly among grazed
and ungrazed sites in a German river delta (Ibanez et al. 1999), but mowed sites had less
colonization by mycorrhizae (Titus & Leps 2000). Decomposition of cotton strips was
significantly greater following logging of a Michigan forested wetland (Trettin and
Jurgensen 1992). Removal of vegetation and the upper soil layer from a Florida
Everglades site resulted in increased activity of microrrhizal fungi for at least 2 years
after disturbance (Aziz et al. 1995). However, removal of streamside vegetation in
Illinois resulted in fewer number of decomposer fungal species per unit substrate and
reduced colonization of substrates (Metwalli & Shearer 1989). Denitrification rates also
have been found to differ between grass vs. forested riparian areas (Groffman et al.
1991).
2.9 Effects of Dehydration, Inundation
Changes in water volume in wetlands have little effect on individual bacteria directly, but
resulting changes in nutrient levels and salinity, brought on by dilution or concentration
of elements, can cause profound changes. Bacterial densities in wetlands often increase
shortly after runoff events that bring in nutrients (White et al. 1998). Following the end
of a drought, bacterial densities in a clearwater lake exposed to acidic precipitation were
unchanged, whereas in a darkwater (naturally stained) lake, bacterial densities declined
possibly as the result of interactive effects of high dissolved organic carbon
concentrations and reduced pH (James 1991). Leaf litter decomposed faster, and nitrogen
was mineralized more readily, in a Louisiana crayfish pond where flooding was
manipulated than in natural and impounded wetland forests (Conner and Day 1991).
Decomposition rates were greater in drained than undrained pocosins (peat wetlands) in
North Carolina (Bridgham et al. 1991). In Manitoba, decomposition rates were found to
be higher in a flooded treatment than in the control treatment for some plant species (van
der Valk et al. 1991), but not others (Wrubleski et al. 1997).
Flooding of wetlands stimulates microbial activity that potentially increases conversion
of inorganic mercury to the much more toxic methyl mercury form (Kelly et al. 1997,
Heyes et al. 1998). Either increases or decreases in wetland water levels, if occurring
over sustained periods, can increase bacterial release of methane from organic sediments
(Freeman et al. 1996, Brown & Clair 1998). However, lowering the water level below
the sediment surface in part of the Florida Everglades caused the wetland to shift from a
methane source to a methane sink (Happell et al. 1993).
Although effects of changing water levels on denitrification have not been studied in
prairie wetlands, two recent landscape-scale studies of Saskatchewan fields (Elliott and
de Jong 1992, van Kessel et al. 1993) highlight the key role of soil moisture. Soil water
content was found to be the most dominant factor controlling denitrification
activity, followed by the concentration of ammonium, total soil respiration, and
nitrate (van Kessel et al. 1993). Microbial assemblages that support
denitrification develop rapidly in newly created wetlands (Duncan and Groffman
1994).
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Some studies suggest that colonization of plant roots by vesicular arbuscular mycorrhizal
(VAM) fungi can be influenced by wetness (Cooke et al. 1993, Cantlemos and Ehrenfeld
1999), whereas others suggest it is only mildly influenced by wetness (Aziz et al. 1995,
Turner et al. 2000). The species composition of VAM fungi was found to vary only
slightly by water depth when plant species was held constant (Miller & Bever 1999). No
VAM species was confined to the wettest parts of two wetlands that were studied in
Florida (Miller & Bever 1999). VAM colonization has found to be less where surface
water depth and persistence are great (Rickerl et al. 1994,Wigand et al. 1998, Miller &
Sharitz 2000, Miller 2000).
2.10 Wetland Monitoring
Spatial and Temporal Variation
Bacterial abundance and productivity can vary more than an order of magnitude over an
annual cycle. Bacterial abundance during warmer periods can be measurably affected by
protistan herbivory, although much of this herbivory is respired (Johnson and Ward
1997). In an Alabama wetland, bacterial productivity per mg dissolved organic carbon
(DOC) in spring decreased compared to winter, while dissolved organic carbon
concentrations increased over this period (Mann 1998). In an Ohio fen, microbial growth
rates and cell density peaked in late July and then decreased until at least December
(Gsell et al. 1997). Surface DOC within a southern wetland varied seasonally, with
greatest fluctuations in concentrations through the summer and autumn during intensive
macrophyte growth and bacterial production (Mann and Wetzel 1995). Methane
emission in a Maine peat bog reached its lowest point in winter (Roslev and King 1996).
Methane production in organic-rich wetlands is related more to organic-chemical
components of the peat than to the activity of coincident sulfate-reducing bacteria (Yavitt
and Lang 1990). In salt marshes, the species composition of diazotrophic bacteria
assemblages (which provide nitrogen to plant roots) despite large acute variations in
available carbon (Piceno & Lovell 2000b). Bacterial and fungal populations can occur in
deep alluvial sediments over 1 mile from a river channel, but are greater closer to the
channel (Ellis et al. 1998). Spatial variation in nitrification rates in a wetland was mostly
associated with differences in emergent plant species composition (Eriksson &
Andersson 1999). One study found substrate type to have a greater influence than local
geography on microbial taxonomic composition (Goodfriend 1998).
Techniques and Equipment
Because of the highly dynamic nature of microbial assemblages, obtaining samples that
are spatially and temporally representative of the community's taxonomic composition
and density can be a daunting or impossible task (Kinkel et al. 1992). Instead, diversity
of functional processes is often measured with much less effort. Or, microbial taxa can
be grouped by presumed physiological tolerances, nutritional versatility, genetic
distinctiveness, or other factors, prior to analyzing the data (e.g., Atlas 1991). A
technique using rRNA-targeted oligonucleotide probes was found to be reliable for
characterizing functional composition of microbes in salt marsh sediments (Edgecomb et
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al. 1999, Frischer et al. 2000). A method for assessing populations shifts within
functional groups of microbes (such as change in proportion of nitrogen-fixing microbes)
in response to pollution also is described by Chelius & Lepo (1999). Isotopes are
increasingly being used to identify key sources of nutrition for microbial assemblages in
wetlands, thus better defining the temporally and spatially variable roles of wetland algae
vs. macrophytes (Creach et al. 1999). Microbial taxa responsible for sulphate reduction
and methane oxidation can be identified by use of stable carbon (13C) isotopes (Boschker
etal. 1998).
To characterize wetland microbial diversity according to physiologic characteristics, API
and BIOLOG tests can be performed on samples (Bagwell et al. 1998). Also, DNA,
protein, and lipid synthesis can be assessed by measuring thymidine, leucine, and glucose
incorporation (Gsell et al. 1997). A review of commonly used microbial toxicity tests
(Van Beelen & Doelman 1997) concluded that (1) respiration rate per unit of biomass is
a more sensitive indicator of toxic effects than the respiration rate or the amount of
biomass alone, (2) autotrophic nitrification and acetylene reduction tests can be sensitive
to toxics when short incubation times are used, (3) the nitrogen mineralization,
denitrification and many enzymatic tests are often not very sensitive to effects of toxics,
and (4) urease activity is a relatively sensitive enzymatic test for contaminants. Much
information on microbiological techniques is summarized by Maier et al. (2000).
-------
2.11 Literature Cited
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-------
Chung, KH; Ro, KS; Roy, D. 1996. Fate and enhancement of atrazine biotransformation in anaerobic
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Section 3: Algae
3.1 Use as Indicators
From a habitat perspective, algae are commonly grouped as phytoplankton (algae suspended in
the water column), metaphyton (unattached and floating or loosely associated with substrata),
benthic algae (attached to substratum), and epiphytic algae (attached to plants). Like the other
sections in this document, this section focuses almost entirely on research published since 1989.
For a general discussion of the topic based on pre-1990 scientific studies, and for broader
discussion including advantages and disadvantages of using algal assemblages and processes as
indicators of wetland integrity, readers should refer to Adamus and Brandt (1990) and to EPA's
web pages on use of algae in wetland monitoring: http://www.epa.gov/owow/wetlands/bawwg.
As a source of energy for invertebrates and higher trophic levels, algae often are more important
than vascular plants, at least during the summer when temperature and light conditions are most
favorable (Hanson & Butler 1990, Hargeby et al. 1991, Neill & Cornwell 1992, Murkin et al.
1992, Peterson & Deegan 1993, Vymazal 1994, Campeau et al. 1994).
In the past decade, studies that have used assemblages of algae specifically to indicate condition
of a large series of wetlands have been conducted, for example, in Montana (Apfelbeck 1998),
the Midwest (Mayer & Galatowitsch 1999), and southwestern Maine (inpreparation). Some of
these studies are detailed at http://www.epa.gov/owow/wetlands/bawwg/case.html.
Of particular note is the book Algae and Element Cycling in Wetlands (Vymazal 1994), and the
book Algal Ecology (Stevenson et al. 1996), especially the chapter on wetland algae by
Goldsborough and Robinson (1996). Also, literature specifically on algae in prairie wetlands
was reviewed by Crumpton (1989) and Adamus (1996). Information on use of algae in the
development of biocriteria, and/or on the tolerances of many algal taxa to eutrophication,
acidification, and/or herbicides, is discussed or compiled by Shubert 1984, Dixit et al. 1991,
Whitton et al. 1992, McCormick and Cairns 1993, Dixit and Smol 1994, Harper 1992, van Dam
etal. 1994, Patrick andPavalage 1994, Rosen 1995, Stevenson et al. 1996, Pan & Stevenson
1996, McCormick & Stevenson 1998, and Stevenson et al. 1999.
3.2 Effects of Enrichment, Eutrophication, Reduced Dissolved Oxygen
Processes
Algal production is often constrained by one or two macronutrients (Hannson 1992), most
notably phosphorus and nitrogen. For example, Everglades wetlands are naturally limited by
phosphorus (Browder et al. 1994, Pan et al. 1999, 2000), whereas arctic systems respond to
either nitrogen, phosphorus (Lock et al. 1989) or both (Bowden et al. 1992). A number of
studies involving experimental dosing of actual or simulated wetlands with nutrients have been
conducted in the past decade, and algal responses have been monitored (Table 3.1).
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Table 3.1. Examples of nutrient response studies involving wetland algae
STUDY
SITE
Lab
2mX19cm
Tundra
stream
Marsh
DURATION
61 days
6 years
2 months
NUTRIENT
CONCENTRATION
0,0.1,0.5,1,2,5, 10,
20, 50, 100 (og/L PO3
10 (og/L P, 100 (og/L N
2,3,5,6,7,27,51,76
REFERENCE
Both-well 1989
Bowdene/a/. 1992
Grimshaw et al. 1993
Everglades 5 months
Slough
112 (og/L Dissolved P
0.4,0.8, 1.6,3.2,6.4,
12.8 (og/L NaH2PO4
McCormick and O'Dell 1996
Whereas in some lakes and streams algal mats indicate eutrophication, in the Florida Everglades
algal mats with Utricularia spp. are viewed as indicators of health (McCormick and Stevenson
1998, Craft et al. 1995, Rader and Richardson 1992). With increasing nutrient loading, however,
the polysaccharides that hold algal mats together disintegrate (McCormick et al. 1997,
McCormick and Stevenson 1998, Craft et al. 1995, Rader and Richardson 1992). Although the
mats themselves dissipate, the species responsible for the polysaccharides typically remain,
unless affected by other variables (Rader and Richardson 1992). In some cases where nutrient
loading continues, desmid species that construct the mats are replaced by more nutrient tolerant
species. Craft et al. (1995) found that as algae mats dissipated, Chora spp. became dominant.
Effects on Species Richness
Eutrophication can lead to the simplification of algal communities, especially those already in a
mesotrophic environment. In streams, diatom communities respond to organic enrichment with
decreased species richness, diversity, and evenness (Steinman and Mclntire 1990). When
phosphoric acid was added to an Alaskan river, for example, species diversity and evenness
declined (Bowden etal. 1992).
Effects on Species and Functional Group Composition
Generally, phytoplankton respond quickly to small, repeated additions of nutrients (Jorgensoner
al. 1992, O'Brien etal. 1992, Olssonetal. 1992, Gaboretal. 1994), as does epiphyton (Wetzel
1990) . However, phytoplankton may respond more strongly than epiphyton to a single dose of
nutrients added in the spring (Gabor et al. 1994). Metaphyton respond even more slowly, but
effects are more enduring. At least in deeper wetlands and lakes, the response of benthic algae to
nutrients is relatively muted (Murkin et al. 1994), but nonetheless measurable (Lowe and Pan
1994).
Particular algal taxa seem to respond more quickly than others to nutrient enrichment. For
example, Chlorella, Spirogyra, andAnabaena are often found in nutrient-rich waters, whereas
Cladophora andAnkistrodesmus sometimes occur in less eutrophic situations (Hosseini and van
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der Valk 1989a). When phosphoric acid was added to an Alaskan river, the dominant tax on
Hannea arcus declined, as did Fragillaria., whereas numerous taxa belonging to Achnanthese,
Cymbella and Eunotia increased (Miller et al. 1992). Among 30 Canadian lakes, total
phosphorous and water transparency (Secchi depth) were found to explain a large proportion
(85%) of the species variance (Agbeti 1992).
Among wetlands of the United States, algal species response to nutrients has probably been best
studied in the Florida Everglades. During a nutrient addition experiment there, periphyton
declined strongly after one year of treatment, and their species composition shifted (McCormick
et al. 1997). Chlorophytes (filamentous green algae) and diatoms replaced the normal
assemblages of cyanobacteria (blue-greens), whereas vascular plants showed no response (Rader
& Richardson 1992, Daoust and Childers 1997).
The decline in the Everglades specifically of periphyton was also documented by Jensen etal.
(1995), and its cause has been attributed primarily to phosphorus rather than nitrogen (Vymazal
et al. 1994). Along a 14 km nutrient gradient in the Florida Everglades, shifts in periphyton
community composition were related strongly to decreasing distance from canal discharges and
to increasing total phosphorus (TP) ranging from 10 (8 km from canal discharges) to 150 (ig/L
(near discharges) (McCormick etal. 1996). In particular, diatom species indicative of low TP
(e.g., Anomoeoneis vitrea, Mastogloia smithii) were consistently replaced by eutrophic indicator
species (e.g., Gomhonemaparvulum, Nitzschia amphibia) at TP concentrations between 10 and
20 ng/L. Another study along the same gradient found declines in the relative abundance of five
oligotrophic indicator species (Scytonema hoffmanii, Shizothrix calcicola, Oscillatoria limnetica,
Cymbella lunata and Mastogloia smithii) at 10 |j,g TP/L (McCormick and O'Dell 1996). The
abundance of Anomoeoneis serians displayed a less severe decline at this TP level.
Cyanobacteria/diatom assemblages dominated stations with less than 10 |j,g TP/L, whereas
intermediate TP loading levels (3.2-6.4 g P/m2/yr) encouraged the establishment of floating
Spirogyra sp. mats. As TP levels increased to approximatey 40-50 (ig/L other taxa became
dominant including Nitzschia sp., Lyngbya aestuarii and Oscillatoria minima. Independently,
Raschke (1993) found similar shifts in taxonomic composition in the Everglades below 20 |J,g/L
TP.
The exact nutrient that contributes to algal community shift is often difficult to identify due to
correlations among many nutrients (McCormick and O'Dell 1996). Nonetheless, some studies
(Harper 1992) have reported that diatoms seem to dominate at lower temperatures and when
phosphorus (P) but not silica (Si) is limiting, whereas green algae may dominate at higher
temperatures with moderate or lowN:P and Si:P ratios; Cyanobacteria (blue-green algae)
typically dominate at higher temperatures and at low N:P ratios, and often characterize highly
enriched waters (Hughes and Paulsen 1990, Murkin et al. 1991, Murkin et al. 1994, Biggs 1995).
In the Florida Everglades, phosphorus has the largest impact on algal assemblages, followed by
nitrogen (N) and iron (Fe) (McCormick & O'Dell 1996, McCormick et al. 1998). In enriched
areas of the Everglades, nitrogen, other nutrients, and/or light play a larger role in limiting
growth (Vaithiyanathan and Richardson 1997, McCormick and Stevenson 1998). In nitrogen
limited areas, increases in Rhopalodia gibba and blue-green algae (Nostoc) and typical
(Vaithiyanathan and Richardson 1997).
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Other chemical constituents of water, particularly pH-altering bicarbonates, can regulate the
response of algae to nutrient additions (Fairchild and Sherman 1993). In a Canadian softwater
oligotrophic lake, Stigeoclonium sp., Scenedesmus sp., Cryptomonas sp., Euglena sp. and
Rhodomonas sp. increased in relative abundance when carbonate ions were naturally abundant
during the addition of nitrogen (.15 mol NaNOs) and phosphorus (.015 mol Na2HPC>4), whereas
Mougeotia sp., Oedogonium sp., Nostoc sp. andAnacystis sp. decreased (Fairchild etal. 1989,
1989a). Species responses varied depending on whether nitrogen, phosphorus or bicarbonate
were supplied, indicating that limiting nutrients were species specific. Excessive concentrations
of nitrogen, in the form of ammonia - can be directly lethal to algae. In Michigan, ammonia
contaminated sediments (1.3-54.4 mg/L ammonia) from 13 sites were acutely toxic to
Selenastrum capricornum fAnkley et al. 1990). In addition, physical factors play a role in the
response of alge to nutrients. In Pool 8 of the Mississippi river — a predominantly lacustrine
system with various lentic and slow water lotic microhabitats — algal communities appear to be
shaped by water flow, rather than strictly regulated by nutrient levels (Lange and Rada 1993).
Shifts in nutrients can alter macroinvertebrate populations, leading to a change in the balance
between those that consume algae vs. those that consume vascular plants. This shift in trophic
levels can mask the effects of nutrient additions (Irvine et al. 1989). Enrichment also can shift a
stable epiphyton-dominant system ("open wetland") to metaphyton dominance ("sheltered
wetland") if macrophytes remain sufficiently abundant to provide a substrate for metaphytic
algae (McDougal et al. (1997). Development of the phytoplankton-dominant "lake wetland"
state presumably occurs only when other algal and macrophytic competitors for nutrients are
few. Metaphytic algae may increase in dominance due to their limited palatability to micro and
macroinvertebrates (Neill and Cornwell 1992).
The use of algal species composition to predict ambient TP concentrations is limited by
extensive variation in time and space of the TP concentrations (Chambers et al. 1992, France and
Peters 1992). Nutrients appear to vary even more than pH (Battarbee 1990, 1999).
Effects on Density or Biomass
In response to enrichment, algal biomass increases quickly (Humphrey and Stevenson 1992,
Dodds et al. 1998) — more quickly than does biomass of submerged vascular plants (Crumpton
1989, Klarer and Millie 1992). For example, in a lacustrine wetland in Manitoba, algal biomass
and density were strongly correlated to the degree of wetland enrichment (Murkin et al. 1991b).
In the Florida Everglades, phosphorus concentrations in periphyton corresponded closely with
concentrations in the water across the range of 3 to 112 |j,g TP/L, implying substantial uptake of
phosphorus by the algae (Grimshaw et al. 1993). Uptake by Everglades epiphyton, vascular
plants, and sediments resulted in a ten-fold decline in ambient phosphorus levels (Grimshaw et
al. 1993).
The increase in algae that results from an increase in nutrients depends largely on the intensity of
algal grazing by invertebrates (Harris 1996, Allan 1995, Bourassa and Cattaneo 1998, Mazumder
et al. 1989, Paul et al. 1989, Mulholland et al. 1991). For example, in open water systems, large
cladocerans can suppress algal biomass to low levels despite excess nutrients. Additions of
phosphorus to an arctic river resulted in a net decrease in epilithic algae due to an extreme
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increase in grazers (Miller et al. 1992). Similarly, when nutrients were added to a Canadian
marsh, net algal production (predominantly epiphyton) declined due to grazing from
invertebrates (Hann and Goldsborough 1997). However, in a series of ponds treated with
phosphate, algal biomass was higher - despite the abundance of a snail — than those where
nitrate was added (McCormick and Stevenson 1989).
Shading also influences the magnitude and type of response of algae to nutrients. In 11
Canadian lakes, epiphyton biomass increased up to 39 |j,g total phosphorus/L, after which it
declined due to shading from phytoplankton (Lalonde and Downing 1991). Metaphyton also
create shade, and respond quickly to phosphorus additions (e.g., 179 mg-P/m3 )(Wu and Mitsch
1998).
3.3 Effects of Contaminant Toxicity
Processes
Much of the literature on algal sensitivity to chemical contaminants, as revealed by experimental
dosing (phytotoxicity testing), is summarized by Lewis (1995). Many algal species are more
sensitive than vascular plants to contaminants, especially contaminants that interfere with
photosynthesis. Common pollutants often elicit responses from algae that are well below the
levels that affect vascular plants and other groups (Niederlehner and Cairns 1990, Lewis et al.
1998). Algal sensitivity to contaminants can also be less in eutrophic environments (Chen 1989,
Wangberg andBlanck 1990).
Prior exposure to a contaminant can alter the response of individual cells to a new contaminant
(Niederlehner and Cairns 1990), and consequently entire populations can develop resistance to a
chemical after chronic exposure (Blanck and Wangber 1988). Algal communities can recover
from exposure to toxic pollutants by means of their physiological or genetic adaptations to
current conditions, or by undergoing a slow successional change after the disturbance ends,
which sometimes returns the community to its predisturbance condition (Steinman and Mclntire
1990).
Effects on Biomass
If algal biomass is the only metric considered, then some studies suggest that the stimulatory
effect of compounds remaining in treated wastewater mixtures sometimes outweighs the toxic
effects. For example, Lewis et al. (1998) examined the response of Selenastrum capricornutum
(a freshwater green algae) to in vitro additions of different effluent from two cities, one naval air
station, three forest product plants, two agro-chemical industries, one synthetic fibers industry,
and one steam power generation plant. Algal biomass was stimulated in all but one forest
product effluent and one city effluent.
Effects on Species Composition, Growth
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Herbicides often cause a shift from large, filamentous green algae (chlorophytes) to smaller
diatoms and cyanobacteria species (Gurney and Robinson 1989). Ironically, algal blooms can
occur in wetlands after herbicides are applied to kill vascular plants, because a reduction in shade
from vascular plants can trigger increases in benthic algae and metaphyton (Adamus 1996).
After testing 20 herbicides, 2 insecticides, and one fungicide on many algal species, Peterson and
others (1995) reported adverse effects from 9 of the pesticides (particularly 5 triazine herbicides).
The fungicide propiconazole and the herbicides picloram, boromoxynil, and triclopyr were
relatively harmless to algae. In another study, stream periphyton did not appear to be adversely
impacted by a 12-hour exposure to hexazinone (Kreutzweiser et al. 1995). The insecticide,
Fenitrothion™, caused significant decreases in growth among 12 phytoplankton species (Kent et
al. 1995). Atrazine and bifenthrin had deleterious effects on algal populations (Hoagland et al.
1993).
Toxic levels of copper, lead, and zinc have been documented to cause a decline of many algal
species and an increase of Rhizosoenia eriensis in a contaminated lake (Deniseger et al. 1990), as
well as affect algal metabolic processes (Hill et al. 1997). Phytoplankton from an Ohio lake
were more sensitive to copper toxicity during the summer and fall than in the spring, except for
Crythophyta which was strongly sensitive to copper throughout the year (Winner and Owen
1991). Polynuclear species of aluminum may be very toxic to algae and may represent a
significant proportion of the aqueous aluminum at some conditions of low pH (Hunter and Ross
1991). However, in a stream experiment the addition of aluminum increased densities of
diatoms, green algae, and blue-green algae (Genter and Amyot 1994). Specifically, the diatom
Achnahes minutissima, the green alga Cosmarium malanosporum, the filamentous blue-green
alga Schizohrix calcicola, andNavicula sp. diatoms all experienced increased growth in response
to elevated aluminum (Genter 1995). Another study found Navicula sp. to be unaffected by the
addition of aluminum (Planas et al. 1989).
3.4 Effects of Acidification
Processes
Algae are affected by acidification as a result of (a) direct toxicity (Baker and Christensen 1990,
Fairchild and Sherman 1993), (b) indirect toxicity, from some metals that are mobilized or made
more available by changes in acidity (Genter and Amyot 1994, Schindler 1990, Kingston et al.
1992), and (c) changes in competition with, and predation from, organisms that are less sensitive
(Elwood and Mulholland 1989, Schindler 1990, Locke and Sprules 1994, Feminella and
Hawkins 1995). Changes resulting from algal sensitivity to acidification can be traced through
entire food webs (Havens 1992).
Effects on Species Richness
Either extreme of acidity (too acid or too basic) can dimish species richness of algal
communities. Among 36 lakes in the Upper Midwest, diatom community diversity and richness
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were lowest in the most alkaline, plankton-dominated lakes and in lakes dominated byMelosira
sp. (Cook and Jager 1991).
Effects on Species Composition
Logically, the losses of algae in acidified waters are proportionally greater for acid-intolerant
species (Havens and Heath 1990, Pillsbury and Kingston 1990). Species composition is perhaps
a better indicator of acidification than is biomass (Dixit and Smol 1989).
Acidification, especially when accompanied by increased availability of aluminum, can result in
an increase in green algae (e.g., Mougeotia, Temnogametum) and decrease in cyanobacteria (e.g.,
Lyngbya andAnabaena) and diatoms (e.g., Achnathes) (Turner et al. 1991, Dixit et al. 1991b).
In a series of lake acidification experiments, green algal mats greatly increased as pH decreased
from 6.0 to 5.1 (Baker and Christensen 1991). Mougeotia sp. also grew extensive mats in a
stream experiment that maintained a pH of 5.0 for extended periods of time (Elwood and
Mulholland 1989).
During acidification, the phytoplankton community often shifts from smaller to larger types with
an accompanying increase in transparency of the water (Havens 1991b). Long, thin, pennate
diatoms like Nitzschia gracilis, Syndera rumpens, S. delicatissima are exposed to acidity to a
greater degree due to their larger surface area, whereas smaller diatoms (e.g., Achnathes
minutissima, Coccneisplacentula, Nitzschia fonticola, and TV. lacuum) are presumably less
exposed to acidity and associated aluminum (Genter and Amyot 1994). Phytoplankton
associated with acidic wetlands include the dinoflagellates, Peridinium inconspicuum and P.
limbatum (Baker and Christensen 1991). Under acidic conditions, Merismopedia, Peridinum,
and Gymnodinium are common species (Dixit and Smol 1989). Also, as acidification progresses,
planktonic algae can become less dominant andbenthic algal species can increase (e.g.,,
Mougeotia sp. and Spirogyra sp.; Dixit and Smol 1989).
Diatoms in particular are sensitive indicators of acidity, as shown by studies where they
successfully predicted acidity of New York lakes (Dixit et al. 1992, Dixit et al. 1993, Dixit et al.
1999), southeastern Ontario lakes (Christie & Smol 1993), Montana lakes (Charles etal. 1996),
Montana wetlands (Apfelbeck 1998), streams in the mid-Atlantic Highlands ecoregion (Pan et al.
1996), Chesapeake Bay waters (Cooper 1995), and in other regions (e.g., Birks etal. 1990,
Anderson et al. 1993, Battarbee et al. 1999). Although pH appears to a be an important factor in
determining the diatom composition, other factors such as dissolved oxygen concentrations,
nutrient loading, metal and micronutrient availability, and lake size and shape are also important
(Cook and Jager 1991).
Effects on Biomass, Growth
Algal biomass in a stream declined as pH dropped from 6.5 to 4.5, and the community
composition shifted to larger cells that were less productive (Havens 1992). The combination of
acid and aluminum is much more harmful to growth of diatoms and cyanobacteria than acid
alone. Growth of Schizohrix calcicola, a filamentous cyanobacteria, was inhibited by all
aluminum treatments but not inhibited by acidic conditions (Genter and Amyot 1994). Some
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species, such as Rhopalodia gibba and Synedra rumpens, tend to decrease in abundance with
more acidic conditions but do not decrease further with the addition of aluminum (Planas et al.
1989). Filamentous cyanobacteria (e.g., Schizohrix calcicola) increase under acidic conditions
but decrease under a combination of elevated acidity and aluminum (Dixit et al. 1991b).
3.5 Effects of Salinization
Algal species are sensitive indicators of salinity and conductivity in surface waters and sediments
of lakes and wetlands. Salinity ranges associated with the occurrence and greatest density of
particular species are described by Blinn (1993) and Fritz et al. (1993). Studies of saline lakes in
western North America found that the greatest richness of diatom taxa occured with specific
conductance of less than 45 mS (Blinn 1993), although light, temperature, and nutrient levels
may influence response to salinity in prairie wetlands (Robarts et al. 1992). Among some
lacustrine wetlands in Wyoming, relative production of different algal types (epiphyton,
epipelon, phytoplankton) varied depending on salinity and associated floating macrophytes
(Chora at low salinity, Potomogeton at higher), but total algal production did not. At lower
salinity, epiphyton predominated whereas at higher salinity phytoplankton and benthic algae
assumed a larger proportion of the total production (Hart & Lovvorn 2000). In western
Kentucky wetlands that received acid mine drainage, conductivity — not pH — was a stronger
determinant of diatom assemblages (Pan & Stevenson 1996).
3.6 Effects of Temperature
Apparently no studies have been published in the last decade regarding response of algae to
temperature changes in wetlands. Algal biomass generally increases with rising temperature and
duration of growing season.
3.7 Effects of Sedimentation, Burial
The impacts of sedimentation on algae have not been documented within the last decade in
wetlands. Attached algae such as periphyton and benthic algae are probably affected
disproportiately.
3.8 Effects of Vegetation Removal
The removal of wetland vascular plants will obviously cause a decrease in the abundance of
algae associated with these plants, i.e., epiphytic algae (Murkin et al. (1994). Epiphyton biomass
varies by vascular plant species. For example, epiphyton biomass was found to be higher on
Polygonum sp. than on Typha sp. (Cronk and Mitsch 1994a, b). Vegetation removal also can
have a cascading trophic effect. In a pond experiment in California, the removal of vegatation
resulted in growth of predaceous beetles which reduced chironomid populations. This reduction
in grazers allowed the periphyton population to grow independently of herbivore pressures
(Batzer and Resh 1991).
3.9 Effects of Turbidity, Shade
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Shade tolerance differs among algal species, allowing some species to thrive at the expense of
others (Steinman and Mclntire 1990, Steinman et al. 1990). In the Florida Everglades, denser
stands of vascular plants had less epiphytic algae due to growth limitation by shade, and
consequently removed less phosphorus (but more nitrate) than more open stands (Grimshaw et
al. 1997). Algal responses to nutrient additions can be confounded by shading (Stevenson et al.
1991). Tannins produced by some wetland plants under acidic conditions can also stain waters
to such a degree that light transmissivity is reduced, with consequent reductions in algal growth
and changes in species composition (Graham 1989, Goldsborough and Brown 1991).
Turbidity-related shading probably impacts benthic algae more than phytoplankton (Mitsch &
Reeder 1991). In a Canadian marsh, shade from metaphyton reduced growth of submerged
macrophytes and presumably their associated epiphytic algae; the metaphyton had expanded in
response to nutrient additions (McDougal et al. 1997). Despite the shading out of some species,
primary production increased two-fold due to the large amount of metaphyton (Hann and
Goldsborough 1997). Shade from high densities of nannoplankton, such as occur when
planktivorous fish are present, can also reduce the amount of benthic and epiphytic algae
(Mazumder et al. 1989). Some prostrate diatoms, such asAchnanthes lanceolata and Cocconeis
placentula, appear to gain competitive advantage under conditions of low light (Steinman and
Mclntire 1990). In shallow lakes in which fish increased turbidity by resuspending sediment,
nanochlorophytes replaced centric diatoms and larger diatoms replaced smaller types (Havens
199 la).
3.10 Effects of Dehydration, Inundation
Processes
Dehydration and inundation affect algal communities by altering the available water volume
(Hough et al. 1991), substrate area (Sand-Jensen and Borum 1991), wind mixing (Robinson et
al. 1997b), scouring (Biggs and Close 1989, Stevenson 1990), temperature, and nutrient
concentrations. Flooding also can alter algal communities by dispersing algae and by increasing
turbidity, thus decreasing light penetration. Water level drawdown can concentrate nutrients,
contaminants, and salinity. However, nutrient availability can increase with flooding as well as
with drought (Cronk and Mitsch 1994). Thus, algal response to nutrients in shallowly- or
temporarily-flooded wetlands is more complex than in lakes (Robarts et al. 1995, Moss et al.
1996).
Some algae have adapted to periodic dessication by achieving the ability to produce thick cell
walls, mucilage sheaths, zygospores, or cysts (Steinman and Mclntire 1990). Taxa that
experience frequent drawdowns, such as those in temporary wetlands, are most likely to have
such adaptations, whereas taxa that are normally submerged year-round are more susceptible to
dessication. Some algal taxa found in soils can survive dessication for over 30 years (Trainor
and Gladyeh 1995). Moderate currents and flooding (e.g., 41 cm of water depth per week) can
provide epiphytic algae with nutrients from outside a wetland while removing wastes, thereby
increasing growth (Cronk and Mitsch 1994).
Effects on Species Composition
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The conversion of shallowly-inundated wetlands to deepwater ponds or lakes reduces epiphyton
and metaphyton, and causes algal communities to shift toward a greater dominance by
phytoplankton (Robinson et al. 1997a,b). This can trigger a shift in the invertebrate community,
from detritivores and/or scrapers to filter-feeding taxa. In intermittent bay wetlands in South
Carolina and Georgia, paleolimnological analysis indicated thatEunotia spp., Luticola saxophila,
and Pinnularia borealis var. scalaris were indicators of sustained drying or draw-down; and
other Pinnularia spp. and Stenopterobia densestriata responded positively to persistent ponding
(Gaiseretal. 1998).
Effects on Density, Biomass, Growth
Initially, an increase in water levels in a Manitoba marsh (a 30 to 60 cm total depth increase)
increased the biomass of floating metaphyton, epiphytes, epipelon and phytoplankton. However,
the algal response did not persist (Hosseini and van der Valk 1989a, Robinson etal. 1997a).
3.11 Effects of Other Human Influences
Humans can indirectly alter the species composition of algae by removing or introducing grazing
and planktivorous fish (Hanson and Butler 1990), but this effect is not inevitable (e.g.,
Ramcharan etal. 1995). Algae are also affected when introduction or harvesting offish, or
activities or events, change the density of wetland invertebrates, because grazing invertebrates
significantly reduce algal biomass in wetlands (Hann 1991, Botts 1993). In an experimental
pond system, grazing from macroinvertebrates resulted in a simplified algal community
dominated by Stigeoclonium sp. (Hann 1991). Grazing by invertebrates causes a
disproportionate reduction in more edible algae species, allowing less palatable species (e.g.,
blue-greens) to prosper (Goldsborough and Robinson 1996, Hann and Goldsborough 1997).
Algal communities also can be physically impacted by waves and scouring. Older colonies (e.g.,
12, 18 and 24 days old) are more resistant to scouring than younger (e.g., 6 day old)
communities, which suffered a 47.6% reduction in biomass following disturbance by waves
(Peterson et al. 1990). Resistance to wave scouring was due to stabilization of algal mats by
diatom mucilages and overlying Oscillatoria sp. surface layers. Recovery of communities was
related to differential settling rates of algal species. Fragilaria sp. andNavicula sp. resettled
quickly while Nitzschia sp. were not replaced as readily.
3.12 Wetland Monitoring
Spatial and Temporal Variation
In rivers, algae sampled using artificial substrates appear to integrate the river environment fairly
well. For example, in one river system the assemblage of algal species that colonized artificial
substrates was remarkably similar among samples collected from 1st through 5th order channels
(Molloy 1992). However, in wetlands the density and species composition of algal communities
often varies over the scale of a few meters horizontally and a few centimeters vertically. These
fine-scale spatial patterns, which confound attempts to characterize entire wetlands or even
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particular habitats within wetlands (Morin and Cattaneo 1992), are sometimes driven by spatial
patterns in nutrient inputs (e.g., Wu andMitsch 1998, McCormick 1998) or physical habitat
structure (Rose and Crumpton 1996).
Temporal variation is also important, at both a seasonal and annual scale (Whitton et al. 1998).
In a riverine study of 186 algal species, the possible effects of water quality were subsumed by
other differences between years (Miller et al. 1992). Within years, many water bodies have
predictable, seasonal shifts in the community structure and spatial distribution of phytoplankton
and other types of algae (Harper 1992, Cloern et al. 1992), and these may be altered by nutrient
additions (Gabor et al. 1994). Diatoms are often more common in the spring and summer, green
algae in the summer, and cyanobacteria in late summer (Harper 1992). Seasonal variation in the
relative abundance of algae species can be dampened by eutrophication (McCormick et al.
1997). Oligotrophic Everglades marshes sampled during wet and dry seasons exhibited seasonal
variation in diversity, with cyanobacteria (e.g., Chroococcus turgidus, Scytonema hofmannii)
dominant during the wet season, and diatoms (e.g. Amphora lineolata, Mastogloia smithii)
during the dry season. Eutrophic marshes that were dominated by Cyanobacteria (e.g.,
Oscillatoriaprinceps) and green algae (e.g., Spirogyra sp.) exhibited comparatively little
seasonality.
Equipment and Techniques
Algae are relatively simple to collect and the EPA and others have developed standardized
sampling protocols for streams, rivers, and lakes (Stevenson andBahls 1999):
http://www.epa.gov/owow/monitoring/rbp/ch06main.html.
These protocols will need to be adapted for use in wetlands, as the equipment designed for other
systems is often hindered by wetland conditions. A variety of equipment and techniques are used
in monitoring algal communities in wetlands, and are summarized by Aloi (1990), Adamus and
Brandt (1990), and Adamus (1996). Artificial substrates in the form of glass "diatometer" slides
(McCormick et al. 1996), polyvinyl substrate (Mazumder et al. 1989), open cell styrofoam
(Bothwell 1989), acrylic rods (Hann 1991), and nutrient-diffusing alginate (Gensemer 1991)
have been used, but not without controversy (Aloi 1990). Colonization of natural wetland
substrates also has been monitored (Batzer and Resh 1991). Metaphyton have been sampled by
removing of a portion of numerous mats and combining them into one sample to mediate the
effects of spatial variation (McCormick and O'Dell 1996). Corers are sometimes used to sample
benthic algae (Glew 1991). Phytoplankton are typically sampled with water bottles or other
volumetric samplers.
Measuring algal biomass, volume, and density (e.g., McCormick et al. 1998, Hillebrand et al.
1999) with high precision is time-consuming and probably unnecessary for most assessments of
wetland condition. Formulas have been devised and tested for relating algal volume in wetlands
to simpler measurements of chlorophyll-a (LaBaugh 1995). Measuring algal metabolic activity
also may be useful for some objectives, but requires repetitive measurements of changes in
dissolved oxygen (DO) concentrations or net uptake of radiolabelled CO2 added to water (Wetzel
and Likens 1991, Keough et al. 1993, McCormick et al. 1997). Numerous published estimates of
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algal production in freshwater and saltwater wetlands were compiled by Goldsborough and
Robinson (1996).
Analysis of diatom remains in lake and wetland sediment cores continues to be a promising but
painstaking method for establishing water quality reference (presettlement) conditions, as
discussed by Battarbee et al. (1990), Van Dam and Mertens (1993), and Dixit et al. (1999).
Identification of Taxa
Some of the more recent references for identifying North American algae include Cox (1996) for
diatoms and Dillard (1989a, b, 1990, 1991, 1993, 1998, 2000) for the southeastern United States.
Opinions are mixed regarding the necessity of identifying algal taxa to species or genus. While
finer-level identification is generally desirable, this decision clearly depends on the available
resources and the objective of the study. Identification of algae to only the genus level or coarser
can produce information useful for some objectives (e.g., Prygiel and Coste 1993).
Metrics for Assessing Impacts to Wetland Algal Communities
Metrics used for characterizing algal community response to pollution in streams are described
by Stevenson and Bahls (1999) (http://www.epa.gov/owow/monitoring/rbp/ch06main.html). By
grouping species according to reputed tolerances, much of the statistical variation inherent in
examining single species can be compensated for (McCormick et al. 2000). Use of weighted
average regression analysis can elucidate relationships with specific causative factors, e.g.,
nutrients (Hall & Smol 1992, Line et al. 1994, Lowe and Pan 1996, Stevenson et al. 1999,
Winter & Duthie 2000).
The effectiveness of 5 indices based on diatoms, that are used to assess condition of European
waters, was evaluated by Kelly et al. (1995). In North America, models using algae to predict
total phosphorus and conductivity were developed and tested by Pan and Stevenson (1996). In
general, their model that used planktonic algae was better at predicting conductivity than models
based on epiphytic diatoms. In contrast, the model using epiphytic diatoms was better at
predicting total phosphorus than models based on planktonic diatoms.
In a study of 356 Montana lakes, reservoirs, and wetlands, algal species percent dominance and
community diversity were not successful in reliably differentiating habitats, whereas pH and
conductivity provided a firmer classification (Charles et al. 1996). The usefulness of diatoms
(Dixit etal. 1992, Stoermer & Smol 1999, Slate & Stevenson 2000) and benthic algae (Whitton
et al. 1991, Lowe and Pan 1996) for detecting long-term environmental change has been
described in various types of wetlands. However, in prairie wetlands, Mayer and Galatowitsch
(1999) found diatoms to be a poor indicator of human-related disturbance. This was attributed to
the high degree of natural variability in the sampled wetlands, which was inadequately addressed
by the wetland classification that was employed in sampling. This prohibited statistical
separation of the relative influences of human-related and natural disturbance.
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van Dam, H., A. Mertens, and J. Sinkeldam. 1994. A coded checklist of ecological indicator values of freshwater
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Vymazal, J. 1994. Algae and Element Cycling in Wetlands. Lewis Publishers, Boca Raton, FL.
Vymazal, J.C., Craft, C.B. and Richardson, C.J. 1994. Periphyton response to nitrogen and phosphorus additions in
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Wangberg S.A. and Blank, H. 1990. Arsenate sensitivity in marine periphyton communities established under
various nutrient regimes. Journal of Experimental Marine Biology and Ecology 139: 119-134.
Whitton, B. A., Yelloly, J. M., Christmas, M., and I. Hernandez, I. 1998. Surface phosphatase activity of benthic
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Section 4: Vascular Plants
4.1 Use as Indicators
This section addresses woody and herbaceous plants that grow in non-tidal wetlands. For a
general discussion of the topic based on pre-1990 scientific information, and for discussion of
advantages and disadvantages of using plants as indicators of wetland integrity, readers should
refer to Adamus and Brandt (1990). Also, some of the recent literature on use of wetland plants
as indicators of water quality is summarized, especially for Europe, by Doust et al. (1994).
Vascular plants, also termed macrophytes, are commonly classified as woody or herbaceous.
Woody plants may be classified further as trees, shrubs, or vines. Herbaceous plants (which in
practice sometimes include non-vascular mosses and ferns) may be classified as submersed,
floating-leaved, or emergent. Plants are such obvious components of wetlands, and they are so
sensitive to wetland hydrology, that they are commonly used to delineating wetland boundaries.
A good background on the ecology of wetland vegetation is provided by Keddy (2000) and
Cronk and Fennessy (2001). Information on plants in riparian systems and floodplain wetlands is
compiled and summarized by Malanson (1993), Galatowitsch and McAdams (1994), and Patten
(1998).
Wetland plants, and especially herbaceous species, are increasingly being used as indicators of
wetland ecological condition in North America. The 1990's have seen increased use of plants as
indicators of wetland ecological condition. Following the lead of European botanists (e.g.,
Robach et al. 1996), wetland botanists in North America have increasingly attempted to identify
patterns of plant community response to human-related alteration of streams (Small et al. 1996)
and wetlands (Wilcox 1991). Recently, attempts have been made to develop and/or test
multimetric indices for the purpose of estimating condition of wetlands, e.g., in North Dakota,
Minnesota, Ohio, Pennsylvania, Delaware, Oregon, and perhaps elsewhere. These indices are
often termed "floristic indices" (e.g., Andreas & Lichvar 1995). As a partial starting point for
such indices, Adamus and Gonyaw (2000) compiled literature and prepared a documented
species database for EPA's internet web site, that categorizes many wetland species as tolerant or
intolerant, with regard to overall sensitivity, and/or specific sensitivity to excessive nutrients and
hydrologic alteration: http://www.epa.gov/owow/wetlands/bawwg/publicat.html
In the past decade, several studies have used plant assemblages specifically to indicate the
ecological condition of a large series of wetlands. Results have been published from such
studies, for example, in the Seattle area (Cooke & Azous 2000), Massachusetts (Carlisle et al.
1998), Montana (Apfelbeck 1998), Minnesota (Gernes and Helgen 1999), and western Oregon
(Magee et al. 1999, Adamus 2001). Most of these studies are detailed at:
http://www.epa.gov/owow/wetlands/bawwg/case.html.
Also, in Minnesota wetland plants were used to represent the condition of landscapes that
contained a large wetland/ riparian component (Galatowitsch et al. 1998, Mensing et al. 1998).
Sampling 15 wetlands belonging to each of 8 wetland types, the investigators found positive or
negative correlations of the several metrics with a site disturbance score and/or various land cover
types measured within 500, 1000, and 2500 m of each wetland. Results are reported by wetland
type at: http://www.hort.agri.umn.edu/mnwet/
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4.2 Effects of Enrichment/ Eutrophication/ Reduced Dissolved Oxygen
Processes
Excessive nutrients can affect wetland plant communities in a variety of ways (Wisheu et al.
1990, Weisner 1990), including:
(a) shifting the species composition away from species that take up nutrients slowly, to
those that are able to exploit nutrient pulses more rapidly or which have high nutrient
requirements (Hough et al. 1989, Arts et al. 1990, Gopal and Chamanlal 1991, Wetzel and
van derValk 1998);
(b) triggering algal blooms that can shade out many submersed herbaceous plants (Mason
1990, Crowder and Painter 1991, Stevenson et al. 1993, Srivastava et al. 1995, Short and
Burdick 1995, Coops and Doef 1996);
(c) causing dead plant material to accumulate faster than it can decompose completely,
thus altering understory and soil structure (Neill 1990, Craft & Richardson 1993).
Over the long term, nutrient additions to most wetlands tend to reduce species richness and
increase the dominance of a few species. Often, non-native species are most capable of invading
rapidly changing environments. Consequently they frequently come to dominate some enriched
wetlands.
Woody plants do not usually show an immediate obvious visible response to nutrients. Nutrient-
related shifts in community composition, if they occur at all, do so over long time periods.
However, seed germination, seed production, foliar nutrient concentrations, shoot length, and
annual growth of woody plants are all potentially affected by enrichment and can be measured to
document short-term nutrient exposure. In relatively unaltered watersheds, floodplain trees
farthest from the channel may suffer the least from nutrient deficiencies, due to less frequent
flooding, whereas in watersheds with largely agricultural land cover, near-channel trees may be
less prone to nutrient limitation (Friedman et al. 1996).
Response of wetland plants to nutrient additions can be influenced by many factors, including
type of nutrient, dosing rate, nutrient concentration, soil or water pH, hydrologic conditions,
season, plant species, and life stage. Together, these factors may determine whether a wetland
plant community is limited more by a lack of nitrogen, phosphorus, potassium, calcium, or other
elements (e.g., Craft et al. 1995). Bog vegetation is often limited the most by lack of nitrogen
(Bridgham et al. 1996), whereas vegetation in tidal freshwater wetlands may be limited more by
phosphorus.
Species Richness
Enrichment can increase or decrease species richness of plants within a wetland, depending on
the initial species mix, nutrient loading rates, season, and other factors. In France, wet meadows
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receiving agricultural fertilizer (30 to 160 kg N/ha) had lower species richness, and their dominant
species were common in other wetlands (Grevilliot et al. 1998).
Species Composition
Because the database on EP A's Biological Assessment of Wetlands Working Group (BAWWG)
web site (Adamus and Gonyaw 2000) compiles the literature on nutrient-related species
composition shifts, we have limited the discussion of autecological sensitivities in this document.
Many wetlands are naturally eutrophic. Plants that typify such situations often include species
such as Typha sp., Phragmites sp., Lythrum salicaria, and Lemna (Grace 1989, Davis 1991,
Huebert and Shay 1991, Davis 1994, Otto et al. 1995, Weiher et al. 1996, Doren et al. 1997) which
are relatively invasive and tend to form monotypic stands. Some species occur only in eutrophic
or oligotrophic waters, whereas others span the full range of nutrient conditions (e.g., Srivastava
et al. 1995). Among the species that characterize nutrient-rich wetlands, some dominate not
because they require high nutrient levels, but because they have life history characteristics that
allow them to invade and spread rapidly into enriched habitats where other species are or would
be struggling. Nutrient limitation of Florida Everglades plants was explored by Daoust and
Childers (1999). They found wet prairies were highly P-limited at N:P ratios above 36:1 and
Cladium jamaincense remained dominant, with sub-dominants including Peltandra virginica,
Pontedaria cordata, Saggitaria lancifolia and Panicum hemitomon. When N:P ratios dropped
below this threshold, Typha spp. became increasingly dominant. Hymenocallispalmeri was
shown to be N-limited and may signal a change in nutrient regime.
Bogs and other nutrient-poor wetlands are logically the most sensitive to nutrient additions
(Moore et al. 1989). There, the increased availability of nutrients allows grasses and common
opportunistic plants to out-compete the rare, nutrient-poor specialists such as sundews, orchids,
and pitcher plants. However, plant community response to enrichment of boreal wetlands
depends largely on the species present (Nams et al. 1993).
Nutrient additions to wetlands do not inevitably cause shifts in species composition, at least not
in the short term, in wetlands with just a few dominant (and perhaps "adapted") species. When
nutrients were added to an Everglades fen, areas with moderate (500-750 mg/kg) and low (<500
mg/kg) phosphorus soil concentrations maintained their original plant composition over many
years (Richardson et al. 1999).
Impacts of enrichment are often confounded or even obscured by simultaneous effects of
correlated disturbances. In a southeastern Michigan lake (Hough et al. 1991), a combination of
water level and nutrient declines resulted in a shift from a Potamogeton-dominated community
to one dominated by Nuphar, Nymphaea and Myriophyllum. In the Everglades, shifts in plant
growth and species composition in response to nutrients depend partly on water levels and/or fire
history (Grace 1989, Davis 1989, Davis 1991, Urban et al. 1993, Maceina 1994, Craft et al. 1995,
Newman et al. 1996, Kludze & Delaune 1996, David 1996, Doren et al. 1997).
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Because enrichment can increase densities of phytoplankton and epiphytes, and this reduces the
amount of light available to submersed plants, eutrophic lakes are often dominated by
phytoplankton and non-rooted macrophytes, whereas oligotrophic lakes are dominated by rooted
macrophytes and a higher proportion of submersed plants (Hough et al. 1989, Srivastava et al.
1995). When nutrient additions are curtailed, submersed macrophytes may increase and
consequently stabilize sediments (Scheffer 1991, Stevenson et al. 1993).
Density, Biomass, Growth, Productivity, Germination
Many aquatic plant species respond to nutrient additions by increasing their growth, biomass,
and productivity. Growth responses to enrichment have been documented for only about 80
wetland-associated species in North America, and of these, most have tolerated enrichment or
responded to enrichment with increased biomass or growth (Adamus & Gonyaw 2000). Wetland
macrophytes may be more nutrient-limited than are algae (Duarte 1992). In nutrient limited
systems, growth increases can be dramatic. Fertilization of an Alaskan river resulted in an
increase in bryophyte biomass from 17g dry mass/m2 to 322 g dry mass/m2 (Bowden et al. 1994).
Fertilization of the common bog plant Calluna vulgaris with a garden fertilizer led to increased
flowering per shoot and a greater proportion of flowering shoots (lason and Hester 1993). In
alpine wet meadows of Colorado, fertilization increased the biomass of grasses at the expense of
forbs (Bowman et al. 1993). Overall nutrient uptake, photosynthesis, and growth of the dominant
species were not strongly affected by application of 500-1000 g N and/or P to each of five 2 x 2 m
plots.
The response of plant growth to nutrient addition can vary depending on the degree to which the
species that are present allocate nutrients to roots vs. shoots, as well as their overall life history
strategies (Carter and Grace 1990, Grace 1990, Keddy 1990, Olffetal. 1990). Many plant species
allocate more biomass to shoots during competition for light and more to roots when competing
for nutrients (Tilman 1990, 1991, Poorter and Remkes 1990, Poorter and Lambers 1991). For
example, Phalaris arundinacea and Echinochloa crusgalli have greater root to shoot biomass
ratios under lower nutrient levels (Figiel et al. 1995). Potamogeton nodosus responds to nutrient
additions by allocating biomass predominantly to tubers while Hydrilla verticillata concentrates
nutrients in aboveground structures (McFarland et al. 1992).
Increases in aboveground biomass can smother other plants following senescence of tissue, thus
helping maintain the dominance of species that exploit nutrients the most (van Auken and Bush
1992). When nitrogen and phosphorus fertilizers were added to a Typha glauca and a
Scolochloafestucacea marsh during two growing seasons, this resulted in increased biomass of
both Scolochloafestucacea and Typha glauca. However, biomass of Scolochloafestucacea
declined in the second year due to large accumulated amounts of Typha glauca litter (Neill 1990).
Fertilizer applications of up to 90 kg/ha can increase the aboveground productivity of sedge-
dominated wet meadows, but applications higher than that rate had little added effect (Reece et
al. 1994). Submersed macrophytes in non-eutrophic waters generally increase in response to
moderate nutrient additions, even though most derive their nutrients from sediments rather than
the water column (Spencer 1990, McFarland and Barko 1990, Barko et al. 1991, Spencer et al.
1993, Spencer and Ksander 1995).
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Carbon dioxide is another type of nutrient utilized by wetland plants. During a laboratory
experiment Callitriche was unable to grow under ambient conditions, but CO2 enrichment with
500-1000 uM of carbon dioxide led to growth of 0.089 to 0.124/day. Under the same
concentrations Elodea growth rates increased from 0.046-0.080 to 0.076-0.117/day (Vadstrup and
Madsen 1995).
Enrichment also affects germination rates of some macrophytes. Typha latifolia germinated in
fewer days than did Typha domingensis at high phosphate levels (.200 mg/L). However, the
germination rate of each species was unaffected by other nutrient levels (Stewart et al. 797).
Phosphorus amendments led to improved establishment of Sphagnum and Eriophorum
angustifolium in laboratory experiments (Ferland and Rochefort 1997).
Cottonwood trees treated with 50 kg/ha fertilizer increased stem volume four-fold (van den
Driessche 1999). In Appalachian peatbogs, spatial dominance of bristly dewberry, Rubus
hispidus, was positively related to nutrient levels, but dominance of the Ericaceae shrubs was
negatively related (Stewart andNilsen 1993).
4.3 Effects of Contaminant Toxicity
Much of the literature on aquatic plant sensitivity to chemical contaminants, as revealed by
experimental dosing (phytotoxicity testing), is summarized by Lewis (1995). In general, dose-
response relationships are less well known for vascular plants than for algae, and most
experiments have used floating-leaved plants (especially duckweed, Lemna spp.) rather than
rooted plants.
Processes and Symptoms
Most vascular plants are relatively tolerant of contaminant toxicity. When effects occur, they
usually result from the effects of contaminants on plant metabolic pathways, enzymatic
reactions, and growth (Fitter and Hay 1987). Symptoms of toxicity can include growth
reduction, small leaves, necrotic, chloritic or discolored leaves, early leaf fall, stunted root growth,
suppressed growth of lateral roots and death of root meristems (Pahlsson 1989, Rhoads etal.
1989, Vasquez etal. 1989, Alloway 1990, Kiekens 1990, O'Neill 1990, Kabata-Pendias and
Pendias 1992, Dushenko et al. 1995). Acidic conditions in some wetlands can increase the
harmful effects of many heavy metals (e.g., Carlson and Carlson 1994).
Effects on Species Composition
Shifts in wetland plant species composition in response to contaminants have not been widely
documented. Thus, the ability of plant-based multimetric indices to represent chemical
contamination of wetlands is, at best, currently very limited. Submersed species tend to
accumulate contaminants and also are perhaps the most sensitive plants to contaminants
(Outridge andNoller 1991). For example, in acidic lakes of New Jersey, submersed species
contained higher trace element levels than did floating-leafed species, although one floating-
leaved species (water shield, Brasenid) bioaccumulated zinc and cadmium substantially
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(Sprenger and Mclntosh 1989). In an Ontario lake, cattail stands appeared to tolerate acid mine
drainage and associated heavy metals. Toxicity of acid mine drainage to macrophytes often
depends on local environmental and geological features that alter contaminant bioavailability
(Fyson et al. 1991). Differences also exist among species with regard to their sensitivity to
particular herbicides.
Effects on Biomass, Growth, Health
The invasive submersed plant, Myriophyllum spicatum, did not grow when exposed to cadmium
concentrations above 7.63 |j,g /mL (Sajwan and Ornes 1996). Arsenic, cadmium, copper, lead,
and zinc inhibited growth in hybrid poplar (Populus) and several other tree species (Lejeune et al.
1996). Iron and manganese, although not usually toxic to wetland plants, do affect species in
some wetland types. For example, laboratory experiments revealed differences among 44 fen
species with regard to the influence of iron on growth (Table 4.1) (Snowden and Wheeler 1993).
Boron can also be toxic. When added at rates of 0, 2, 4, 8, and 16 mg/L in laboratory
experiments, it caused significant decreases in the growth of seedlings of several wetland trees:
Betula nigra, Nyssa aquatica, Platanus occidentalis and Taxodium distichum. Severe leaf
damage, but no reduction in growth, occurred at the higher boron levels in Quercus alba.,
Quercus falcatavar.pagodaefolia, Quercus nigra, Quercus michauxii and Quercus phellos
(McLeod and Ciravolo 1998). In a lacustrine wetland exposed to high arsenic levels, cattails
(Typha latifolia) were shorter and had necrosis of leaf tips and reduced micronutrient
concentrations in root tissues. These symptoms were observed at sediment and water
concentrations exceeding 300 |J,m/g and 400 |J,m/g arsenic respectively (Dushenko et al. 1995).
However, another study found that seedlings of Typha latifolia were able to tolerate and
accumulate zinc (1.0 ng/ml), lead (10.0 ng/ml) and cadmium (0.2 ng/ml) (Ye et al. 1998).
Laboratory experiments indicated that Eriocaulon septangulare was unaffected by tissue
cadmium concentrations of less than 2.6 |j,g/g dry weight in shoots and less than 45 |j,g/g in
(Stewart and Malley 1999). The emergent herbs Bacopa monnieri and Scirpus lacustris were
tolerant of cadmium and copper additions of up to 5 (jM/mL, although decreases in chlorophyll
concentration occurred (Gupta et al. 1994).
Oil spills can have long-lasting effects on wetland plant communities (Obot et al. 1992). In a
greenhouse experiment, oil and a detergent used to clean up oil spills were applied to Sagittaria
lancifolia, Scirpus olneyi and Typha latifolia. The leaves on all of the study plants died
following oiling, but new leaves soon developed on those plants subjected to oil and subsequent
cleaning with the detergent. Scirpus olneyi was the least sensitive of the three species whereas
Typha latifolia appeared to be the most sensitive (Pezeshki et al. 1998).
Nitrile and volatile organic acids in culture media were very toxic ioAzollafiliculoides plants. In
secondary effluent containing nitrile, the plants shed their roots, suffered fragmentation of their
fronds, and eventually died (Kitoh etal. 1993).
The herbicides, Rodeo™ and Garlan 3A™, applied to control Lythrum salicaria., also decreased
growth rates of non-target species such as Lemna gibba (Gardner and Grue 1996). The herbicide
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Triclopyr™ has been reported to be relatively harmless to wetland vascular plants (Gabor et al.
1993).
The germination of the forbs Echinochloa crusgalli and Sesbania macrocarpa was unaffected
by unspecified pollutants from a coke plant, pulp mill, and a wastewater treatment plant.
Seedling growth of Echinochloa crusgalli increased after exposure to pollutants from one
wastewater treatment plant but not from another (Walsh et al. 1991).
Bioaccumulation
Arsenic, cadmium, copper, lead, aluminum, and zinc often accumulate in plants growing near
industrial areas (St-Cyr & Campbell 1994) and mining sites, frequently at levels toxic to other
ecosystem components and the plants themselves (O'Niell 1990, Kabata-Pendias and Pendias
1992, Ton et al. 1993). In Montana, upland soils with high levels of arsenic and other metals from
smelter emissions had reduced cover and vertical diversity of plants, lower species richness, and
increased dominance of weedy species (Galbraith et al. 1995). In Colorado, riparian conifers and
Populus tremuloides died when exposed to high levels of iron and manganese, whereas Populus
angustifolia and Salix monticola remained healthy (Barrick and Noble 1993). In Florida,
constructed marshes effectively removed methyl-mercury from Everglades water (Miles and Fink
(1998).
Plants do not inevitably bioaccumulate or biomagnify metals from sediments. Moreover, when
they do accumulate, effects are not always obvious. In marshes of northern Canada, no
correlation was found between spent gunshot and lead in soil and lead in plant tissue, which
remained at background concentrations (Tsuji and Karagatzides 1998). In a southeastern
wetland, the dominant tree, Pinus taeda, accumulated metals and other trace elements, but high
levels of metals in the soil did not prevent growth of seedlings (Carlson and Carlson 1994).
Extensive spatial and species variation in tendency of plants to accumulate contaminants has
been noted in Lake Ontario marshes (Crowder and Painter 1991).
Much of the literature pertaining to plant bioaccumulation of heavy metals was reviewed by
Crowder (1991), Nellessen & Fletcher (1993), and Odum et al. (2000). Plant uptake of metals
appears to be largely influenced by pH, Eh, soil organic content, oxide and carbonate content,
and cation exchange capacity (Crowder 1991), as well as plant species characteristics (Jackson &
Kalff 1993, Thompson et al. 1999). Shallow-rooted marsh species may be more effective sinks
for lead than are woody plants (Ton et al. 1993). Wetland plants can also take up some toxic
hydrocarbons (Gobas et al. 1991).
Table 4.1. Tolerance groupings of fen plant species exposed to various iron concentrations for a
two week period (Snowden and Wheeler 1993).
INSENSITIVE SLIGHTY SENSITVE
Juncus articulatus Valeriana dioica
Eriophorum angustifolium Holcus lanatus
Carex echinata Juncus subnodulosus
Juncus effusus Phalaris arundinacea
Juncus acutiflorus Lysimachia vulgaris
Iris pseudacorus Carex lepidocarpa
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Molinia caerulea Juncus inflexus
Juncus articulatus Potentillapalustris
Oryza saliva Ranunculus flammula
Agrostis stolonifera Briza media
Eriophorum latifolium Carex diandra
Pedicularis palustris Caltha palustris
Parnissia palustris Phragmites australis
Carex pulicaris
MODERATELY SENSITIVE VERY SENSITIVE
Galium palustre Filipendula ulmaria
Carex appropinquata Lychnis flos-cuculi
Lotus uliginosus Rumex acetosa
Trifolium pratense Scrophularia auriculata
Primula farinosa Rumex hydrolapathum
Epilobium palustre Epilobium hirsutum
Thalictrum flavum
Galium aparine
Eupatorium cannabinum
Valeriana ojficinalis
4.4 Effects of Acidification
Processes
Acidic conditions in wetland soils increase the toxicity of aluminum and manganese (Rendig and
Taylor 1989, Crowder 1991). Acidification can directly impact plants by limiting the availability
of some inorganic nutrients and carbon (Farmer 1990). Acidic conditions also promote the
conversion of nitrates into ammonium. In regions with carbonate soils, acidification can mobilize
phosphorus (Reddy et al. 1993). Acidic conditions also can impact plants indirectly by reducing
densities of grazing and detritus-processing invertebrates. Mosses can alter the pH of water,
increasing the bioavailability and toxicity of some metals to other wetland species, and thus
potentially causing shifts in species composition (Vitt & Chee 1990, Vedagiri & Ehrenfeld 1991,
Kooijman andBakker 1994).
Effects on Species Composition and Richness
Effects of acidification (or its reversal by liming) on wetland species composition are not
consistent among wetland types or even within individual wetlands (Farmer 1990, Baker and
Christensen 1991, Mackun et al. 1994, Weiher et al. 1994). Many plant species that inhabit bogs
and pocosin wetlands are, of course, adapted to tolerating acidity levels that would kill most
wetland plant species. Species whose decline or disappearance from a lake coincided with
acidification include Lobelia dortmanna, Isoetes riparia, Myriophyllum tenellum., Nuphar sp.,
Utricularia vulgaris, and Potamogeton epihydris (Farmer 1990). Species whose relative
abundance increased include Leptodictium riparium, Eleocharis acicularis, Sphagnum sp., and
Eriocaulon septangular e (Farmer 1990). Within a two year period after 1100 mg of lime was
applied to a 100 hectare Adirondack watershed, only 6 of 64 wetland plant taxa increased in
cover, frequency, or importance. The cover of Sphagnum and Carex interior decreased in
control plots compared to limed plots. Cover of Cladium mariscoides tripled in extent. In
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contrast, Drosera intermedia., Hypericum canadense, andMuhlenbergia uniflora all were
negatively impacted by the liming.
Among 51 Maine peatlands, plant species richness tended to increase with decreasing acidity and
rising pH (Anderson & Davis 1997), as was also the case in Minnesota peatlands (Glaser et al.
1990) and northern Ontario wetlands (Jeglum et al. 1995). Species associated with very or
moderately acidic conditions included Smilacina trifolia, Carex oligosperma, Chamaecyparis
thyoides, Pinus strobus, Utricularia cornuta, Vaccinium angustifolium, V. oxycoccus,
Gaylussacia dumosa, and Kalmia angustifolia. Species associated with less acidic conditions in
the Maine peatlands included Alnus rugosa, Campylium stellatum, Trichophorum alpinum,
Rhamnus alnifolia, Betulapumila, Thuja occidentalis, Abies balsamea, Aster borealis,
Muhlenbergia glomerata, and Onoclea sensibilis.
Effects on Density, Biomass, Growth, Germination
Biomass of the submersed aquatic plant, Vallisneria americana, when transplanted to acidic
lakes quickly fell to 0.008 g dry mass, whereas plants transplanted to alkaline lakes grew
vigorously to 7.5 g dry weight (Overath et al. 1991). In laboratory experiments, tuber growth was
decreased by 97% for Vallisneria americana at pH 5 compared to normal growth at pH 7.5. At
the same low pH, Najas flexilis produced no flowers and few tubers (Titus and Hoover 1993). In
another round of experiments, low pH (pH 5) reduced the growth of Vallisneria americana, but
0.15 mM KHCO3 stimulated growth by 2.8 to 10 fold. CO2 availbility appeared to be an important
control on the growth of this species (Titus et al. 1990). Growth of Carex exilis seedlings was
mostly unaffected by varied acidity (Santelmann 1991). In a laboratory experiment, the
germination of Typha latifolia seeds exposed to cattail ash, leaf extracts, and a variety of pH
levels was unaffected by any of the treatments (Rivard and Woodard 1989).
4.5 Effects of Salinization
Processes
High concentrations of soluble salts in soil water are lethal to plants, and sub-lethal levels may
impair growth (Rendig and Taylor 1989). Woody plants tend to be less tolerant than herbaceous
plants because they do not have mechanisms for removing salt, other than accumulating salts in
leaves and subsequently dropping them.
Effects on Species Composition
Many plant species that inhabit inland saline wetland and coastal tidal wetlands are, of course,
adapted to tolerating salt levels that would kill most wetland plant species. A survey of inland
lakes in western Canada which spanned a salinity gradient identified relative tolerance to salinity,
and specific salinity tolerance thresholds, of many wetland species (Hammer andHeseltine 1988).
Short-term salinity pulses (or fresh water pulses) that occur during storm events can affect spatial
patterns and species composition of plant communities in tidal marshes (Howard and
Mendelssohn 1999b).
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Effects on Growth, Germination
The susceptability of plant species is highly variable with some suffering signs of stress almost
immediately whereas others can tolerate elevated salinity for 6-8 weeks (Howard and
Mendelssohn 1999a). Salinity concentrations of as little as 3 ppt result in a substantial stress in
several southeastern wetland trees (e.g., Nyssa aquatica, Taxodium distichum) (see review in
Pezeshki et al. 1990). Biomass of Taxodium distichum var. distichum seedlings declined after
flooding with water of 6-8 ppt salinity (Allen et al. 1997) but seedlings survived 10 ppt salinity for
8 weeks in another experiment (Conner et al. 1997). Cephalanthus occidentalis and Nyssa
sylvatica var. biflora are both sensitive to salinity in excess of 2 ppt. Responses of these species
to salinity are most evident in gross photosynthesis, stomatal conductance, water pressure
potential, and stem and root biomass (McCarron et al. 1998).
Among emergent plants, the growth of Typha domingensis was reduced to zero when salinity
exceeded 6 ppt, and 15 ppt salinity caused 75% mortality (Glenn et al. 1995). In contrast, Carex
exilis seedlings were virtually unaffected by minor differences in salinity (Santelmann 1991).
Some species of plants are stimulated by low salinity levels. Chenopodium rubrum growth was
stimulated by low concentrations of sodium sulfate and sodium chloride although increased
concentrations reduced dry mass and leaf area (Warne et al. 1989). Panicum hemitomon and
Leersia oryzoides tolerated 9.4% salinity for up to a month in laboratory conditions whereas
Sagittaria lancifolia showed damage at only 4.8% salinity (McKee and Mendelssohn 1989).
Effects of salinity may interact with inundation to influence plant mortality. At low salinities,
Leersia oryzoides growth appears to be inhibited by flooding, whereas Panicum hemitomon
growth was not significantly affected by flooding (McKee and Mendelssohn 1989). Growth may
also be reduced due to the combination of salinity with reductions in soil oxygen associated with
flooding, as with seawater flooding of Taxodium distichum and Nyssa aquatica (Pezeshki 1990,
Pezeshki et al. 1995). In southern forested wetlands, Taxodium distichum var. distichum
tolerated flooding with low salinity water (salinity 2 g/L) whereas biomass decreased after
flooding with water of 6-8 g/L salinity (Allen et al. 1994). Taxodium distichum collected from
coastal zones of the southern United States were more tolerant to salinity stress than those
collected from inland zones (Allen et al. 1997).
Salinity influxes may harm the ability of seeds to germinate, even after the salinity stress is
relieved (Khan and Ungar 1999). When seeds of Triglochin maritima from an inland salt marsh
were exposed to various salinity levels, a level of 400 mol/m3 NaCl resulted in no germination.
Even seeds that were then transferred to distilled water suffered some degree of mortality.
4.6 Effects of Sedimentation, Burial
Processes
Sedimentation is a naturally occurring process in wetland systems, but accelerated rates of
sediment deposition (or erosion) can tax the ability of wetland plant communities to adapt
(Kantrud et al. 1989, Jurike^a/. 1994, Wang et al. 1994). Sedimentation can affect wetland
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systems by the addition of sediment-born pollutants, the burial of established vegetation, and the
burial of seed banks (Neely and Baker 1989, Childers and Gosselink 1990, Pucket etal. 1993).
Burial of leaves has the direct effect of removing light needed for photosynthesis, and restricting
foliar gas exchange (Ewing 1996). Buried plants expend energy elongating their shoots in an
attempt to outpace sedimentation, seeking oxygen and light, and consequently may be less
robust (lower biomass). Over the long term, sedimentation can shrink the proportion of shallow
wetlands that remains suitable for wetland plants, or increase the suitable habitat area in ponds
that previously were too deep to support many wetland plants. Such long-term changes in water
depth (relative elevation) due to sedimentation also result in shifts in species composition, as has
been documented in the Mississippi River floodplain.
Moderate rates of sedimentation are also required by several species native to floodplain
wetlands. For example, the endangered Boltonia decurrens, a perennial member of the aster
family, appears to require bare and sandy alluvial deposits for germination and survival in Illinois
(Smith et al. 1993, Stoeker et al. 1995). In the riparian zones of Alberta, cottonwood seedlings
display a high degree of recruitment following a record flood in early June of 1995 (Rood et al.
1998). The flooding and the accompanying sediments that were deposited with seeds induced
germination ofPopulus angustifolia, Populus balsamifera, and Populus deltoides. The presence
of naturally-deposited islands and bars in large rivers was considered by Dykaar and Wigington
(2000) to be a useful indicator of river-floodplain integrity because of the role of these features in
sustaining stands of cottonwoods and several other riparian trees.
Effects of sedimentation on particular wetland plant species are not well documented (van der
Valk and Jolly 1992, Bartel and Maristany 1989). Many mature plants, and especially woody
species, apparently are not harmed by a small amount of sedimentation (Wang et al. 1994).
Adult plants of Vallisneria americana tolerated burial to depths of up to 10 cm but none
survived burial under sediment depths of 25 cm (Rybicki and Carter 1986). Growth of the
invasive reed, Phragmites australis, typically keeps pace with moderate rates of sedimentation
(Pyke & Havens 1999). However, seeds, seedlings, and plants that have evolved in wetland types
in which sedimentation is rare (e.g., bogs) are highly sensitive to burial. The size of particles that
are being deposited, not just their amount, also may influence plant survival (Dittmar and Neely
1999).
Effects on Species Richness, Species Composition
Significant declines in seedling species richness were observed in wetland plots receiving as little
as 0.25 cm of sediments (Jurik et al. 1994). Deposition (in floodplains) of sediments to a depth
approaching 1 m can prevent shallow rooted species from becoming established. Such
deposition also can result in a shift to species capable of being sustained only by local
precipitation, as well as species unaccustomed to the severe natural disturbance regimes of
floodplains (Shafroth et al. 1995).
Sedimentation can result in significant community change as the germination and growth of the
most sensitive species are suppressed. Species with larger seeds appear to be better able to
survive burial with excessive amounts of sediment (Dittmar and Neely 1999, Jurik et al. 1994,
Wang et al. 1994). Accelerated sedimentation of backwater wetlands was at least partly
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responsible for allowing a non-native plant, Colocasia esculenta, to proliferate (White 1993). Of
14 taxa experimentally buried with sediments of various sizes by Dittmar and Neely (1999), only
Bidens coronata, Polygonum amphibium, Ludwigiapalustris, and a Carex species were
negatively affected. The seemingly unaffected (or positively affected) species included
Epilobium coloratum, Eupatorium perfoliatum, Galium tinctorium, Impatiens capensis,
Lycopus uniflorus, Polygonum pennsylvanicum, Polygonum persicaria, Polygonum punctatum,
Polygonum sagittatum and Verbena hastata. In a survey of several Pennsylvania wetlands
where sedimentation rates were also measured, only 6 of the 35 species were considered
intolerant of sedimentation (Wardrop and Brooks 1998) (Table 4.2). Although the sedges Carex
rostrata and Carex stipata were mostly resilient to cycles of flooding and drying, sediment
deposits resulted in decreased biomass, which was diminished further by high water levels
(Ewing 1996). Among woody plants, saplings ofAlnus rubra tolerated burial less well than those
of Fraxinus latifolia (Ewing 1996). In Florida wetlands receiving sediment-laden stormwater,
several invasive species (Typha latifolia., Ludwigiaperuviana andMikania scandens) were most
dominant nearest the input pipe (Carr 1994).
Table 4.2. Sediment tolerance categories assigned to Pennsylvania wetland plants experiencing
sedimentation rates of 0 to 8 cm/year (Wardrop and Brooks 1998).
Intolerant Moderately Tolerant
Asclepias syriaca Brachyelytrum erectum
Aster vimineus Carex emoryi
Cirsium arvense Carex folliculata
Lysimachia nummularia Carex retroflexa
Mentha arvensis Carex prasina
Poapratensis Carex stricta
Carex vulpinoidea
Phalaris arundinacea
Solidago sp.
Symplocarpus foetidus
Thelyoteris noveboracensis
Triadenum virginicum
Slightly Tolerant Very Tolerant
Juncus canadensis Aster novae-angliae
Euthamia graminfolia Dipsacus sylvestris
Sagittaria latifolia Dulichium arundinaceum
Eleocharis sp. Impatiens capensis
Verbena hastata Leersia oryzoides
Equisetum arvense Polygonum sagittatum
Carex intumescens Solidago patula
Solidago Canadensis Solidago uliginosa
Ulrica dioca
Effects on Density, Biomass, Germination
Excessive sedimentation can reduce seedling recruitment (Jurik et al. 1994). For example, Typha
seedling density and biomass decreased as sediment loads increased from 0.2 to 1.0 cm. One
study found a fourfold greater density of annuals (vs. perennials) in some heavily sedimented
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sites (Neely and Wiler 1993). Older and larger seedlings tolerated burial better (Wang et al.
1994).
Sediment additions have been found to reduce germination rates of wetland herb species by 34%
(Neely and Wiler 1993), 80% (Jurik et al. 1994), and 90% (Wang et al. 1994). Sedimentation can
significantly alter the species composition of wetland plant communities, as seeds of the most
sensitive species fail to germinate (Dittmar and Neely 1999). Less than 1 cm of sediment can
inhibit germination of Typha sp., Echinocola crusgalli., Leersia oryzoides, and Carex sp. (Jurik
et al. 1994). In contrast, burial by 2 cm of sediment does not interfere with germination of
several non-native plant species (Blackshaw 1992, Reddy and Singh 1992). Sedimentation
inhibits the germination of Typha latifolia seeds more than Sparganium eurycarum seeds (Neely
et al. 1993). Typha germination (as well as species abundance and total number of individuals)
was inhibited by burial under as little as 0.25 cm layer of sediment; this species germinates best
under low oxygen conditions but with light present (Jurik et al. 1994). Germination of cattail
(Typha xglauca) seeds decreased by 60-90% when sediment loads of 0.2 to 0.4 cm were applied
to the surface of the soil (Wang et al. 1994).
Germination of emergent herbaceous species is typically promoted by cold stratification, seeds
positioned in the light at the surface of wet but not flooded soils, and fluctuating temperatures
(Shipley and Parent 1991). Deposits of sediments on seeds lying on exposed soils can alter these
conditions and reduce seed germination. Seeds submersed underwater are also vulnerable to
sedimentation, but might be slightly more tolerant (Neely et al. 1993, Clevering 1998).
Germination of a submersed plant, Myriophyllum spicatum, was reduced by sediment depths of
2 cm or more (Hartleb et al. 1993).
4.7 Effects of Thermal Alteration
Temperature affects wetland plants mainly by influencing seed germination, and by extending or
shortening the growing season. Growth (biomass accumulation) of some species also may be
affected (e.g., Phalaris arundinacea, Bernard & Lauvel995), giving particular species a
competitive advantage or disadvantage (Landhausser and Lieffers 1994). Gradually rising
springtime temperature triggers the germination of many wetland herbs (Leek 1989, Hogenbirk
and Wein 1992). Abnormally increased temperatures during the dormant season can potentially
hinder germination of some seeds that require cold temperatures to alter hormones related with
germination. Heated effluents from power plants (Crowder and Painter 1991, Taylor and Helwig
1995), wastewater treatment facilities, landfills (Bernard & Lauve 1995), and other sources can
support localized populations of species that otherwise occur farther south (e.g., but may
interfere with germination of species . Global climate warming also is anticipated to cause
geographic shifts in wetland plant communities.
Germination of wetland plants is affected not only by temperature magnitude, but also by
temperature fluctuation (amplitude), at least during particularly crucial times of the year. For
example, in experiments involving 45 combinations of diurnal mean temperature and amplitude,
Ekstam and Foresby (1999) found Phragmites australis to require a high amplitude (>10 °C) for
germination over the entire range of mean temperatures, whereas germination of this species was
less sensitive to mean temperature than was germination of Typha latifolia.
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Temperature can also influence species composition by influencing the amount of oxygen
persisting in soils or sediments (Callaway & King 1996, Crawford & Braendle 1996). Higher
levels of oxygen are typically associated with cooler temperatures, and create an environment
that may be more favorable to facultative and upland species than to wetland obligates.
4.8 Effects of Vegetation Removal
Processes
By definition, removal of any vegetation from a plant community causes at least a short-term
change in plant biomass and possibly species composition. Removal can occur as a result of fire,
tillage, mowing, herbivory (including grazing by ungulates and aquatic animals — see reviews by
Newman 1991, Naiman & Rodgers 1997), mortality from contaminants (e.g., herbicides), logging
or beaver activity, dredging or construction activities, or damage from wind (Loope et al. 1994),
ice, or flooding.
Vegetation in wetlands typically is adapted for the usual kinds and intensities of disturbances that
have occurred for centuries prior to the advent of human influence. That is, many wetlands
affected by natural disturbances eventually recover to a condition somewhat resembling their
prior state, provided surrounding landscapes have not been drastically changed by humans. For
example, fires are historically a natural phenomenon in Carolina bays and pocosins, and no
significant change was found in species richness, evenness, or diversity of plant communities
following one fire (Kirkman and Sharitz 1994). In other situations, particularly when wetland
plants are already exposed to drought, floods, or other severe stress, fire can induce large shifts in
species composition and sometimes biomass (Tilman and El Haddi 1992). For example,
although fire alone had little effect on the emergence of a fire-tolerant grass (P. hemitomon),
winter fire followed by spring inundation significantly decreased emergence (Kirkman and
Sharitz 1993). Vegetation removal can also facilitate the introduction or expansion of weedy
species (see Species Composition, below). Fire can mobilize some contaminants from soils (e.g.,
boron, Busch and Smith 1993). Removal of vegetation - especially woody vegetation - by fire,
cutting, or severe floods can result in a rise in local water tables, with consequent implications for
plant biomass and species composition (Hodgkinson 1992).
The type of removal process appears to influence the type, duration, magnitude of the effect on
plants. When removal is total or nearly total (such as with herbicides), recovery occurs mainly
via seedling recruitment. When removal is by non-lethal modes (such as herbivory), recovery
often is by vegetative growth. These differential effects of lethal and nonlethal disturbances are
partly responsible for vegetation patterns in some wetlands (Baldwin and Mendelssohn 1998).
Effects on Species Composition and Richness
Fire is perhaps the most-studied vegetation removal factor affecting species composition in North
American wetlands. Fire suppression can lead to the establishment of longer-lived shrubs at the
expense of herbaceous vegetation (Shedlock et al. 1993). In Atlantic white cedar wetlands in
coastal Massachusetts, the dominant Chamaecyparis thyoides became much more common
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after years of fire suppression, but in the last century light timber thinnings have favored its
partial replacement by red maple (Acer rubrum). Recently, neither Chamaecyparis thyoides nor
Acer rubrum have regenerated well where an undisturbed cedar overstory persists (Motzkin et al.
1993). Cattail stands in the Everglades often expand following fires (Urban et al. 1993). The time
interval between fires also can shape tallgrass prairie marsh communities (Collins and Gibson
1990, Johnson and Knapp 1995). Annual fires tend to reduce diversity to a greater degree than
less frequent fires, due their causing a reduction in forb diversity (Johnson and Knapp 1995).
The effects of fire on species composition and biomass depend significantly on the season when
burning occurred, as was found in the Delta Marsh wetlands of Manitoba (Thompson & Shay
1989) and in southeast Missouri wetlands. There, burning of wetlands in spring improved habitat
and food conditions for waterfowl, whereas burning in summer benefitted shorebirds and helped
keep invasive plant species in check (Laubhan 1995).
In some Louisiana wetlands, herbivory and fire individually affect the structure and composition
plant communities, but their effects are not necessarily interactive. No significant differences
were found in species richness between herbivory treatments or between fire treatments (Taylor
et al. 1994). Another Louisiana study found Spartinapatens to be less resilient to herbivory
than Scirpus americanus (Broome et al. 1995).
The highly invasive, often non-native species that typically colonize sites where vegetation has
been removed non-selectively can have profound effects on species composition of wetland
communities (Swetnam 1990, Busch and Smith 1993, Crins 1989). Native sedges and grasses
responded positively following removal of purple loosestrife from wetlands using herbicides; this
was attributed partly to increased sun exposure (Gabor et al. 1995). In the Florida Everglades,
light transmittance by Typha spp. was measured as 15% compared with 65% by Cladium
jamaicense. Consequently, reduced light available for periphytic photosynthesis was predicted
to influence the replacement of sawgrass by cattail (Grimshaw et al. 1997).
After 25 years of regeneration following mining of peat and associated vegetation, four mined
areas of an Ontario bog had attained 50% similarity with the species composition of a relatively
unaltered bog area (Jonsson-Ninniss & Middleton 1991).
Effects on Density, Biomass, Growth
Burning of a Louisiana lacustrine marsh reduced the aboveground biomass of Spartina patens
and Bacopa monnieri. After recovery, none of the species present before the burn increased in
biomass as a result of the burn. Biomass was over 1.5 times greater in the plots that remained
unburned than in those that were burned. Biomass was almost twice as great in plots protected
from herbivory than in plots subject to natural herbivory after the burn (Taylor et al. 1994). Fires
can increase cone production and seedling survival of baldcypress (Taxodium distichum) by
removing competition (Conner & Toliver 1990, Cook andEwel 1992, Conner 1993). Several
other studies have examined the regeneration of southern bottomland and cypress swamps after
thinning (Ewel and Davis 1992), clearcutting (Kennedy and Meadows 1993), and hurricane
damage (Putz and Sharitz 1991).
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Effects of grazing depend partly on density of grazers, duration of presence in the grazed area,
availability of food and water in nearby alternative habitats, and season (Popolizio et al. 1994,
Clary 1995, Fitch & Adams 1998). Excessive herbivory from deer populations may have caused
Chaemacyperis thyoides to be replaced by Acer rubrum in swamps of the New Jersey Pine
Barrens (Stoltzfus 1990). Herbivory by nutria (Myocastor coypus) can limit regeneration of
baldcypress stands (Brantley and Platt 1992, Myers et al. 1995). Short-term grazing of riparian
vegetation after more than 30 years of cattle exclusion stimulated growth of herbaceous
vegetation (Popolizio et al. 1994).
Nonetheless, biomass and production of the herbaceous community as a whole can increase
following temporary introduction of grazers at low densities in some riparian and wetland
communities (Heitschmidt 1990, Matches 1992, Clary 1995). Also, response of riparian
communities to partial vegetation removal depends on the type of wetland or riparian community
involved (Clary 1995). For instance, a riparian site dominated by the grass, Agrostis stolonifera,
which was subjected to mowing (when 10, 5, and 1 cm high) in spring, fall, or both seasons,
increased or maintained aboveground biomass as measured the following year. Also, plant
biomass in mowed sites dominated by Carex stayed the same or decreased, following spring,
mid-summer, and late summer cuttings. In one of the three Carex dominated sites, forbs
increased in response to cuttings at 1 cm, probably due to reduced competition (Clary 1995).
On lakeshores and river margins, ice commonly damages or partly removes wetland vegetation,
especially during low water years or intentional winter drawdown (Crowder and Painter 1991,
Begin and Payette 1991, Belanger andBedard 1994, Scott et al. 1997). Emergent species seldom
tolerate removal of their tops to the extent that their tops no longer protrude above the water
surface, whereas some floating-leaved species do (Middleton 1990).
When wetland plants are removed by dredging or excavation, full recovery of plant biomass in
the disturbed areas may require more than 10 months, with vegetatively-reproducing species
sometimes becoming more dominant than annuals in the disturbed areas (McKnight 1992).
Similarly, perennial species gained dominance over annuals, mosses, and ferns in Montana
wetlands that were cultivated (Borth 1998). Also, in a subset of these wetlands that were grazed,
species found prior to grazing only in shallow areas and along the upland edge tended to be
found more often in deeper waters once grazing was initiated.
4.9 Effects of Turbidity/Shade
Turbidity (i.e., decreased water clarity) almost by definition means decreased availability of light
to submersed aquatic vegetation, killing many species. Conversely, increased water clarity may
result in increased cover of submersed aquatics (Scheffer et al. 1993), which in turn can improve
water clarity even further by reducing resuspension of sediments (James and Barko 1990).
Nutrient additions often increase phytoplankton growth and consequently turbidity, whereas
control of nutrient sources may favor vascular plants and increase light penetration of the water
column (Dushenko 1990, Hanson and Butler 1990). Turbidity in wetlands also can result from
resuspension of bottom sediments or erosion within a wetland, sediment-laden runoff or channel
water sources, or windborne inputs. Many introduced bottom-feeding fish, e.g., carp, stir up
sediments and consequently have caused changes in submersed aquatic vegetation cover and
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species composition (Richardson et al. 1995). Severe turbidity typically shifts plant community
structure towards floating and emergent species and away from submersed species. Differences
in turbidity tolerance exist among submersed species, e.g., Potamogetonpectinatus appears to
be relatively tolerant of murky waters (Kantrud 1990, Nichols and Lathrop 1994). Propagules of
Egeria densa growing at 1.8 m depth grew well at suspended solid levels up to 25 g/m3 in spring
and autumn, and 35 g/m3 in summer (Tanner et al. 1993).
Canopy cover can have much the same effect as turbidity, reducing the area of underwater
vegetation and some terrestrial plants as well (Pukkala etal. 1991, 1993, Sims andPearcy 1993,
Small etal. 1996).
4.10 Effects of Dehydration or Inundation
Processes
Topographic variation on the order of a few centimeters can shape the composition and richness
of the plant community by influencing the duration (Dicke & Toliver 1990, Merendino & Smith
1991, David 1996, Vivian-Smith 1997, Silverton et al. 1999), timing (Merendino etal. 1990,
Squires and van der Valk 1992, Scott et al. 1996, 1997, Gladwin and Roelle 1998), and frequency
of saturation (van der Valk 1994, Pezeshki etal. 1996, 1997, Smith 1996, Pollock etal. 1998) in
the root zones of wetland plants. The amplitude and rate of water level fluctuation also
influences plant species composition, biomass, and germination (Hull etal. 1989, Hudon 1997,
Shay et al. 1999). This is true even when the substrate beneath the plants is not dewatered.
Effects of fluctuations on particular species are influenced partly by oxygen status of the
sediments, with anaerobic sediments benefiting the early growth of some species (Spencer and
Ksander 1997). Minnesota lakes whose water levels were not artificially manipulated supported
much more diverse communities of submersed plants (Wilcox & Meeker 1991). Among 26
Seattle-area wetlands, the degree of seasonal water level fluctuation had no statistically significant
effect on species richness in the forested wetlands, but was negatively associated with richness
found in emergent and shrub wetlands. Fluctuation during the early spring seemed to have an
especially detrimental effect on plant richness in the emergent and shrub wetlands (Cooke &
Azous 2000). A lack of water level fluctuation can be just as damaging as excessive fluctuation
to some wetland species (Rood and Mahoney 1990). This is because many species need a period
of desiccation in order to germinate.
In riverine wetlands, inundation is often accompanied by severe scouring of the substrate by
major floods. Such scouring reduces the biomass of many species (at least temporarily) but also
allows increased germination of understory or underwater species by reducing dense stands of
plants (especially herbs) that otherwise would crowd or shade out seedlings (Streng 1989,
Friedman et al. 1995, Spink & Rogers 1996, Osterkamp 1998). The influence of periodic floods
on floodplain vegetation is so profound that structural patterns of floodplain vegetation
sometimes can be used to indicate the nature of prior flood events (e.g., Hupp 1992). In a river in
western Colorado, the duration of flooding was found to be more important than the magnitude
of annual peak flow in driving lateral channel migration, which in turn was a major influence on
patterns of floodplain vegetation (Richter & Richter 2000).
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Wetland plants have many adaptations for coping with prolonged flooding or drought (Rubio et
al. 1995). Nonetheless, inundation and/or saturated soil conditions potentially kill plants when
sediment oxygen deficits alter plant metabolic processes or allow buildup of substances toxic to
plants. Thus, soil texture, which influences soil oxygen levels, can influence the sensitivity of
some species to inundation (Wallace et al. 1996).
Inundation also may increase or decrease the exposure of plants to competitors and herbivores
(Wilson & Keddy 1991) and cause a shift in the location of plant communities within a wetland
(van der Valk et al. 1994). The opposite extreme — dehydration - kills plants partly by removing
the pathway for taking up nutrients and maintaining tissues, and may also increase or decrease
competition and plant exposure to herbivory. Interruption of water corridors between wetlands
(e.g., by dams or water diversions) can hinder the spread of seeds of some wetland plants, e.g.,
Carex sp. (Budelsky and Galatowitsch 1999), whereas periodic floods can assist recolonization of
denuded areas (Spink & Rogers 1996).
Differences exist among plant species with regard their ability to resist drought and flooding.
These differences are related to plant life history and physical characteristics (Earnst 1990,
Koncalova 1990, Voesenek etal. 1993, Kirkman and Sharitz 1993, Teutsch and Sulc 1997).
Specifically, the seed dispersal and germination characteristics of plants may have the greatest
effect on the relative abundance of species, according to a model simulation exercise conducted
by Ellison & Bedford (1995) using 6 years of data from a southern Wisconsin sedge meadow.
The size of seeds and differences in the timing and methods of seed dispersal can help explain the
occurrence of some species (Grillas et al. 1991). In bottomland hardwood forests, smaller seeded
species have more seeds dispersed, more germinants, more established seedlings, but not
necessarily more surviving seedlings at the end of one year (Jones et al. 1994). Also, some
species, such as cat-tail, are able to keep pace with rising water levels because they are able to
rapidly elongate their stem tissue to a greater degree than other species (Waters and Shay 1992,
Galatowitsch et al. 1999) or sprout adventitious roots (Voesenek et al. 1989)
Feedback mechanisms exist, inasmuch as plants themselves, via transpiration, can potentially
dehydrate wetlands and increase slightly the magnitude of water level fluctuations (Dube et al.
1995). Plants can have the opposite effect as well, sheltering surface water and exposed
sediments from the evaporative influences of solar radiation and wind, thus maintaining moisture
that benefits their growth.
Effects on Species Richness
Inundation and dehydration have variable effects on plant community richness. In a California
subalpine wetland, species diversity within Carex rostrata, Scirpus acutus, and Nuphar
polysepalum communities was highest during dry years whereas biomass was lowest then
(Rejmankova et al. 1999). In Alaskan riparian wetlands, sites with intermediate flooding were the
most species-rich, whereas those with no flooding or high or low flood frequency were species-
poor (Pollock et al. (1998). Species richness tended to peak (at 100-110 species) when mean
flooding frequency reached 11-12 events per year. In the riparian zones of six mid-sized streams
in Vermont, an average of 90% of the total species present out to 50 m from the high water mark
were found within the first 30 m of the high water mark (Spackman & Hughes 1995).
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Water table levels often decline when local or regional groundwater resources are depleted, and
plant species composition may change as a consequence, e.g., Sonenshein & Hofstetter (1990),
Segelquist et al. (1993), Rochow & Rhinesmith (1992), Stromberg et al. (1996). Water table
levels also change when river or lake levels are regulated, e.g., Nilsson et al. (1991), Hughes
(1990), Gregory et al. (1991), Naiman and Decamps (1997), Jansson et al. (2000). Many plant
species inhabit only wetlands that are inundated briefly, and so can provide a major contribution
to regional plant richness in regions where many such wetlands have been destroyed by water
table declines or converted to other uses (Hoagland & Collins 1997).
In Texas, April drawdown of wetland water levels produced the greatest species richness
(McKnight 1992). In a study of 26 Seattle-area wetlands, wetlands whose contributing
watersheds became developed during the multiyear study experienced a decline in plant species
richness, whereas urban and rural reference wetlands changed little between years (Cooke &
Azous 2000).
Seed bank richness decreases with increasing water depth (Wilson et al. 1993, Haukos & Smith
1994) and permanency of flooding (Baskin et al. 1996). Old and young beaver ponds in Quebec
were found to have similarly rich and abundant seed banks (Le Page & Keddy 1998).
Effects on Species Composition
Because the database on EPA's BAWWG web site (Adamus and Gonyaw 2000) compiles the
literature on hydrology-related shifts in plant species composition, we have limited the discussion
of autecological sensitivities in this document.
Species that can move vertically with floodwaters (e.g., Utricularia vulgaris, which is not rooted
to the substrate), or which grow quickly enough to keep their leaves above water, are better able
to flourish with increasing water levels (Murkin et al. 1991). Woody plants are particularly
sensitive to prolonged inundation (especially >80 days) (Niswander & Mitsch 1995, Toner &
Keddy 1997, Sharitz & Gresham 1997). Their seedlings consequently are most affected during
years when flooding occurs at or shortly after the beginning of the growing season, or when
flooding persists for >40% of the growing season (Toner & Keddy 1997). Annual (as opposed to
perennial) species tend to increase proportionately in response to drought and some other severe
disturbances (Poiani and Johnson 1989), and species richness tends to be lower where lake and
river communities are dominated by annuals (Shipley et al. 1991). A relatively high proportion of
plant species that are characteristically wetland "facultatives" (as opposed to obligates), including
most woody plants, also suggests relatively dry conditions during at least part of the growing
season in a wetland. Species with small, light seeds seem particularly adept at colonizing
mudflats exposed during drawdowns and after disturbances (Poiani and Johnson 1989, Ellison
and Bedford 1995), and tend to emerge early in the season, thus usually increasing their
probability of success due to greater light availability (Toner & Keddy 1997). Successive years of
annual drawdowns can favor the spread of many non-native plant species within wetlands (van
der Valk 1994). Dominance of a wetland by just a few species is sometimes a sign that the
wetland has experienced prolonged drought or drawdown (Wilcox 1995).
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Restoration of natural hydrologic regimes to regulated rivers can have dramatic effects on plant
species composition (Busch & Smith 1995, Poff et al. 1997). Following a large flood on the
Hassayampa River in Arizona, several native species that previously had been uncommon
increased (Stromberg et al. 1997). With careful planning, restoration of historical flows to
regulated rivers, and historic water table levels to isolated wetlands, can allow characteristic native
species to resume dominance (Stromberg et al. 1991, Toth 1993, Toth et al. 1995, Briggs &
Cornelius 1998, Sher et al. 2000). In the Portland, Oregon, metropolitan area, policies that
allowed out-of-kind mitigation of wetland losses resulted in plant communities in constructed
wetlands that differed significantly from those present in naturally-occurring wetlands (Magee et
al. 1999). Engineering specifications for the constructed wetlands favored plant species more
typical of deeper, more permanent water.
Effects on Density, Biomass, Production, Germination, Growth
Inundation or dehydration can either increase or decrease the germination, survival, biomass,
biomass allocation (roots vs. foliage), growth, and density (basal area or shoots per unit area) of
wetland plants. The effect is very species-specific, and depends on the age (life stage) of the
plant, concentration of oxygen in sediments, direction of water level change (rising or falling),
duration, depth, and season (temperature/light) of flooding (Busch et al. 1998). Many species
have only a narrow "window" in which they can germinate, for example, a few-week period
when favorable water levels (or temporary lack of competitors) must coincide with favorable
temperatures and acceptable water quality (Rood et al. 1998, Roelle & Gladwin 1999).
Woody plant production is typically higher in flowing-water wetlands with natural flood pulses
than in wetlands where water is continuously stagnant (Mitsch et al. 1991). Cumulatively, entire
communities can be structured by gross factors such as annual maximum and minimum flows
(Auble et al. 1994) or very specific hydrologic sequences, e.g., the last date of the first flood of
the season and the first date of the second flood (Toner & Keddy 1997), or prolonged flooding
followed by drawdown followed immediately by moderate inundation (in the case of
baldcypress). Water velocity, channel pattern, and substrate characteristics also are important to
some species (Smith and Wellington 1991). For example, the dominance of cottonwoods over
saltcedar can be determined partly by the distribution of flow across the floodplain (Szaro 1990,
Howe and Knopf 1991, Cuomo 1992).
Responses to specific hydrologic variables of hundreds of individual species that have been
studied are presented in the database at EPA's web site (Adamus & Gonyaw 2000). Evidence
from some studies suggests relative tolerance of water level fluctuations is greatest among several
non-native or invasive species (Figiel et al. 1991, Haworth-Brockman and Murkin 1993, King and
Grace 2000). Among tree species, seed weight may be an indicator of drought tolerance, with
heavier-seeded species being more tolerant (Streng et al. 1989). However, among herbaceous
species, seed weight may not correlate with germination characteristics such as lag time,
maximum germination rate, and final germination proportion (Shipley & Parent 1991c).
4.11 Effects of Other Human Influences
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Several studies have documented changes in wetland vegetation as a result of general watershed
development, without diagnosing in all cases the exact stressor(s) responsible. For example,
plant species richness in forested Ontario wetlands was found to be negatively correlated with
density of roads within 2 km, but the exact mechanism of impact was not defined (Findlay and
Houlahan 1997). Plant species richness also was generally less in urban than in riparian forested
streamside corridors near Portland, Oregon (O'Neill & Yeakley 2000). Adverse effects of
watershed urbanization on wetland plants were detected in Minnesota (Galatowitch et al. 1998,
Mensing et al. 1998, Gernes and Helgen 1999) and western Oregon (Magee et al. 1999, Adamus
2001). Alteration of plant communities of Atlantic white cedar wetlands in southern New Jersey
corresponded to the occurrence of roads and housing developments, especially when stormwater
was directed into these wetlands (Ehrenfeld & Schneider 1990). Over a period of 150 years, the
plant communities in a series of Wisconsin lakes appeared to have lost overall biomass and
diversity, and characteristically disturbance-tolerant species became more prevalent. Changes
were attributed broadly to dredging, filling, introduction of carp and two invasive plants,
herbicides, and shoreline raking to remove nuisance macrophytes. Emergent and floating-leaved
species suffered the greatest losses (Nichols and Lathrop 1994).
Physical disturbance of wetland soils during the dry season, such as through tillage, compaction,
or excavation, can increase the dominance of invasive non-native species (Morin et al. 1989,
Sutton 1996, David 1999, Galatowitsch et al. 1999), as well as destroy much of the viable seed
bank (Lee 1991). In central Florida, soil disturbance by feral hogs significantly reduced plant
cover and biomass in a broadleaf floodplain marsh, but increased plant species diversity and
richness (Arlington 1999). Soil tillage often reduces diversity, including both richness and
evenness, as documented in a Carolina bay wetland (Kirkman and Sharitz 1994). The tillage
treatment disrupted the rhizomes of perennials more than burning and also facilitated germination
of annuals in the seed bank and colonization by several invasive species. Persistent seed banks,
perenniality coupled with early sexual maturation, and favorable response to disturbance, were
influential in maintaining wetland flora after disturbance. Dominant perennials persisted
vegetatively, either above or below ground, and were absent from the seed bank.
Invasive plants, especially non-native invaders, significantly alter the species composition of
many wetlands, sometimes even forming nearly monotypic stands. Among the most
geographically widespread invaders in North America are Typha, Phalaris sp., Lythrum
salicaria., Phragmites sp., Myriophyllum spicatum, and Hydrilla verticillata. Their increased
dominance has frequently been viewed as a partial consequence of physical disturbance of soils
or water levels within a wetland and/or the surrounding landscape, including accelerated
sedimentation, eutrophication, and the construction of mitigation wetlands (Confer and Niering
1992, Magee et al. 1999). In Florida, the invasive vine, climbing hempweed (Mikania scandens\
thrives along the edges of wetlands whose long-term water levels have risen (Moon et al. 1993).
Some invasive species, such as Phalaris arundinacea and Leersia oryzoides, prevent sedge
meadow species from recolonizing when they invade constructed or restored wetlands and
(Wienhold and van der Valk 1989, van der Valk et al. 1999). Attempts have been made to
quantify the ability of particular wetland species to out-compete or displace others, and to
identify traits that could be used as a general guide for identifying such species (Keddy et al.
1994, Twolan-Strutt and Keddy 1996, Keddy et al. 1998).
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Removal or decline of non-native species typically results in expanded species richness within
wetlands, at least temporarily (e.g., Trebitz et al. 1993). However, other than their adverse
impacts on native plant diversity, relatively little is known about consequences of many non-
native herbaceous species on wetland functions and attributes such as nutrient processing, carbon
cycling, water balance, and wildlife habitat support (Anderson 1995). Research on this topic has
recently been initiated for some woody species, e.g., non-native saltcedar vs. native cottonwoods
(Stromberg 1998).
Many European studies of restored wetlands are summarized by Pfadenhauer and Klotzii (1996).
Long-term studies of constructed or restored wetlands anywhere are rare. In one instance, a
reforested gravel pit in Ontario was found to differ from natural forested areas even after 107
years, although it had acquired some similar structural characteristics (Larson 1996). In a series
of newly restored wetlands in northern New York, mowing and plowing treatments increased
wetland plant establishment, but less so than did salvaging and importing soil from other
wetlands (ones being altered) (Brown & Bedford 1997). Plowing also significantly increased the
establishment of cattail. Following the importing of soil from wetlands being altered, the species
richness and percent cover were actually higher in the newly restored wetlands than in natural
reference wetlands.
In contrast, after three years of regeneration, the vegetation of a restored prairie pothole was still
not similar to that of a natural wetland which had higher plant diversity (Galatowitsch and van der
Valk 1995, 1996a,b). The regenerating wetland had significantly fewer sedge meadow species
and more submersed aquatic species than the natural wetlands. The restored wetland also had
fewer species represented in its seed bank. Recently restored prairie pothole wetlands in Iowa
lacked distinct low prairie and wet meadow vegetation zones (Delphey and Dinsmore 1993). The
transition from marsh emergents (e.g., cattail, bulrush) to upland vegetation was often abrupt,
perhaps reflecting differences in the exposure gradients of the sites (Shipley et al. 1991b).
In phosphate-mined lands of central Florida, a period of at least 7 years was required before
restored wetlands mostly resembled natural wetlands (Crisman et al. 1997). Non-native and
floating-leaved species dominated in the years immediately following mining. The frequent
abundance of submersed and floating-leaved species in "new" wetlands might be attributed to
either the transportation of seeds on the feathers and feet of waterfowl, or shading by a canopy of
emergent vegetation in more established wetlands (Galatowitsch and van der Valk 1996). Wet-
meadow plants do not readily colonize prairie potholes that were cultivated before restoration
(Galatowitsch 1993). Abroad series of mesocosm experiments involving manipulation of
flooding and nutrients indicated that wetlands with both high plant diversity and low biomass will
be difficult to create "from scratch" (Weiher et al. 1996).
Natural seed banks, which are important to maintaining plant species diversity (Leek 1989, Leek
et al. 1989, van der Valk et al. 1992), can be damaged easily by a variety of human alterations
(Keddy etal. 1989, Wisheu and Keddy 1989, Wisheu and Keddy 1991). Seeds of wetland plants
can be categorized as persistent (remaining viable for many years or decades) or transient (viable
for less than one year), but little is known regarding which species characteristically have
persistent vs. transient seeds, and the degree to which climate, nutrition, contaminants, and other
factors might influence this (van der Valk & Rosburg 1997). Wetland plant communities that are
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most vulnerable to invasion of non-native species following disturbance include those containing
species that have low rates of seed viability and seed dispersal (Reinartz and Warne 1991,
Galatowitsch and van der Valk 1994). Many taxa that recolonize bare sites have seeds that are
naturally adapted for wind dispersal.
Continuous soil disturbance (such as from compaction and road building) can alter species
composition and lead to a decline in both the biomass of native species and alteration of the soil
conditions that support them (Ehrenfeld and Schneider 1991). Off-road vehicular traffic caused
long-lasting shifts in species composition of arctic tundra wetlands (Felix et al. 1992). Use of all-
terrain vehicles also impacted Atlantic coastal plain wetlands, reducing the density of propagules
and seed in wetland seed banks, and allowing common rushes to displace rare species (Wisheu
and Keddy 1991). Excavation and clearing of gas pipeline rights-of-way through forested
wetlands in Florida resulted in increased species richness within the wetland clearings and
increased percent cover of non-native species, primarily Micrantemum umbrosum andPaspalum
notatum (van Dyke et al. 1993, Shem et al. 1994). Higher densities of roads in the vicinity of
forested Ontario wetlands were associated with diminished plant species richness in the wetlands
(Streng et al. 1989). Effects of roads on wetland flora and fauna are also summarized by Forman
and Alexander (1998). In a study of 19 Seattle-area wetlands, wetlands whose contributing
watersheds became developed during the multiyear study experienced a decline in plant species
richness, whereas reference wetlands in both urban and rural settings without ongoing
development changed little between years (Cooke & Azous 2000). Urbanization, with
accompanying alterations to the physical and chemical environment of wetlands, has been
associated with altered wetland plant communities elsewhere, as well (e.g., Manny & Kenaga
1991).
4.12 Wetland Monitoring
Spatial and Temporal Variability
Choice of appropriate sample sizes depends on measured variation in the target taxa and metrics.
Such coefficients of variation for plants are summarized from various studies at:
http://www.im.nbs.gov/powcase/powvariation.html
Spatial and temporal variation of plant species composition at the scale of individual wetlands is
influenced by connectivity of suitable habitats (Tabacchi et al. 1990), competition (McCreary
1991, Keddy et al. 1994, 1998), water regimes, water chemistry, and other factors described
previously in this chapter. Spatial variation in plant diversity was reported to be greater among
less-degraded than among more-degraded streams near Chesapeake Bay (Small et al. 1996).
Along transects that cross floodplains, plant richness often peaks midway between the base flow
channel and seldom-flooded uplands (Gregory et al. 1991), although in some meandering
lowland rivers, richness can be greatest immediately adjoining the channel (Stromberg et al.
1997). Temporal variation (interannual change) in plant species richness in 19 Seattle-area
wetlands was reported to be statistically insignificant over an 8-year period (Azous & Cooke
1997). Densities of submersed macrophytes, however, often change dramatically from year to
year (Blindow 1992).
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Along streams feeding Chesapeake Bay, information derived from rapid determinations of plant
species richness along a series of only 5-10 survey lines per stream (each line being 5 meters long
and usually 1 m wide, paralleling and touching the channel) was sufficient to distinguish streams
which other data showed were polluted from those that were not (Small et al. 1996). When
species relative abundance was also considered, predictions of water quality were even more
accurate. The best-quality streams averaged 40 species (cumulative) per 20 survey lines. The
survey lines were 15m long and followed the stream bank.
In many wetlands and floodplains, plant richness tends to increase initially as community
biomass increases, but at some point begins to decline with continued increases in biomass
(Shipley et al. 1991, Tilman 1996). However, biomass alone is usually a poor predictor of species
richness in wetlands, often being secondary to more direct effects of environmental factors -
particularly light availability - and evolutionary history (Gough et al. 1994, Grace 1999).
Techniques
Fundamental to the use of plants as indicators of wetland condition are practical procedures for
assessing plant assemblages. With good reason, there is no single accepted approach for
sampling wetland and riparian plants. That is because sampling design should depend largely on
sampling objectives, i.e., for which particular attributes of plant community structure is
information needed, and how quantitative does it have to be? Procedures that are used often
when employing plants to characterize wetlands are summarized in Table 4.3, and include the
following, applied either independently or in combination:
• Unstructured searches
• Systematic transects
• Random plots
• Stratified plot-based surveys
Table 4.3. Types of procedures commonly used to characterize wetland and riparian
plant assemblages
Unstructured searches have the advantage of being the quickest and least restrictive option. They generally involve
one skilled botanist walking the entirety of a site while keeping a running list of species noticed (sometimes called a
"random walk," Planty-Tabacchi et al. 1996). This approach is applicable when the only objective is to assess plant
richness and species composition, not percent cover. Disadvantages include the fact that results are strongly
influenced by (a) time spent searching per unit wetland area, (b) size and complexity of the site, (c) keenness and
taxonomic skills of the searcher, (d) inaccessibility of parts of the site, e.g., deep water. To improve somewhat the
comparability of estimates among different sites, searches can be restricted by time (e.g., 10-minutes per acre, or
search until no more species found after 3 minutes of searching) and/or by stratifying the search by recognizable
habitats within a site.
Systematic transects are commonly used, especially in research studies, to assess wetland and riparian plant
communities (e.g., Winward 2000). Transects consist of sample plots or observation ("intercept") points located,
usually at even intervals, along generally straight lines. The transect lines also are usually spaced at even distances
apart. They may be oriented perpendicular to the long axis of a site, may radiate from the centerpoint of the site, or
be oriented in some other manner intended to span moisture gradients. If the number of transects and/or the number of
plots or points per transect is sufficient, information on spatial dominance of particular species throughout the site
can be obtained. The even spacing of an inadequate number of transects within a site can be a disadvantage because
it will be insensitive to (and may totally miss) important environmental gradients that influence a site's plant
communities. The transect may unknowingly follow a linear feature of the vegetation, such as a former farm road that
has since overgrown, and as a result a row or column of plots may be spatially correlated, compromising the
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statistical reliability of the data. These issues can be addressed somewhat by sampling large numbers of plots or
points along the transects, and/or by using numerous transects per unit of wetland area. Transect approaches are
specified for assessing relative dominance of wetland-associated plants in the Corps of Engineers wetland
delineation manual. The manual's procedures, when applied to an average 2-acre site, would require 6 plots. At
least 100 plots per site would be required using procedures employed in studies of Portland-area wetlands by EPA's
National Health and Environmental Effects Research Laboratory (Magee et al. 1999). Researchers studying restored
wetlands in West Eugene, Oregon, calculated that 200 intercept points, spaced equally throughout a site, were
required to derive estimates of species cover that are confidently within 5-10% of the true values in their wetlands.
Random plots are typically used at sites perceived either as lacking recognizable environmental gradients, or with
highly complex gradients. Standard-sized plots are situated according to X-Y coordinates generated by a random
numbers table, or other random number generator. No assumptions are made regarding locations of particular
gradients that may influence plant distribution. If the number of plots is sufficient, statistically-sound information on
spatial dominance of particular species throughout the site can be obtained.
Stratified plot-based procedures also are sometimes used to assess wetland plant communities. This involves using
professional judgment, rather than solely systematic or randomized designs, to situate plots or observation points.
One option is to place plots in "representative" locations, using judgment to identify locations that seem most typical
of the site. This may encounter problems with repeatability. Another option is to stratify a site according to plant
communities or associations (i.e., commonly-correlated assemblages of species), and then sample each association
with one or more plots, located randomly or systematically within each association. This requires judgment to
recognize and delimit what constitutes plant "associations," inasmuch as no generally-accepted list exists in many
regions. To avoid problems with defining associations, one can place plots to include every plant species that
appears, from an initial site reconnaissance, to constitute more than a prespecified acreage or percent of the site. To
conserve sampling effort, and if a primary objective is to assess species richness, one can select the fewest plot
locations that will produce the largest species list. However, this requires careful screening of the site and strategic
planning to identify - quite subjectively — the most complementary and species-rich locations. Finally, one can
stratify a site by observable physical and chemical features thought to influence plants, such as shade and/or expected
duration of inundation (elevation), and then allocate plots randomly or systematically within each "zone." This also
requires judgment to define and locate such gradients, and may have low repeatability when used by different
researchers studying the same wetland at different times.
It should be noted that the intensity and statistical rigor appropriate for some plant surveys (e.g.,
research studies, or periodic monitoring of mitigation sites) is not necessarily appropriate or
necessary for studies whose aim is solely to develop plant IBI's or to conduct one-time
assessments of relative ecological condition of a series of wetlands. If plant surveys must be
limited to only brief and/or infrequent visits to a site (as is often the case when implementing a
regional assessment program), it is highly unlikely that enough plots can be surveyed to yield
statistically-sound estimates of the percent of the site occupied by each species. This will be true
regardless of whether plots are located systematically or randomly. On the other hand, a one-day
visit, especially during the growing season and using unstructured searches or stratified plots, will
usually be sufficient to determine presence of a large percentage of species occupying a site.
However, this implies a tradeoff between speed and repeatability (consistency) of results.
Metrics for Assessing Impacts to Wetland Plant Communities
Species richness (and diversity) can increase (Grassle 1989, Kaczor and Harnett 1990, Phillips et
al. 1994) or decrease (Brown andBrussock 1991, Englund 1991, Wilson and Tilman 1991) along
disturbance gradients, whether natural or human-related. Wetland plant richness also varies
latitudinally within continents but shows no clear geographic pattern (Crow 1993). Plant species
richness in raised bogs is remarkably uniform along a latitudinal gradient from the northeastern
US into central Canada, averaging 20 to 26 species (Glaser 1992). The richest bog community
occurs on maritime islands off of the northeastern US, with species richness ranging from 50 in
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the south to 32 in the northeastern Canadian sites. The southern range of some bog species is
determined by maximum summer temperatures which accelerate metabolic consumption rates to
lethal levels (Crawford 1989). Thus, to be useful as metrics, richness or diversity should be
calibrated regionally, and in any case may not be good indicators of human alterations to wetland
plant communities if they are used alone. Data on species composition and other metrics is
generally needed as well.
A survey of 22 forested wetlands in the St. Jones watershed of Delaware, using plant species
composition and richness as indicators, failed to find strong statistical relationships to apparent
watershed condition (Emslie & Clancy 1999). In Minnesota, a wetland plant IBI reflected
contamination of wetland water with chloride and sediment contamination with copper (Gernes
& Helgen 1999). A statistically weaker response of wetland plants to ambient phosphorus was
noted. In western Oregon, the percent of total plant species at a site that are native was found to
correlate negatively with several indicators of human alteration estimated at both the site-scale
and landscape-scale (Adamus 2001).
Species accumulation curves sometimes can be used to assess degradation and to identify the
minimum number of plots per site required to determine this, e.g., by testing for significant
differences in slopes of curves from presumably degraded vs. undegraded sites. In streams of the
Chesapeake watershed, the cumulative total of shoreline plant species tended to increase slowly
as more degraded sites were surveyed, whereas the cumulative total of shoreline plant species
increased significantly more rapidly, and reached a higher total, as a comparable number of less-
degraded sites were surveyed (Small et al. 1996). Graphs of the number of individual plants per
species vs. number of species ("Preston curves") demonstrated separation of altered and
unaltered streams. In Chesapeake Bay itself, the distribution and biomass of submersed aquatic
macrophytes were found to be a reliable indicator of water quality (Dennison et al. 1993).
For years, botanists have attempted to identify species traits that would allow rapid categorization
of species according to sensitivity to specific impacts or to habitat alteration generally. Examples
include attempts by Grace (1990), Wood & Tanner (1990), Mclntyre et al. (1995), and Boutin &
Keddy (1993) as elaborated by Keddy (2000). Such categorizations of species into "guilds" or
"functional groups" could greatly expedite the development of successful multimetric indices.
However, for an enormous number of wetland plant species, essential prerequisite information
about life history is unknown, as is also the case regarding characteristics that reliably distinguish
sensitivity to human influences from sensitivity to natural phenomena.
Paralleling the increasing use of plants for assessing wetland condition has been the development
of improved conceptual and computer models for predicting long-term sustainability of
populations and restored wetland plant communities, given particular hydrologic scenarios for
woody plants (e.g., Stromberg et al. 1993, Richter & Richter 2000), herbaceous plants (Shipley et
al. 1991a), or wetlands generally (Poiani & Johnson 1993, Weiher & Keddy 1995, 1999, Weiher et
al. 1998, Hill et al. 1998). These hold promise for characterizing disturbance gradients across
landscapes, as well as for improving water management and the design of constructed wetlands,
so that wetland plant communities are better protected. Also, increased attention has focused on
use of satellite sensors for detecting stands of stressed vegetation, and especially stands of easily
recognizable non-native or invasive species. The potential for use of biomarkers (chemical
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signatures of individual plants) in detecting wetland alteration also is being explored, e.g., Miller
et al. (1993).
4.13 Literature Cited
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Environmental Research Laboratory, Corvallis, OR.
Adamus, P.R. 2001. Guidebook for Hydrogeomorphic (HGM)-based Assessment of Oregon Wetland and Riparian Sites.
I. Willamette Valley Ecoregion, Riverine Impounding and Slope/flats Subclasses. Volume IB. Technical Report. Oregon
Division of State Lands, Salem, OR.
Adamus, P.R., andK. Brandt 1990. Impacts on Quality of Inland Wetlands of the United States: A Survey of Indicators,
Techniques, and Applications of Community-Level Biomonitoring Data. EPA/600/3-90-073. Office of Research and
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http: //www. epa. go v/o wo w/wetlands/wqual/intro web. html
Adamus, P.R. and A. Gonyaw. 2000. National Database of Wetland Plant Tolerances. Prepared for the USEPA. Internet
address: http://www.epa.gov/owow/wetlands/bawwg/publicat.html
Albers, P.H., andM.B. Camardese. 1993. Effects of Acidification on Metal Accumulation by Aquatic Plants and
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seedlings to salinity. Forest Ecology & Management 70(1-3): 203-214.
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Anderson, D.S. and R.B. Davis. 1997. The vegetation and its environments in Maine peatlands. Can. J. Bot. 75:1785-
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Barko, J.W., D. Gunnison. and S.R. Carpenter. 1991. Sediment interactions with submersed macrophyte growth and
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Barrick, K.A. and Noble, M.G. 1993. The iron and manganese status of seven upper and montane tree species in Colorado,
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Baskin, C.C., Chester, E.W. and Baskin, J.M. 1996. Effect of flooding on annual dormancy cycles in buried seeds of two
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Begin, Y. and Payette, S. 1991. Population structure of lakeshore willows and ice-push events in subarctic Quebec,
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Borth, C.S. 1998. Effects of land use on vegetation in glaciated depressional wetlands in Western Montana. Thesis,
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Section 5. Invertebrates
5.1 Use as Indicators
This section describes invertebrates (including insects) that are commonly associated with wetlands,
and their response to human-associated changes to wetlands. This section includes many species of
(for example) midges, aquatic worms, dragonflies, snails, and water beetles. This section summarizes
only the scientific findings on this topic that have been published since 1989. For a general discussion
of the topic, and for scientific information published before 1990, readers should refer to Adamus and
Brandt (1990).
Of particular note is the book recently published by Batzer et al. (1999) and compilations of
information on invertebrates in prairie wetlands (Adamus 1996). Sampling protocols, metrics, and
indices have been well developed, at least for stream and lake biomonitoring (Plafkin et al. 1989,
Klemm et al. 1990, Barbour et al. 1992, Barton & Metcalfe-Smith 1993, Rosenberg & Resh 1993,
Kerans & Karr 1994). However, these are not uniformly transferable to wetlands. Nonetheless, a
growing body of literature is addressing the sampling of invertebrates in wetlands. Advantages and
disadvantages of using macroinvertebrates as indicators of wetland condition were summarized by
Adamus and Brandt (1990).
In the past decade, several studies have used invertebrate assemblages specifically to indicate the
ecological condition of a large series of wetlands. Results have been published from such studies, for
example, in the Seattle area (Ludwa 1994, Ludwa & Richter 2000), Massachusetts (Carlisle et al.
1998), Montana (Apfelbeck 1998), and Minnesota (Gernes & Helgen 1999, Mensing et al. 1998), and
the Great Lakes (Burton etal. 1999). Most of these studies are detailed at:
http://www.epa.gov/owow/wetlands/bawwg/case.html.
In Minnesota, wetland invertebrates were used to represent the condition of landscapes that contained a
large riparian wetland component (Galatowitsch et al. 1998, Mensing et al. 1998). Sampling 15
wetlands belonging to each of 8 wetland types, the investigators found positive or negative correlations
of several metrics with a site disturbance score and/or various land cover types measured within 500,
1000, and 2500 m of each wetland. Results are reported by wetland type at:
http://www.hort.agri.umn.edu/mnwet/ .
Other efforts to develop wetland indices of biotic integrity (IBI's) using invertebrates are underway in
Ohio, Michigan, Maine, Florida, and elsewhere. As a partial starting point for such indices, Adamus
and Gonyaw (2001) compiled literature and prepared a documented species database for EPA's
internet web site, that categorizes many wetland invertebrates as tolerant or intolerant, with regard to
overall sensitivity, and/or specific sensitivity to excessive nutrients and hydrologic alteration:
http://www.epa.gov/owow/wetlands/bawwg/publicat.html
Florida Department of Environmental Protection has developed a wetland "Bio-
scores macroinvertebrate taxa present in dip sweeps using a weighted index for sampled taxa. Three
metrics are currently computed: total taxa richness, total lake index, and total number of
Ephemeroptera, Trichoptera, and Odonata taxa. Wetland health or impairment is determined based on
deviation from target values. The scoring and selection of metrics is based on similar procedures for
Florida streams and lakes and judgment of local biologists. Florida also is developing a biomonitoring
program for canal systems using a modified stream condition index protocol.
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In drier wetlands (those not inundated for long periods), it is often impossible to use many of the
traditional assemblages of aquatic invertebrates as indicators of ecological condition because these
assemblages require permanent inundation. As such, their use would erroneously imply that "good
condition" corresponds with "long duration flooding." In drier wetlands, surveys of soil fauna (e.g.,
earthworms, nematodes), including their dormant stages, have the potential to accurately represent
ecological condition (Linden et al. 1994). This has been demonstrated in many European studies (e.g.,
Goede & Bongers 1994, Korthals et al. 1996, van Straalen & Verhoef 1997, van Straalen 1998, Gyedu-
Ababio et al. 1998, Ritz & Trudgill 1999, Bongers & Ferris 1999, and Urzelai et al. 2000) but only
recently in some North American studies (Lau et al. 1997, Ettema et al. 1998, 1999).
5.2 Effects of Enrichment, Eutrophication, Reduced Dissolved Oxygen
Processes
Excessive nutrients can cause long-term or short-term shifts in invertebrate community richness,
abundance (density), and species composition. These changes are typically triggered when excessive
nutrients lead to greater growth of aquatic plants, and in particular the increased dominance of certain
kinds of algae. Invertebrate species that happen to be specialized to feed on these algae, or which
characteristically find shelter and attachment sites in the aquatic plants, are then favored (Murkin et al.
1991, Campeau etal. 1994, Moore et al. 1993). However, excessive nutrients can trigger severe
outbreaks in bacterial taxa harmful to invertebrates; bacterial infestations covering more than 25% of
the exterior of individual mayfly specimens were especially lethal (Lemly & King 2000). In addition,
respiration and decay of extensive algal biomass can reduce dissolved oxygen in the water column and
sediments to levels critical to many wetland invertebrates. Taxa with shorter generation times are
especially likely to respond to nutrient increases, and thus maintain a competitive advantage, because
the blooms of algae associated with enrichment are often short-lived. In Massachusetts (Carlisle et al.
1998) and Minnesota (Gernes & Helgen 1998), indices of wetland biological integrity using
invertebrates were correlated negatively with nutrient-laden stormwater inputs to wetlands.
Ammonium fertilizers commonly used in agriculture can be lethal to earthworms (Linden et al. 1994).
Long-term shifts can occur when excess nutrients continue to cycle even after enrichment is abated
(Graves et al. 1998). Temporary shifts can result from episodic or chronic inputs to wetlands that are
less retentive of nutrients (Gabor et al. 1994). There may be some level of nutrient input that has no
measurable affect on community structure as it lies within the environmental tolerance range of all the
major organisms of a wetland's food web (Cooper 1993).
Effects on Species Richness
Up to some point, nutrient inputs to wetlands can lead to increased invertebrate richness, as more food
sources become available to predatory invertebrates (Moore et al. 1993, Rader and Richardson 1992,
Campeau etal. 1994, Cieminski & Flake 1995, Gernes and Helgen 1999). However, invertebrate
richness in a series of highly enriched wastewater wetlands was found to be lower than in a less
enriched reference wetland (Nelson et al. 2000). Excess nutrients in lakes have caused local extinction
of some macroinvertebrates and an overall decline in species diversity (Mason 1991), partly due to
associated diminuation of dissolved oxygen. Similar responses have been noted in enriched streams
(Belong and Brusven 1998). However, a study of Florida lakes found no correlation between nutrient
levels and zooplankton (copepod) diversity (Blancher 1984).
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Effects on Species Composition
Studies conducted in the last decade continue to strengthen the premise that significant enrichment
causes the species composition of wetland invertebrate communities to shift, and that in some cases the
characteristics of such shifts can be used to diagnose enrichment as the cause of altered wetland
condition. Literature describing the sensitivities of individual species of North America wetland
macroinvertebrates to nutrient enrichment has recently and exhaustively been compiled for EPA in a
public database (Adamus and Gonyaw 2000). Such information is especially useful because
characterizations of invertebrate species according to the trophic state of lakes, or as pollution-tolerant
or intolerant in streams (e.g., Hilsenhoff 1982a, 1982b, Rosenberg and Resh 1993, Patrick and
Palavage 1994), cannot be assumed to always be accurate for wetlands.
Because the above-mentioned EPA database covers the recent literature regarding species composition
shifts, the discussion of that topic here will be limited. Exposure to organic enrichment and
eutrophication frequently causes an increase in macroinvertebrate grazers (such as Tanypodinae
midges), and herbivores, detritivores, predators, and "miners" that burrow into macrophytes. These are
groups that typically increase with increasing growth of periphyton and emergent aquatic plants
(Campeau et al. 1994). In Florida, Graves et al. (1998) noted that oligochaetes, ancylid gastropods,
and midges were more dominant in enriched wetlands, while in similar unenriched wetlands (e.g., 8
ppb phosphorus and 783 ppb nitrogen) the dominant taxa were caddisflies (Trichoptera), talitrid
amphipods, and a different midge. Differences were possibly due to the intolerance of the latter taxa to
low oxygen conditions present in the substrate and water of the enriched wetlands. Also in the Florida
Everglades, Rader and Richardson (1992, 1994) reported a greater number of coleopteran species
(especially in the Hydrophilidae and Dytiscidae families) in enriched and intermediate areas than in
unimpacted sites (total mean annual density of macroinvertebrates at enriched and intermediate sites
was 6.1 and 3.5 times greater, respectively, than in the unenriched area). Except for decapods,
especially Palaemonetespaludosus, the density of each order or class was higher within enriched and
intermediate areas. Percent composition measured 2.6 times higher and density of dipterans as 16.2
times greater at enriched and intermediate sites than at the unenriched site. Dominant dipterans at
enriched sites were Dasyhelia spp., Goelkichironomus holoprasinus, Larsia decolorata, Polypedilum
trigonus, Pseudochironomus spp., and Tanytarsus sp. J. The number of taxa (primarily Chironomidae)
did not increase, and was very similar for all sites.
Other studies, specifically focusing on midges, have found that at the subfamily level, Chironominae
and Tanytarsinae, which contain hemoglobin and thus are more tolerant of reduced dissolved oxygen
levels, appear to replace Orthocladinae as eutrophy increases (King and Brazner 1999). A study of
four lacustrine/bay wetlands bordering Lake Michigan also found that midge communities shifted
across nutrient gradients. Oligotrophic bays with low conductivity (190-230 |j,S/cm) were dominated
by Cladotanytarsus sp., Orthocladius sp. and Heterotrissocladius changi, whereas more eutrophic
bays with higher conductivity (390-450 |j,S/cm) were dominated by Chironomus sp., Tanytarsus sp.
and Cricotopus sp. Based on dosing experiments by Murkin et al. (1994) and Campeau et al. (1994),
it can be surmised that a major shift in invertebrate species composition can occur in northern
lacustrine marshes at concentrations of 60-200 mg/L phosphorus and 1600-4000 |J,g/L nitrogen.
Effects on Total Abundance or Biomass
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Recent literature continues to substantiate the tendency of total invertebrate density to increase with
increased nutrients, as algal production becomes less of a limiting factor in the invertebrate community
(Murkin et al. 1991, Campeauetal. 1994, Moore et al. 1993). Invertebrate populations in nutrient-
poor systems, especially in the arctic and subarctic, are especially quick to respond (Hershey 1992,
Hiltner & Hershey 1992, Hinterleitner-Anderson et al. 1992, Bartsch 1994). A survey of 20 Nova
Scotia lakes also found a positive relationship between zooplankton density (1-5 individuals/m3) and
total phosphorus (5-20 mg/m3) (Kerekes et al. 1990). At least in wastewater systems, total density
may continue to increase, or at least not decrease, even when dissolved oxygen deficits that are
associated with enrichment become severe (Nelson et al. 2000). However, in enriched Everglades
wetlands, depletion of dissolved oxygen supplies caused by high oxygen demand appears to be the
most important mechanism leading to the decrease in macroinvertebrate abundance (Mason 1991).
Similarly, nutient-rich feedlot runoff can be lethal to several invertebrates due to its high oxygen
demand (McCahon et al. 1991). A study of 3 Canadian marshes found that enrichment initially caused
a reduction in total abundance of invertebrates, but as vegetation became more fully decomposed and
oxygen levels rose, so did invertebrate abundance (Gabor etal. 1994).
Removal of nutrients can have a cascading effect of the trophic structure of a water body. Following
the cessation of sewage inputs to an English lake, phosphorus levels fell from an average of 155 |j,g/L
P to 78 |j,g/L (Moss et al. 1996). This resulted in an almost immediate drop in the total chlorophyll
level and a subsequent drop mDaphnia sp. numbers.
Effects on Metal Bioaccumulation
Nutrients appear to influence the tendency of aquatic invertebrates to accumulate heavy metals, and the
type of metals that are accumulated. For instance, zinc, iron and manganese concentrations were
higher in midges from nutrient-rich wetlands, whereas high copper concentrations were found in
midges from nutrient-poor wetlands (Bendell-Young et al. 1994). This may be due at least partly to
the bioavailability of various metals being influenced by sediment oxygen conditions, which in turn are
partly the result of decomposition of algal blooms triggered by high nutrient concentrations.
5.3 Effects of Contaminant Toxicity
The following subsections first review the effects of metals and then the effects of organic and
synthetic compounds such as pesticides.
Effects of Heavy Metals
Processes
Heavy metals such as mercury, lead, zinc, copper, and cadmium can be directly toxic to wetland
invertebrates, or can impact invertebrate communities by altering the species composition and
abundance of algae and aquatic plants upon which invertebrates depend for food and shelter. Wetlands
store heavy metals to an even greater degree than lotic systems (Gambrell 1994), so exposure of
invertebrates to metals may be greater in wetlands. Even when present at low background levels,
metals can easily be bioaccumulated in wetland invertebrates (Hare et al. 1991). Growth, larval
development, and reproduction of invertebrates can also be harmed by long-term exposure to sub-
lethal concentrations of trace metals (Timmermans 1993). Relatively little is known of the sub-lethal
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effects and the fate and balance (e.g., metabolization and accumulation) of metal pollutants in
freshwater wetlands (Johnson etal. 1993). Under some conditions common in wetlands (especially
wastewater wetlands), high levels of iron are toxic to invertebrates, both directly and through alteration
of habitat structure with floe layers (Rovers 1998).
The extent to which heavy metals are toxic to wetland invertebrates depends largely on the acidity of
the wetland and the particular form of the metal involved. Acidic conditions can mobilize and increase
the toxicity of some metals, such as cadmium (Wright & Welbourn 1994), and decrease the toxicity of
others, such as aluminum (Wren and Stephenson 1991). However, in a British stream, acidic
conditions, combined with high aluminum concentrations increased the mortality of the amphipod
Gammaruspulex and the mayflies Baetis rhodani and Ephemerella ignita (McCahon and Pascoe
1989). Some metals, such as iron and aluminum, can to some degree protect invertebrates from
otherwise toxic effects of heavy metals in acid mine drainage (Whipple and Dunson 1992).
Effects on Richness
Declines in aquatic invertebrate richness have been documented in watersheds with greater percent
urban land cover and presumably larger loadings of heavy metals, compared with less-urbanized
(control) watersheds (Shutes et al. 1993, Casper 1994, Winter and Duthie 1998, Gernes and Helgen
1999). A Montana stream exposed to elevated levels of heavy metals from mining operations had
significantly reduced species richness (Poulton et al. 1995). More than 20 years after cadmium and
cobalt discharges to a freshwater marsh in New York were curtailed, invertebrate richness remained
lower than at a control (less-polluted) site (Klerks and Levinton 1993). Moderate recovery of
invertebrates from metal contamination was demonstrated in the Coeur D'Alene River in Idaho. Over
22 years following cessation of contamination by zinc and other metals, the number of taxa grew from
0 to 18, while the proportion of mayflies, stoneflies, and caddisflies relative to proportion of midges
rose (Holland & Rabe 1992, Holland etal 1994). Recovery of aquatic systems from toxic pollutants is
summarized by Cooper (1993).
Effects on Species Composition
In general, gastropods, crustaceans, and molluscs are more sensitive than insects to metal exposure
(Johnson etal. 1993). Amphipods, midges, and mayflies, have been used successfully in field and
laboratory studies to detect metal-related sediment toxicity (Burton 1992, Adams et al. 1992a, Cain et
al. 1992, Clements et al. 1989a). Some studies show herbivores and detritivores being the most
sensitive to metal additions (Kiffney and Clements 1994a, Leland et al. 1989), whereas others have
reported scrapers being the most sensitive group (Clements 1994).
In general, mayflies and some stoneflies of western streams are sensitive to metals, whereas caddisflies
and midges are relatively tolerant (Clements 1994, Kiffney and Clements 1994b, Leland et al. 1989,
Nelson and Roline 1996). Following abatement of zinc pollution in a Colorado stream, the mayfly
Rhithrogenia hageni and the dipteran Pericoma recovered and became the dominant benthic organism.
The mayflies Baetis andEpeorus, as well as Chloroperlid stoneflies, the caddisfly Artctopsyche
grandis, and midges did not appear to be significantly affected by the pollution abatement (Nelson and
Roline 1993). A Montana stream exposed to elevated levels of metals lacked mayflies and stoneflies,
and had more acid-tolerant/metal-tolerant taxa such as the midge Cardiocladius sp. (Poulton etal.
1995). However, midge richness and density (with the exception of Procladius sp. and Chironomus
sp.) decreased along a gradient of increasing trace metals in the Buffalo River, New York (Diggins and
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Stewart 1998). Among the midges, Orthocladinae are known to tolerate moderately elevated
concentrations of metals in some situations (Clements 1994) whereas Tanytarsini are sensitive to
copper (Clements 1989a). Some mayflies are more sensitive early in their development and become
less sensitive as they mature (Kiffney and Clements 1994a, Diamond etal. 1992). Freshwater
amphipods (Gammaruspulex) suffered higher mortality and behavior alterations when exposed to high
doses of copper (Taylor et al. 1994). An index of macroinvertebrate community sensitivity to copper
and other heavy metals was proposed by Clements etal. (1992).
Effects on Abundance, Growth, Behavior, Deformities
There was no significant difference in the density of midges and oligochaetes between a polluted
freshwater marsh in New York and a control (less-polluted) marsh. Laboratory tests on the midges and
oligochaetes from the two wetlands suggested that the oligochaetes from the heavily polluted site were
genetically more resistant to cadmium toxicity (Klerks and Levinton 1993). Agricultural drainage
water containing arsenic, boron, lithium, and molybdenum and entering the Stillwater Wildlife
Management Area in Nevada proved acutely toxic to many wetland invertebrates (Hoffman 1992,
Hallock and Hallock 1993a,b). Copper and some other heavy metals appear to be more damaging to
aquatic communities in the spring and summer rather than in the fall (Leland et al. 1989). It is thought
that the summer dosing coincides more closely with hatching of many macroinvertebrates and that
early developmental periods may be more susceptible.
Deformities of midge mouthparts have been used as an indicator of heavy metal contamination (Bird
1995) and pollution generally (Lenat 1993). Wetland invertebrates exposed to fertilizer factory waste
had an increased rate of developmental deformities (Clarke 1993), perhaps as a result of metals
incidentally associated with the fertilizer. However, invertebrates in some Canadian wetlands
contaminated with oil did not have an unusually high rate of deformities (Bendell-Young et al. 2000).
In Ontario wetlands and streams, morphological deformities in the labial plates of midges were
associated with agricultural, industrial, and domestic pollutants (Dickman etal. 1992b, Dickman and
Rygiel 1996). Incidence of deformities at a control site, upstream from the source of pollutants, was
9% compared to 47% downstream from the source. Similarly, in waters contaminated with coal tar the
midge Chironomus anthyracinus had significantly higher levels of head structure deformities
(Dickman et al. 1992a). Another study recommends that measures of both the weight and head
capsule width of Chironomus larvae can be used as endpoints for toxicity tests to differentiate reduced
growth from retardation of instar development (Day etal. 1994). Observation of behavioral responses
of macroinvertebrates has been proposed as one rapid, cost-effective means of assessing sub-lethal
exposures to contaminants (Heinis et al. 1990)
Bioaccumulation
In general, higher metal concentrations are found in animals from polluted sediments (vanHattum et
al. 1993). Moreover, increasing the benthic concentrations of heavy metals, such as zinc and copper,
generally increases the concentrations within the invertebrates (Miller et al. 1992, Kiffney and
Clements 1993). In 3 Canadian lakes, concentrations of heavy metals in the crayfish Cambarus
bartoni were correlated with sediment/water concentrations (Alikhan et al. 7990). In roadside ditches
of Louisiana, concentrations of aluminum, lead and cadmium in the crayfish Procambarus clarkii were
significantly higher than in commercial catches unaffected by road influences (Madigosky et al. 1991).
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Monitoring the accumulations of heavy metals in invertebrate tissues can be a more accurate measure
of the effects of metals than measuring metal concentrations in the water column (Kiffney and
Clements 1996). In at least some instances, the aquatic insects that feed on periphyton or detritus
accumulate higher concentrations of metals than predatory insects (Kiffney and Clements 1993). The
aquatic mite Limnesia maculata and the caddisfly Mystacides accumulate trace metals from both the
surrounding water and from the contaminated midge larvae they feed on (Timmermans et al. 1992).
Bioaccumulation often affects the condition of individual animals and thus, presumably, population
health and abundance. For example, the dry weight of individuals of the mollusc Anodonta grandis
declined as cadmium concentrations in tissues of the species increased (Couillard etal. 1993).
Several factors have been found to influence bioaccumulation rates. In northwestern Ontario, methyl
mercury concentrations in tissues of invertebrate taxa from the Odonata, Corixidae, Gerridae,
Gyrinidae, and Phryganeidae/ Polycentropodidae exhibited increases in concentrations in response to
flooding (Hall et al. 1998). Conductivity explained much of the variability in mercury concentrations
of crayfish from 13 Ontario lakes (Allard and Stokes 1989). Warmer temperatures caused increased
accumulation of copper and cadmium in the isopodAsellus aquaticus, whereas lead accumulation
decreased with warming temperature.
Acidification did not appear to affect copper, cadmium, and lead accumulation (van Hattum etal.
1993). However, aluminum can interact with acidity to produce greater toxic effects. In one study, the
combination of acidity (pH 4.9) and aluminum caused the highest invertebrate mortality, and the
mortality decreased when aluminum was complexed with citric acid (McCahon and Pascoe 1989).
Another study suggests that copper has greater toxicity to macroinvertebrates in water that is acidic
and soft than in water that is more alkaline and hard (Clements et al. 1989b). Cadmium concentrations
in freshwater clams were most closely associated with water column pH in 21 Ontario lakes (Campbell
and Evans 1991). Cadmium absorption by a freshwater clam, Uniopictorum, was influenced by
cadmium concentration and water temperature. The process of cadmium accumulation by freshwater
clams is rapid and appears to be irreversible (Jenner et al. 1991).
Acidic conditions can also affect bioaccumulation of metals. Midges from acidic lakes had markedly
higher concentrations of aluminum, cadmium, magnesium, and zinc in their tissues than midges from
less-acidic lakes (St. Louis 1993).
In general, wetlands receiving acid-mine drainage with low pH and large amounts of dissolved
minerals (aluminum, copper, iron, manganese, and zinc) have considerably lower macroinvertebrate
abundance and diversity (Short et al. 1990). After reclaiming a wetland at an abandoned mine site, the
acidity and toxicity of metals decreased and the abundance of aquatic invertebrates increased (Fucik et
al. 1990). When acid-mine drainage was allowed to reenter the wetland, the aquatic invertebrates
decreased in abundance as water quality deteriorated. Aquatic invertebrate populations appeared to be
better indicators of the degree of recovery than water samples.
Macroinvertebrate communities in streams exposed to mine tailings and seepage from a molybdenum
mine had somewhat lower diversity and richness than control sites, and tended to be dominated by
pollution-tolerant species (Whiting et al. 1994). Platyhelminthes (Polycelis sp.), Ephemeropterans,
Trichopterans, and chloroperlids declined in abundance in sites receiving mine tailings. In contrast, the
orthoclad midges Eukiefferiella sp. and Parametriocnemus sp. exhibited the opposite trend. The sites
immediately downstream from the tailings had larger numbers of tubificid worms, psychodids and
tipulid larvae. In general, collector-gatherers made up a larger proportion of the community in
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contaminated sites, while grazers and predators were less prevalent (Whiting et al. 1994). Sites
receiving open mine pit drainage, appearing to contain insufficient molybdenum and cyanide to
severely impact the invertebrate community, had somewhat higher densities and richness than control
sites. The mayfly Ameletus sp. and the caddisfiies Allomyia sp. and Chyrandra sp. were absent from
the sites that received open pit drainage. The open pit drainage sites were dominated by naidids,
Hydracarina, Ostracods, Amphipods, hydroptilids, and ceratopogonids, which were all rare or absent at
control sites. It is unclear why the open pit sites had higher richness, although it is thought that it
might contain higher aqueous nitrogen concentrations that led to increased algal growth.
Effects of Pesticides, Oil, and Other Contaminants
Processes
These substances can alter community structure by (a) being acutely or chronically toxic to
invertebrates, (b) altering algal communities and aquatic macrophytes upon which some invertebrates
depend for food and shelter, (c) altering predation on invertebrates by decimating numbers of other
crustaceans, fish, and amphibians, (d) reducing available dissolved oxygen (i.e., chemical oxygen
demand) or oxygen diffusion rates (e.g., oil), and (e) altering the effects of other potential stressors,
such as acidity. Macroinvertebrates can bioaccumulate some complex pesticides. For example,
Chironomus riparius bioaccumulates flouranthene and benzo[a]pyrene (BAP) (Clements et al. 1994).
Toxicity Effects
With the advent of the mosquito-borne West Nile virus in parts of the United States, nearly all
wetlands in some localities are being dosed with non-selective hydrocarbon pesticides. Few field
trials of these pesticides have closely examined biological effects in wetlands, so inference must be
made from the limited number of laboratory and stream studies.
In laboratory tests the insecticide diflubenzuron was most toxic to crustaceans, followed by mayflies,
midges, and caddisflies (Eisler 1991b, 1992). Showing moderate sensitivity were the larvae of
corixids, dragonfly adults and larvae, spiders, dytiscids, and ostracods. Also lethal to invertebrates
were paraquat (Eisler 1990), cyanide (Eisler 199la), fenvalerate (Eisler 1992b), and acrolein (Eisler
1994). Following toxaphene application in an Alberta lake, the zooplanktonic invertebrate Bosmina
sp. was reduced by 88% (Miskimmin and Schindler 1994). The synthetic pyrethroid insecticide
fenvalerate, when present in saltmarsh sediments, did not cause significant mortality to any life stage
of the copepods Microarthridion littorale, Paronychocamptus wilsoni, and Enhydrosoma
propinquum, even after 7 days of exposure. However, fenvalerate concentrations as low as 25 ppb
depressed egg production by 50 to 100% and mean clutch size by 40-100% (Chandler 1990). Thus,
although a pesticide may not have an immediate effect on a community, restriction of reproductive
capacity may lead to a decline in abundance. Chlorothalonil, an agricultural fungicide, did not appear
to significantly harm Limnephilus sp., Pisidium sp., Haliplus sp., Gammarus sp. and midges.
However, Sigara alternata did experience increased mortality (Ernst et al. 1991). Phorate, an
organophosphorous insecticide, can result in significant mortality of aquatic macroinvertebrates even
when applied at recommended rates. In the Prairie Pothole Region, macroinvertebrates that were
particularly sensitive to phorate included hemipterans, mosquitoes, flies, mayflies, water mites, and
water beetles. Less sensitive were leeches, snails, aquatic worms, and ostracods (Dieter etal. 1996).
The insecticide esfenvalerate, when applied to 12 small Alabama ponds, significantly reduced
populations of invertebrates, although rotifers seemed to be less affected (Webber etal. 1992). The
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lampricide 3-trifluoromethyl-4-nitrophenol (TFM) had a significant affect on invertebrates in a small
stream (Lieffers 1990). Fenithrothion, applied for forest insect control, reduced emergence of aquatic
insects for 6-12 weeks. Densities of most invertebrates, and especially predatory species, midges, and
some other dipterans, were reduced by as much as 50% for more than one month after treatment. The
wetland sediments came to be dominated by aquatic worms and water mites. Although in many
streams and large lakes fenithrothion has transitory affects, some residual toxicity remained in bog
wetlands during the winter and into the next year (Fairchild and Eidt 1993). Effects on invertebrates of
insecticides used in rice fields are reviewed by Roger (1995).
In a stream contaminated by various organic pollutants, many mayflies, stoneflies, and caddisflies
decreased in abundance (Hachmoller et al. 1991). The heptageniid mayflies and most stoneflies were
especially sensitive to organic pollution. A few mayfly genera, such asBaetis sp. and
Paraleptophlebia sp., were more tolerant to pollution or at least were more widespread. Of caddisflies,
Hydropsyche sp. was more common in the unpolluted section of the stream and Parapsyche sp. was
more common in polluted sections. Mussels are especially sensitive to the combined effects of
pesticides, organic compounds, and excessive nutrients (Keller 1993, Metcalfe & Chaarlton 1990).
The chronic release of a timber preservative that contained three pesticides drastically reduced
invertebrate diversity in a river in England (McNeill 1989). In a Pennsylvania stream subjected to
industrial pollution, including contamination with PCB's, isopods, oligochaetes, and craneflies were
the main survivors, compared with non-urbanized control segments (Kemp and Spotila 1996). After 25
days, an oil spill in a Missouri stream had reduced the macroinvertebrate population to less than 0.1%
of normal densities. Recovery of some species of stoneflies, mayflies, and caddisflies did not occur
until at least nine months later (Crunkilton and Duchrow 1990). In another study, the burrowing
mayfly (Hexagenia sp.) had reduced densities where sediments contained visible petroleum oil
residues (Schloesser et al. 1991b). Vinyl chloride discharges from a factory severely degraded the
macroinvertebrate population in the Niagra River watershed in Ontario (Dickman and Rygiel 1992).
Effects of poly chlorinated dibenzo-p-dioxins (PCDD's) and dibenzofurans (PCDF's) are reviewed by
Fletcher & McKay (1993). Effects of the ingredients of common detergents are reviewed by Lewis
(1991), who concluded that detergents using linear alkylbenzene sulfonate (LAS) surfactants are
perhaps among the more benign ones, in terms of aquatic biological impacts. In laboratory tests, a
surfactant was found to be approximately 100 times more toxic than the herbicide glysophate, with
which it is commonly applied (Henry et al. 1994).
A biological control agent — Bacillus thuringiensis var. israelensis (B.t.i.) ~ appears generally to have
minimal adverse effects on non-target insects in streams (Wipfli and Merritt 1994, Kreutzweiser et al.
1994a) although mortality has been observed in Lepidoptera (Jackson et al. 1994), some midges
(Merritt et al. 1989), craneflies (Wipfli and Merritt 1994, Waalwijk et al. 1992), caddisflies, and
mayflies.
5.4 Acidification
Processes
Acidification can alter community structure by (a) being acutely or chronically damaging to tissues of
invertebrates - species that easily lose sodium ions when pH is reduced tend to be most sensitive
(Steinberg and Wright 1992), (b) altering algal communities and aquatic macrophytes upon which
some invertebrates depend for food and shelter, (c) altering predation on invertebrates by decimating
numbers of other crustaceans, fish, and amphibians, (d) altering the bioavailability of some other
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potential stressors, such as heavy metals (Brett 1989, Stokes etal. 1989, Feldman and Connor 1992,
Stephensen etal. 1994). The effects of acidity also depend on the seasonal life cycles of
macroinvertebrates and water temperature (Pilgrim and Burt 1993). In areas with snow, the greatest
acid stress often occurs during snowmelt. Young larvae were more susceptible than older larvae at that
time (Gorham and Vodopich 1992). Metals and acidity also can interact to alter the toxicity of either
or both (e.g., Havens 1994a).
Effects on Richness
Acidity often decreases the richness of macroinvertebrates in aquatic habitats (Schell andKerekes
1989, Hall 1994a). Reductions in acid emissions from some Canadian smelters was followed by
significant increases in richness of invertebrates in water bodies downwind of the smelters (Griffiths
and Keller 1992). Invertebrate richness in a survey of 20 streams increased as pH increased from 4.2 to
5.7. Above pH 5.7, a reversal occurred and richness decreased, at least when aluminum levels also
decreased and humic content increased (Kullberg 1990). Montana wetlands with naturally low
acidities were found to have greater dominance by a few taxa, and greater percent Amphipoda and
Hilsenhoff Biotic Index. Wetlands with circumneutral pH had greater taxonomic richness, number of
Chironomidae taxa, and percent filterer-collectors (Stribling et al. 1995).
Effects on Species Composition
Acidification effects on aquatic invertebrate communities have been researched extensively. Much of
the information from European literature (including autecological responses) is compiled by Johnson et
al. (1993). Researchers in Wales even developed a dichotomous key based on invertebrate indicator
species to classify and rank streams according to their acidity (Wade et al. 1990). Derived mainly the
North American literature, Table 5.1 categorizes some taxa as more or less tolerant of acidification. It
should also be noted some invertebrates are sensitive to pH increases. For example, stormwater input
to a Florida freshwater marsh increased phosphorus levels, lowered oxygen levels, and raised pH and
hardness, resulting in macroinvertebrate population shifts toward species that otherwise are intolerant
of typical acidic and oligotrophic conditions in the studied wetland (Graves et al. 1998).
Table 5.1. Relative Tolerance of Invertebrate Taxa to Acidification
More Tolerant (Less Sensitive):
some Odonata, at least damselflies (Parker et al. 1992, Johnson et al. 1993, Baker and Christensen 1989)
some water beetles (Parker et al. 1992, Johnson et al. 1993) especially hydrophilid and dytiscid beetles (Baker and
Christensen 1989)
some water bugs, at least Notonectidae, Gerridae, Corixidae (Johnson et al. 1993, Baker and Christensen 1989)
some caddisflies: Cheumatopsyche pettiti (Camargo and Ward 1992).
some Megaloptera (Sialis)
some Diptera, at least phantom midges (Chaoboms) (Johnson et al. 1993), midges (Havens 1994a, Baker and Christensen
1989, Tuchman 1993), blackflies (Baker and Christensen 1989).
some stoneflies (Tuchman 1993) such as Amphinemura wALeuctra (Griffith et al. 1995)
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the mayfly Eurylophella funeralis (Griffith et al. 1995)
some water mites (Havens 1994a)
some zooplankters, such as Daphnia galeata mendotae, D. retrocurva, Skistodiaptomus oregonensis (Havens 1993) and the
rotifers Gastropus stylifer, Keratella taurocephala, Polyarthra renata, Symchaeta sp. (Fore et al. 1998)
scrapers and collectors (Smith et al. 1990); filter feeders (Pisidium casertanum) and predators (Enallagma sp.) (Mackie
1989)
Less Tolerant (More Sensitive):
some water bugs (Parker et al. 1992)
some caddisflies (Parker et al. 1992), such as Lepidostoma sp. (Hall 1994a), and some in the scraper and predator guilds
(Williams 1991)
some midges, such as Tanytarsus, Microtendipes, and Nilothauma (Griffiths 1992),Micropsectra sp. (Hall 1994a)
some blackflies, such as Simulium sp., Prosimulim sp. (Hall 1994a)
some mayflies (Balding 1992, Steiner 1993), especially Baetis, Serratella, Drunella, Epeorus, Paraleptophlebia (Melack
and Stoddard 1991, Hall 1994a)
some Odonata (Enallagma civile, Giberson and MacKay 1991)
Dugesia dorotocephala (Camargo and Ward 1992)
molluscs (Grapentine and Rosenberg 1992, Gibbons and Mackie 1991, Balding 1992), including clams (Schell and Kerekes
1989, Melack and Stoddard 1991) and mussels
snails, leeches (pH >5.0, Schell and Kerekes 1989)
the amphipod Hyalella azteca (Havens 1994a, Mackie (1989; pH must remain above 5.8, Grapentine and Rosenberg 1992)
Gammarus minus (Griffith et al. 1995)
many stoneflies, e.g., Peltoperla arcuata (Griffith et al. 1995)
shredders (Tuchman 1993) and deposit feeders (Smith et al. 1990)
Bosmina longirostris (Havens 1993)
themtifersAsplanchnapriodonta, Collotheca mutabilis, Conochiloides sp., Conochilusunicornis, Gastropus hyptopus,
Kellicota longispina, Keratella cochlearis, Keratella crassa, Polyarthra dolichoptera, Trichocera cylindrica (Fore et al.
1998)
Effects on Abundance, Density, Biomass, Productivity
With increased acidity, many aquatic invertebrates declined in numbers and biomass, especially in
wetlands with pH < 5.0 (Parker et al. 1992). Invertebrate densities in a pH 4.5 channel were
significantly lower than in channels of pH 5.9 or 7.4 (Griffiths 1992). Reductions in acid emissions
from some Canadian smelters were followed by significant increases in densities of invertebrates in
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water bodies downwind of the smelters (Griffiths and Keller 1992). In many aquatic habitats the
abundance and biomass of macroinvertebrates seems to be controlled more by nutrient availability than
by acidity (Schell andKerekes 1989).
5.5 Effects of Salinization
Salinity and specific conductance (conductivity) can markedly influence species composition,
richness, and abundance of invertebrates, particularly along coastal rivers and among non-riverine
inland wetlands. Effects of salinity on invertebrate communities are perhaps less noticeable in inland
streams and rivers, where the range of salinity is less (Williams etal. 1991). Salinization of wetlands
can occur as a result of cutting off wetlands from some types of groundwater inflow (Swanson et al.
1988), from increasing evaporative water loss, or from discharge of effluents (especially irrigation
return water), or from routing runoff of relatively high conductivity into wetlands. Among some
lacustrine wetlands in Wyoming, relative production of different invertebrate functional groups
(scrapers, deposit-feeders) varied depending on salinity and associated floating macrophytes (Chora at
low salinity, Potomogeton at higher), but total invertebrate production did not. At lower salinity,
scrapers and epiphytic deposit-feeders predominated whereas at higher salinity, filter-feeders and
benthic deposit-feeders assumed a larger proportion of the total biomass (Hart & Lovvorn 2000).
Processes
High levels of salinity can alter structure of freshwater invertebrate communities by (a) being acutely
or chronically damaging to tissues of invertebrates, (b) altering species composition and structure of
algal communities and aquatic macrophytes upon which some invertebrates depend for food and
shelter, (c) altering predation on invertebrates by decimating numbers of other crustaceans, fish, and
amphibians, (d) altering the bioavailability of some other potential stressors, such as heavy metals and
nutrients.
Effects on Species Richness
Even at low concentrations, increases in chloride (a correlate of salinity, and often associated with
road salt applications) among 27 Minnesota wetlands were significantly correlated with declines in
species richness among the wetlands (Gernes and Helgen 1999). In a survey of East African lakes,
zooplankton taxa richness began to decline at a conductivity of 1000 |j,S/cm and declined to just 2-3
rotifer taxa at salinities above 3000 |j,S/cm (Green 1993).
Effects on Species Composition
In Wyoming wetlands of fairly low salinity (0.8 - 30 mS/cm), the dominant macroinvertebrates are
amphipods and epiphytic snails. Above 30 mS/cm, they are mostly replaced by midges, as well as
predatory dragonflies and water bugs (Wollheim and Lovvorn 1995). In Utah, as salinity in the Great
Salt Lake declined from 13-23% to 6%, the brine shrimp Artemia salina and the brine fly Ephydra
cinerea declined in abundance while Ephydra hians increased in dominance (Stephens 1990). Table
5.2 summarizes some of the recent literature on invertebrate salinity preferences and tolerances,
categorizing some taxa as more or less tolerant of salinity levels atypical of the habitats they normally
inhabit. Other recent species-specific salinity data for wetland invertebrates are presented by Walker
et al. (1995), Parker and Wright (1992), and Lovvorn et al. (1999). In Minnesota, as chloride increased
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across a spatial gradient of 27 wetlands, dragonflies, mayflies, and caddisflies decreased significantly
(Gernes and Helgen 1999).
Table 5.2. Relative Tolerance of Freshwater Invertebrate Taxa to Salinity
More Tolerant (Less Sensitive):
Artemia franciscana (Wurtsbaugh and Berry 1990)
Trichocorixa reticulata, Tanypusgrodhausi (conductivity >53.5 mS/cm, Euliss et al. 1991)
Trichocorixa reticulata, Tanypus sp., Ephydra sp., Brachionusplicitilis (salinity of1,361 to >70,000 mg/L TDS, Parke
and Knight 1992)
Dolichopodidae, Ephydra hians (salinity >63 ppm, Hammered al. 1990)
Ephydra, Culicoides (salinities >10%, Short et al. 1991)Bezzia sp., Hygrotus salinarius, Cricotopus ornatus (salinity 11-63
ppm, Hammer et al. 1990).
Lymnaea elodes (at conductivity >5,000 |j,S/cm, Swanson et al. 1988)
Less Tolerant (More Sensitive):
Gammarus sp., Gyptotendipes sp., Chironomusplumosus (salinity 3-10 ppm, Hammer et al. 1990)
Branchionus sp., Cletocampus sp., Diaptomus sp., Trichocorixa verticalis (Wurtsbaugh and Berry 1990)
Lymnaea stagnalis (conductivity <5,000 |j,S/cm, Swanson et al. 1988)
Toxicity Effects
High salinity in irrigation water entering the Stillwater Wildlife Management Area in southwestern
Nevada was toxic to Hyalella azteca amphipods and Daphnia magna (Ingersoll et al. 1992). The
mussel and the quagga mussel (D. bugensis) can survive over 18 days in salt concentrations higher
than 5%0(Spidle etal. 1995).
5.6 Sedimentation/ Burial
The accelerated covering of plants and other natural substrates with inorganic particles (i.e.,
sedimentation) can result in reduced invertebrate richness and density, and alteration of species
composition (Hellawell 1986). .
Processes
Exposure to sediments affects behavior and survival of individual invertebrates (Taylor and Pascoe
1994), and consequently invertebrate communities. Because wetlands are basically deposit!onal
environments, many wetland invertebrate communities are able to tolerate occasional deposition of
small amounts of sediment, whereas constant or severe deposition causes major changes. Excessive
sedimentation affects invertebrates because it (a) buries essential detrital and algal food sources, and
excess time required to move through deposited sediment and collect scarce food items from a younger
substrate may result in lower survival, (b) reduces flow of interstial water necessary to supplying
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invertebrates with adequate dissolved oxygen, and (c) kills macrophytes that otherwise provide
attachment structures and shelter to invertebrates (Hellawell 1986, Newcombe and MacDonald 1991,
Ryan 1991). Sediments often contain metals and other contaminants at toxic levels. Approaches for
characterizing toxicological and physical risks of sediments to invertebrates are summarized by Adams
etal. (1992). Once deposited, sediments can further damage wetland invertebrate communities if they
are resuspended by wind mixing or fish, making water turbid. For example, benthic feeding carp,
Cyprinus carpio, noticeably increase water column turbidity both directly (as they move along the
bottom) and by consuming aquatic plants that otherwise would stabilize and trap sediments (Lougheed
et al. 1998). Planktonic invertebrate biomass declined in Utah ponds after introduction of carp (Huener
and Kadlec 1992).
Effects on Richness, Abundance, and Density
Prolonged (>10 years) siltation of backwater lakes in the Mississippi River (Iowa) at rates of 1.5
cm/year led to significant declines in populations of the mussel, Musculium transversum (Eckblad and
Lehtinen 1991). Deposition of 5-10 cm of sediment in willow wetlands in northeastern Missouri
resulted in lower invertebrate community richness and density, compared with non-sedimented sites
(Magee 1993). In some instances, invertebrate density and perhaps richness can increase over the long
term if sedimentation of coarser-particled substrates creates fine-particled substrates that better support
establishment of rooted plants. In temporarily flooded prairie pothole wetlands, only caddisflies
seemed relatively unaffected by surrounding land use, whereas ostracods, cladocerans, and snails
(planorbiids, lymnaeids, physids) were diminished, presumably in part due to the effects of
sedimentation (Euliss and Mushet 1999). Across a gradient of increasing land use intensity in
Minnesota, the snail Physa decreased (Gernes and Helgen 1999).
Effects on Species Composition
Burrowing, tube-forming worms and midges commonly predominate where sediments accumulate
(Magee 1993). Filter-feeding and bottom-grazing taxa are most sensitive (Evans 1996, Lougheed and
Chow-Fraser 1998). However, invertebrate size and behavior also influence tolerance (McClelland
and Brusven 1980). Taxa that characteristically occupy the water column, and especially the smaller
forms of such taxa, tend to be less sensitive to sediment deposition than benthic or epiphytic taxa, but
may still be highly sensitive to turbidity (Newcome and MacDonald 1991). Substrates newly created
by sedimentation may attract tolerant individuals and species that are poor competitors on older, more
crowded substrates (Soster and McCall 1990).
Some studies (e.g., Hogg and Norris 1991, Ludwa 1994, Lamberti and Berg 1995, Carlisle et al. 1998,
Ludwa & Richter 2000) have linked changes in invertebrate communities to development of
watersheds, and development often is accompanied by increased export of sediment to water bodies.
Some water beetles (e.g., Stenelmis crenata, Optioservusfastiditus), mayflies (e.g., Baetis tricaudatus,
Stenonema sp.), and even some stoneflies (e.g., Taeniopteryx nivalis) can increase in response to
watershed development despite the accompanying sedimentation (Lamberti and Berg 1995). However,
severe and rapid sedimentation is inevitably lethal to nearly all aquatic invertebrates. In North Dakota,
wetlands surrounded by cropland were virtually devoid of the resting eggs of zooplankton, whereas
such eggs were present extensively in wetlands surrounded by mostly natural grassland, which
presumably minimized erosion and sedimentation (Euliss and Mushet 1999).
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Unionid mussels are one group that is sensitive to increased sedimentation (Goudreau et al. 1993, Box
and Mossa 1999). Numbers of the mussel, Musculium partumeium., and amphipods were reduced in
willow wetlands in northeastern Missouri where 5-10 cm of sediment had been recently deposited
(Magee 1993). However, some bivalves seem to tolerate moderate levels of turbidity and periodic
sedimentation. For example, in a laboratory experiment where the mussels Amblemaplicataplicata
and Fusconaia ebena were exposed to total suspended solids up to 120 mg/L for five minutes every 0.5
and 3.0 hours, no consistent effects were observed (Payne and Miller 1999).
In a tidal freshwater wetland, several taxa were relatively unaffected by single depositions of large
amounts of sediment (0.3 meters of fluid mud). These included Limnodrilus sp., Limnodrilus
hoffmeisteri, Ilyodrilus templetoni, Corbicula fluminea and Coelotanypus scapularis - all species that
typify this dynamic type of wetland. Somewhat greater depths of deposition increased the mortality of
midges and small (< 10 mm) Corbicula fluminea (Diaz 1994).
5.7 Vegetation Removal
Vegetation has been shown to be a major factor shaping wetland invertebrate communities (Krieger
1992, Wissinger 1999). Indeed, wetland managers often manipulate vegetation structure — e.g., by
mowing, burning, plowing, planting — to encourage or discourage populations of desirable or
undesirable invertebrates (Batzer and Resh 1992a, Kirkman and Sharitz 1994, de Szalay et al. 1996, de
Szalay and Resh 1997, Gray et al. 1999).
Processes
Removal of aquatic or riparian vegetation affects invertebrates because it (a) removes attachment
substrates that otherwise provide additional vertical space in the water column for colonization, (b)
removes shade, thus increasing water temperature and enhancing growth of algae, (c) increases water
circulation and perhaps velocity, with accompanying increases in dissolved oxygen and possible
resuspension of sediments, (d) reduces inputs of leaf litter that provide food to some invertebrate taxa,
(e) reduces structures that otherwise shelter invertebrates from predators (Jordan et al. 1994) and
erosive forces (Roman etal 1994), and (f) reduces a source of dissolved oxygen, i.e., plants that are
capable of oxidizing sediments.
Effects on Species Richness
Wetland emergent and submerged plants often support a higher richness of taxa than open water areas
and sediments, at least during some seasons (Brady 1992, Brady and Burton 1995). Complete
removal of vegetation generally reduces richness of the wetland invertebrate community, but patchy
removal or moderate grazing sometimes increases richness (McLaughlin and Harris 1990, Gray et al.
1999), such as where rights-of-way cross forested wetlands (King et al. 2000).
Effects on Species Composition
Logically, the taxa that are most closely associated with vascular plants are the ones likely to be
adversely affected by vegetation removal. These primarily include algae-feeding species (grazers) and
their predators, although removal of part of a forest canopy can result in increased dominance of algae-
feeding species (King et al. 2000). Also, over the long term, removal of vegetation from wetlands with
little input of leaf litter from uplands can decimate detritivorous taxa.
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In a Washington pond, Helobdella leeches, Asellus isopods, mayflies, and some dragonflies (especially
the large-bodied Anax) were more commonly associated with emergent vegetation than with
submerged vegetation or open water areas. Midges, freshwater shrimp (Hyalella azteca), and molluscs
(especially Lymnaea sp., Gyraulus sp., andAnodonta sp.) were more common on the submerged plants
(Parsons and Matthews 1995). A decrease in early season flooded plant cover can reduce mosquito
larvae (Wood et al. 1992) and confine remaining larvae populations to the perimeter of the marshes
(Batzer and Resh (1992a). Such a reduction in plant cover also can increase the density of midges and
dytiscid beetle larvae.
The type of vegetation along a wetland can influence the species composition within the wetland. A
laboratory experiment demonstrated that leaves of an introduced shrub (Multiflora rose) were less
nutritious to Leptophlebia cupida than leaves from some kinds of native vegetation, but either
increased or had no effect on the stonefly Soyedina carolinensi ('Sweeny 1993).
Effects on Abundance, Density, Biomass
Vegetation has been demonstrated to support a greater abundance and/or biomass of invertebrates than
open water or bottom habitat in a wide range of wetlands, for example: Florida Everglades,
Kissimmee River (Florida) wetlands, flatwoods marshes, cypress domes, Southeastern and
Pennsylvania beaver pond wetlands, New England riparian sedge meadows, northern prairie marshes,
temporarily flooded wetlands in Missouri, seasonal and semipermanent California wetlands, tidal
freshwater wetlands, Gulf Coast bayous, and Great Lakes coastal wetlands (Batzer et al. 1999).
The larger growths of epiphytic algae on submerged vegetation may partly explain this greater
abundance of macroinvertebrates. In particular, submerged aquatic vegetation with finely divided or
thin, submerged leaves and large surface area per unit support higher densities and biomass of
macroinvertebrates than emergent vegetation (Parsons and Matthews 1995, Olson et al. 1995). In a
Wisconsin lake, densities ofAmnicola limnosa and Gyraulusparvus were 162 and 48 times greater in
Ceratophyllum demersum beds than in vegetation-free areas (Beckett et al. 1992). Areas of the
Chesapeake Bay containing the introduced submersed plant, Hydrilla verticillata, had greater densities
of invertebrates (Posey etal. 1993).
However, densities of invertebrates are not always greater in denser stands of wetland vegetation. In a
created Florida marsh, dipteran abundance was greater in unvegetated areas (Streever et al. 1995).
Dense stands of vegetation can harbor some predatory fish, and especially in the case of floating-
leaved plants, can be associated with reduction in dissolved oxygen. Thus, consideration should be
given when monitoring mitigation wetlands to include not only measures of plant stem density, but
measures of invertebrate community composition (e.g., Sewell andHiggins 1991, Garono andKooser
1994).
5.8 Thermal Alteration
Processes
Although lethally hot temperatures are seldom encountered by entire invertebrate communities, the
growth and emergence of many wetland insects is closely tied to temperature. In northern climates,
freezing can affect invertebrate communities both directly and indirectly. Some invertebrates avoid
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being frozen by migrating to microhabitats that do not freeze completely, but which often have low
concentrations of dissolved oxygen (Euliss et al. 1999). Others are physiologically tolerant of freezing,
at least for short periods (Block 1991). Ice also can cause lasting physical and chemical alteration of
wetlands. Less dramatic temperature differences between low-elevation wetlands and higher-elevation
wetlands can influence invertebrate richness and species composition, so should be factored into
monitoring plans (Land & Reymond 1993).
When development encroaches on wetlands, excavations and fills often change the proportion of
groundwater vs. surface water inputs to wetlands, and the amount and timing of these inputs. Such
changes can alter wetland thermal regimes, and consequently invertebrates (Williams 1991), because
each type of water source has characteristic thermal properties - groundwater usually being cooler in
summer and warmer in winter, compared with surface runoff. Also, removal of shading vegetation
(Section 5.7) affects wetland thermal regimes, as does the direct discharge of heated effluents.
Effects on Richness, Abundance, Density, Growth
Warmer springs tend to have fewer aquatic insect taxa than springs closer to ambient air temperature
(Pritchard 1991). Heated effluents also generally decrease the abundance and diversity of
macroinvertebrates (Robinson and Craven 1993, Payne 1991). However, in a pair of experimental
channels in Ontario, heated vs. unheated channels, after two years, differed little with regard to total
invertebrate density, although densities of dipterans were slightly less in the heated channel (Hogg et
al. 1992). Growth rates of midges (Chironomini, Tanytarsini, Orthocladiinae) in a Georgia blackwater
swamp reached a maximum at 21-24°C (Hauer and Benke 1991). Birth rates of the amphipod Hyalella
azteca, declined when summer temperatures in a Florida lake reached 25-34°C (Edwards and Cowell
1992).
Effects on Species Composition
By helping species at the northern limits of their range to successfully overwinter, warmer winter
temperatures can facilitate the northerly range expansion of such peripheral species. For example, the
heated discharge of a power plant was critical for the overwinter survival of the Asiatic clam,
Corbiculafluminea (French and Schloesser 1991). Some species (e.g., the stonefly, Nemoura
trispinosa, and the caddisfly, Lepidostoma vernale) may have mechanisms that enable them to
compensate for small changes in thermal regimes, whereas other groups (e.g., Orthocladinae and some
other midges) may be more sensitive (Hogg et al. 1992).
5.9 Dehydration/ Inundation
Processes
Some of the most dramatic changes to wetland invertebrate communities occur when (a) pools or
channels - even seasonal ones — are introduced into wetlands that seldom or never contained surface
water, or (b) wetlands that seldom or never went completely dry are subjected to drought or complete
drawdowns. Usually less dramatic are changes to invertebrate communities that occur when slight
changes occur in the timing, duration, predictability, and depth of surface water (Eyre 1992, Heicher
1993, Giberson et al. 1992). Permanent water in a wetland can act as a refuge to many species during
drought, but also as a disturbance that limits plant growth and development of some invertebrates
(Golladay etal. 1997). The degree to which desiccation alters invertebrate communities is partly
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influenced by the severity of the local climate (e.g., maximum and minimum temperature of the
wetland sediments during dried-out phase), the ability of plant root systems and detritus to provide
moist refuges, and the speed with which water levels fall or rise (Riley & Bookhout 1990).
Wetland water regimes — including subsurface water table levels — influence wetland invertebrates by:
(a) altering the amount and pattern of horizontal and vertical habitat space available for
colonization;
(b) changing the types of algae and vascular plants that occur, the proportions of these two
major energy sources, and the seasons in which they occur (Murkin et al. 199la);
(c) changing the extent of contact between plants and water, thus influencing attachment space,
availability of detrital foods (Ross and Murkin 1993, De Szalay etal. 1999), shade, and shelter;
(d) altering physical (e.g., temperature, turbidity, oxygen) and chemical regimes (organic
carbon, nutrients, metals);
(e) influencing access of predators (Reice 1991, Martin et al. 1991, Mallory et al. 1994,
Johnson etal. 1995, Wellborn et al. 1996) as well as the intensity of competition (Wissinger et
al. 1999);
(f) in the case of complete desiccation and freezing, being directly lethal to many species
(Layzer etal. 1993).
Adaptations of major invertebrate groups to desiccation are shown in Table 5.3. Some wetland
invertebrates can respond to changing water regime by moving within or among wetlands (Jeffries
1994). An amphipod, Corophium spinicorne, and a snail, Jugaplicifera moved to deeper water within
a backdune lake that experienced fluctuating water levels (as much as 2.5 m) in the Oregon Dunes
National Monument (Wones and Larson 1991). In a West Virginia stream, the mayfly as
Paraleptophlebia sp. burrowed toward the water table in response to receding surface water levels
(Griffith and Perry 1993). Likewise, in a New York stream the stonefly Paracapnia opis survived by
following the water table into the hyporheic zone (Delucchi and Peckarsky 1989). In intermittent
streams of the Sonoran Desert, some insects escape desiccation by moving up- or downstream (Stanley
et al. 1994). Also, invertebrates in tidal freshwater wetlands have evolved strategies for dealing with
the threat of desiccation (from falling tides) on a daily basis, e.g., Yozzo and Diaz (1999). Taxa that
typify many tidal freshwater wetlands include midges and other worm-like taxa, e.g., Tubificidae,
Naididae, Enchytraeidae.
Other wetland invertebrates can respond to water level drawdown or drought by laying drought-
resistant eggs or by burrowing down to the water table and aestivating. For example, the dessication
tolerance of eggs of three mosquitoes — Aedes vexans, Aedes trivittatus and Psorophora sp. - allowed
them to colonize temporary woodland pools in Michigan (Higgins and Merrit 1999). Resting eggs
(also known as epiphia) of some species can survive in complete drawdown conditions for a year or
longer. For example, after a complete drawdown that lasted 247 - 346 days, the copepods Diacyclops
haueri, Diacyclops crassicaudis brachycercus and Acanthocyclops vernalis emerged from resting eggs
within weeks of the return of flooding (Wyngaard et al. 1991). After two years the resting eggs of the
cladoceran Diaptomus stagnalis were able to hatch in laboratory conditions of remoistening (Taylor et
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al. 1990). The stonefly Amphinemura delosa and the mayfly Ameletus ludens survived a complete
drawdown in New York streams by having desiccation-resistant eggs (Delucchi and Peckarsky 1989).
Indeed, the density and viability of dormant stages of some invertebrates might be used to determine in
advance whether (and how rapidly) the restoration of a drained wetland will restore its functional
characteristics (Euliss and Mushet 1999). If sediment samples from a drained or farmed wetland are
incubated for about 8 weeks in a moistened laboratory setting, yet fail to produce such hatchings, it
might be assumed that degradation has been so severe as to make full functional restoration of the
former wetland impractical.
Table 5.3. Adaptations of macroinvertebrates to drawdown or drought (from Smock 1999).
Taxon
Amphipoda
Bivalvia
Caecidotea sp.
Ceratopogonidae
Midges
Coleoptera
Crangonyx sp.
Diaptomus stagnalis
Gastropoda
Isopoda
Ostracoda
Paratendipes sp.
Polypedilum sp.
Sphaeriidae
Tabanidae
Adaptation
Burrowing, Aestivation
Burrowing, Aestivation
Burrowing, Aestivation
Flight
Flight
Flight, Aestivation
Burrowing, Aestivation
Resting stage
Burrowing, Aestivation
Burrowing, Aestivation
Resting stage
Flight
Flight
Burrowing, Aestivation
Flight
Effects on Species Richness
Wetlands with a wide variety of hydrologic zones often have a wide variety of vegetation types, and
this in turn can support a wider variety of invertebrates (Kirkman and Sharitz 1994, Williams et al.
1996). Drought or complete drawdown reduces invertebrate richness in many wetlands, and the effect
may be temporary or long-term, i.e., still noticeable after wetter conditions return. In particular, the
richness of midges diminishes in wetlands following drought (Hershey et al. 1999).
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When water is restored to drained former wetlands, invertebrate richness increases during the first few
years following restoration (Nilsson and Danell 1981, Hemesath 1991). Richness also tends to
increase when water levels are raised in existing wetlands. Similarly, richness (at least of midges)
tends to be greater in wetlands having longer durations of standing water during the growing season,
whether isolated (Nelson and Butler 1987) or part of a floodplain (Smock 1999). This is partly
because wetlands with longer hydroperiods generally are deeper, larger, and more likely to contain
submersed and floating-leaved plants that diversify the range of habitats available. Also, wetlands
with longer durations of flooding are less likely to experience deep freezing of sediments and types of
human activities (e.g., soil compaction, cultivation) that sometimes reduce habitat quality for
invertebrates (Swanson et al. 1974). In Wisconsin, only 4 taxa were found in short duration ponds
whereas 65 were found in long duration ponds; richness of diving beetles and other predators
accounted for much of the increase (Schneider and Frost 1996, Schneider 1999). However, short
duration ponds sometimes support more rare species than do longer duration ponds (Collinson et al.
1995). In temporary pools of Colorado, macroinvertebrate richness increased with increasing number
of days that seasonal wetlands contained standing water according to the following equation (Wisinger
etal. 1999):
y = 2.52e('0232x)
Where x = number of days with standing water; y = number of species present in the wetland
As water permanence increased among wetlands in Grand Teton and Yellowstone parks, food webs
became more complex and taxa numbers increased (Duffy 1999). Temporary subalpine wetlands
contained 10 species, with the trophic levels culminating in the predaceous beetles Liodessus sp. and
Oreodytes sp.. Seasonal wetlands contained 13 species with the top predator being the longer-lived
dragonfly Lestes unguicalatus. Semipermanent wetlands supported 14 taxa. Use of emergence traps in
19 Seattle-area wetlands also yielded more taxa from permanently flooded than seasonally flooded
wetlands (Ludwa & Richter 2000).
However, after an initial year of continual flooding the invertebrate richness in some inundated
wetlands declines. For example, flooding of Manitoba marshes at first increased the variety of both
nektonic and benthic invertebrates in vegetation, although not in open water (Murkin et al. 1991,
1992). A short time later, richness of benthic taxa declined (Murkin and Kadlec 1986b). This is
perhaps typical. Especially when inundation persists for years with little fluctuation in water level,
sediments often become anoxic and light deficits caused by algal blooms can reduce the amount and
variety of aquatic plants available as invertebrate habitats, thus reducing invertebrate richness (Neckles
etal. 1990).
Effects on Species Composition
A searchable species database on hydroperiod relationships of North American wetland taxa has
recently been compiled and is accessible via the internet (Adamus and Gonyaw 2000), so discussion of
that topic here will be limited.
Species composition can indicate how long and in what seasons a wetland has contained surface water.
This requires that each species found at a wetland first be classified as to its hydrological requirements
— a relatively simple procedure using life history categories such as defined by Hartland-Rowe (1966);
McLachlan (1970, 1975, 1985); Wiggins et al. (1980); Jeffries (1989); Eyre et al. (1991); and Batzer
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and Wissinger (1996). The usefulness of species composition for inferring hydrologic conditions, at
least of prairie wetlands, has been demonstrated with midges (Euliss et al. 1993), water beetles
(Hanson and Swanson 1989), and macroinvertebrates generally (Neckles et al. 1990, Bataille and
Baldassarre 1993).
In general, wetlands can cautiously be deduced to be of greater hydrologic permanence when they
contain a higher density and richness of longer-lived and/or relatively immobile species (e.g., snails,
mollusks, amphipods, worms, leeches, crayfish), as compared with short-lived species (e.g.,
anostracans, conchostracans), species that survive the winter as drought-resistant eggs (e.g., Daphnid),
and/or species that are relatively mobile (e.g., midges, some water beetles and bugs). Drought and
drawdown renders the less mobile species more vulnerable to predation, as well as causing their direct
loss due to desiccation and related factors (e.g., Stanley et al. 1994). Drought also seems to provide
competitive advantage to many non-insect invertebrates, and during drought that component can
increase at the expense of the insect component of the invertebrate community (Hershey et al. 1999).
In a particular seasonal wetland, a dominance of taxa that reputedly survive drawdown by resisting
desiccation (e.g., with resistant eggs) rather than by flying away, might suggest that the dry-season
microclimate of that seasonal wetland is less harsh than that of wetlands where aerial dispersers
dominate (Wissinger 1999). Dispersal characteristics of many taxa have been cataloged in the "Pond-
internet database of invertebrate life history characteristics:
http ://www. ent3 .orst. edu/PondFX/pondlife_main.htm
Among prairie pothole wetlands, a recent shift to greater hydrologic permanence is suggested by a shift
from herbivorous to detrivorous species of macroinvertebrates, and in a shift from open-water forms
(e.g., zooplankton, water striders) to forms that characteristically dwell in vegetation (e.g., some
mayflies) (Murkin and Kadlec 1986, Murkin et al. 1991). In particular, densities of non-predatory
midges (Chironomidae) increase greatly during the first year after flooding, and within this family,
species characterized by the greatest tolerance for low oxygen levels increase the most (Murkin and
Kadlec 1986b). Densities of swimming (nektonic) and bottom-dwelling (benthic) predatory
invertebrates do not increase with flooding as much as do numbers of nektonic and benthic herbivores
and detritivores. Predatory species can even decrease after flooding (Murkin et al. 1991), and they
often increase as drought or drawdown progresses.
In Minnesota marshes during dry years, molluscs, rotifers and cladocerans were more abundant than
in wet years. In wet years, midges, ceratopogonids, copepods, and ostracods were more abundant than
in dry years. Stratiomyid flies, water beetles, and craneflies showed no obvious relationship to prior
drought (Hershey etal. 1999). Flooding of a Manitoba marsh to 1 meter above normal levels led to an
increase in the Trichopteran Agraylea multipuntata, most likely due to increased algal production and a
large increase in decaying macrophyte tissue (Ross and Murkin 1993). In a temporary wetland in
California, populations of Chironomus stigmaterus expanded during wet years due to increased
availability of detritus. Predatory beetles then responded to these larger prey populations (De Szalay et
al. 1999).
In a year-long comparison of three Manitoba pothole wetlands with different hydroperiods, the
permanent wetland was dominated by cladocerans, the semipermanent wetland by ostracods, and the
seasonal wetland by copepods (Bataille and Baldassarre 1993). Considering just the emerging aquatic
insect component, the permanent wetland was dominated by midges; the semipermanent wetland by
water beetles (early season) and midges and other fly species (mid- and late-season); and the seasonal
wetland by midges (mid-season) and other fly species (late season). Among temporary pools in
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Missouri floodplains, the pools with extended flooding had greater densities of the water beetle
Berosus; the midges Dicrotendipes, Endochironomus, Polypedilum, and several Orthocladinae; and
more copepods and oligochaetes, whereas less permanent pools had more of the Chironomini midges
and the midge Chironomus, the dipteran Palpomyi, and ostracods (Corti et al. 1997). In Everglades
sloughs, the crayfish Procambarus alleni was found mostly in sloughs with short hydroperiod
(approximately 168 days) whereas P.fallax was found in sloughs with longer hydroperiods (Hendrix
and Loftus 2000). Mosquitoes generally are more abundant in temporarily inundated pools, as well as
in permanent pools that remain shallower than about 60 cm during the growing season (Batzer and
Resh 1992b, Neckles et al. 1990). In the Seattle area, richness of emerging mayflies, and insect
richness generally, was greater in perennially flooded wetlands than in seasonal wetlands (Ludwa
1994, Ludwa & Richter 2000). Lists of species that have been found in various types of seasonally or
temporarily inundated wetlands throughout North America are provided by Batzer et al. (1999).
Among wetlands that are not permanently flooded, the duration and timing of seasonal flooding
strongly influences species composition (Schneider & Frost 1996, Wissinger 1999). In a Manitoba
marsh, when water persisted only for a few months during the summer (as opposed to the entire
growing season), densities were greater of cladocerans, midges and ostracods, whereas diving beetles,
corixids, ceratopogonids and ephydrid flies were present in about equal densities regardless of
hydroperiod (Neckles et al. 1990). In California, marshes flooded in early September had higher
winter populations ofEogammarus confervicolus and Berosus ingeminatus than those flooded in late
October. Flooding in September and maintaining water at 40 cm depth, or flooding in October and
maintaining water depth of 20 cm, favored the midge Chironomus stigmaterus (Batzer et al. 1993).
This midge became even more abundant if wetlands were flooded in August (Batzer et al. 1997).
Situations sometimes occur where wetland water regime has little affect on species composition (e.g.,
Neckles et al. 1990). This seems to be the case when the invertebrate fauna is dominated by taxa that
characteristically overwinter as adults or larvae, e.g., species of Dytiscidae, Corixidae,
Ceratopogonidae, Ephydridae, and some Chironomidae. For example, a survey of five isolated,
temporary pools in New York found the "immobile" clams Pisidium casertanum, Sphaerium
occidentale, and Caecidotea racovitzai in three of the pools (Batzer and Sion 1999). Caution also is
required in interpreting species composition data because some species with supposedly minimal
dispersal abilities are frequently carried passively into reflooded areas by mobile waterbirds (Swanson
1984).
Effects on Abundance, Density, Biomass
Flooding generally increases invertebrate densities as well as richness in wetlands, but perhaps only for
about a year after initiation of flooding. For example, flooding of Manitoba marshes containing cat-
tail, hardstem bulrush, and common reed to a level 1 m above normal caused a major year-long
increase in numbers of nektonic invertebrates in both vegetated and open water areas. Densities of
benthic invertebrates increased in flooded vegetation but not in open areas. Biomass of nektonic
invertebrates increased only in the vegetated areas (Murkin et al. 1991). On a year-round basis,
invertebrate biomass and production in prairie pothole wetlands is probably greatest in semipermanent
wetlands (Duffy andBirkelo 1993; Nelson 1989, 1993; Bataille and Baldassarre 1993), but sometimes
can reach greater seasonal peaks in temporary and permanent wetlands. In Everglades sloughs,
macroinvertebrate densities were three to five times greater in long-duration flooded sloughs (which
maintained a water level above 20 cm) compared to short hydroperiod sloughs (levels dropped below
10 cm five times during seven years)(Loftus et al 1990). In Virginia, a floodplain pond that was
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flooded for 9 months annually had higher absolute abundance, biomass, and production of
invertebrates than one that was flooded for much shorter periods (Gladden and Smock 1990). Deeper
ponds (>60 cm depth) in California and the Great Plains have greater macroinvertebrate density and
lower mosquito abundance than shallower ponds (Batzer and Resh 1992b, Neckles et al. 1990). Also,
annual reproductive effort can be affected by shortened or lengthened periods of standing water.
Populations of the mussel Musculium partumeium in permanently flooded Minnesota ponds were able
to produce two generations per year, whereas ponds that had seasonal (autumn) or complete drawdown
produced only one generation per year (Hornbach et al. 1991).
Increasing the duration or depth of standing water does not always increase the density of
invertebrates. In Florida Everglades cypress pools, total density of invertebrates may be similar for
pools with permanent (perennial) water and pools that dry up for at least a month annually (Leslie et
al. 1999). In temporary pools in Missouri, invertebrate density is inversely related to water depth,
most likely due to interactions with season and reduced algal abundance (Magee et al. 1999). In those
temporary pools, invertebrate abundance peaked in May and June when water depth was 10-20 cm,
and was lowest in the fall when depth was at 60 cm.
Aside from duration and depth of inundation, wetland invertebrate densities can be decimated by rapid
water level fluctuations, especially when those are more frequent and severe than historically
encountered in the wetland. For example, Missouri floodplain pools that experience water level
fluctuations at extreme frequency and amplitude tend to have lower invertebrate density (Magee et al.
1993). Repeated exposure to desiccation in a short period of time can lead to a marked reduction in
invertebrate density. In an Arizona stream that experienced twelve flash floods between August and
December of a single year, densities of all invertebrates were reduced from 75 to 100% (Boulton et al.
1992). In particular, water spiders, midges, Probezzia, Helicopsychidae, Leptohyphes, andPhysella
numbers declined whereas oligochaete populations appeared to be unaffected, perhaps due to rapid
colonization from upstream source populations.
In contrast, some taxa appear quite resilient to periodic spates. In a British Columbia river, populations
of the mayflies Rhithrogena and Baetis, as well as the caddisfly Hydropsyche, survived flows that
increased rapidly during flooding from 500 m3/sec to 6500 m3/sec (Rempel et al. 1999). Survival was
assisted by wetlands along the stream that provided temporary refuge from the severe currents, and
were a probable source of colonizing individuals immediately after the flood subsided. In an
Oklahoma intermittent stream where spring and fall floods reduced invertebrate densities 90%, the
mayflies Caenis sp., Leptophlebia sp. andBaetis sp. were especially resilient and midges were less so
(Miller and Golladay 1996). When such catastrophic events happen, the apparent "survivors" often are
not actual survivors but species with great dispersal capabilities that are able to colonize immediately
afterwards from habitats that were less affected by the catastrophe. The pool of available colonizers
and the speed at which they recolonize a disturbed area depends partly on the season when severe
flooding or drought occurs (Grimm and Fisher 1989). Maintaining minimum water levels (and in
streams, flow rates) can increase invertebrate densities, at least during the short term and in the part of
a wetland that is not permanently inundated (Weisberg et al. 1990, Janicki et al. 1990, Troelstrup and
Hergenrader 1990).
5.10 Effects of Other Human Influences
Invertebrates such as molluscs that complete their entire life cycle in a single wetland are especially
vulnerable to disturbances to (and pollution of) their home wetland, because of their limited ability to
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escape. In contrast, the dominance of dipterans and other mobile taxa in disturbed wetlands may be
explained partly by the short generation time of many of these taxa, and their ability to disperse
widely. However, even species that can disperse aerially may suffer the effects of a regional wetland
resource becoming fragmented (i.e., increased distances among suitable wetlands due to wetland loss
or degradation). Coincident with watershed development, these characteristic dispersers increasingly
encounter factors (e.g., vehicles, pesticides, scarcity of protective cover) during their interwetland
movements that increase the risks of interwetland movement, and consequently may suffer reduced
survival. At 25 montane wetlands in Switzerland, wetland area and proximity to other wetlands
positively influenced the number of specialist butterflies (Wettstein and Schmid 1999). In the heavily
agricultural Red River Valley of North Dakota, light-trapping indicated that both the richness of
mayflies, and the abundance of the Caenidae mayflies and the Hydropsy chid caddisflies, were
significantly related to the extent of riverine wetlands (Anderson and Vrondacek 1999). Also
influenced by the extent of riverine wetlands were Scarabaeidae beetles, heptageniid mayflies,
hydroptilid caddisflies, and ceratopogonid dipterans. Seasonal depressional wetlands were
significantly associated with abundance of the water bug Callicorixa and ichneumonid wasps.
Temporarily inundated depressional wetlands were significantly associated with mayfly richness, as
well as abundance of craneflies and hydroptilid caddisflies (Anderson and Vrondacek 1999). These
relationships were true during each of two consecutive years.
Habitat fragmentation at regional and local scales can also involve an undesirable homogenization of
wetland hydrologic variability. As stated by Wissinger and Gallagher (1999):
The loss of permanent [water] habitats from a complex [of wetlands] should reduce the pool of cyclic
colonizers that seasonally invade temporary habitats, thus reducing diversity in those temporary
habitats. Conversely, the loss of temporary habitats should reduce diversity in permanent habitats if
such habitats serve as seasonal refugia for species that otherwise would be driven extinct by predators.
The loss of temporary habitats should also reduce the rate and sequence of recolonization after drought
or other disturbances because they serve as a local source of disturbance-adapted, opportunistic species.
For cyclic colonizers, permanent and temporary habitats should act alternatively in different seasons as
sources and sinks, depending on whether a species is more vulnerable to drought or to vertebrate
predation. The presence of both habitats may be critical for the long-term viability of populations...
The ability of invertebrate habitat in constructed or restored wetlands to compensate for loss of
invertebrate habitat in unaltered wetlands, and thus reduce the hazards to dispersers that otherwise
would be exacerbated, remains uncertain. Much depends on project design (especially soil treatments
and hydrology) and location or the compensatory wetland to other wetlands. Data from 10 natural and
10 constructed marshes in Florida indicated no significant difference in densities of 20 major dipteran
taxa (Streever et al. 1996, Evans et al. 1999). A comparison of some recently restored and natural
wetlands in New York found that invertebrate communities in the restored wetlands differed initially
from natural control sites, but after 3 years were mostly similar (Brown 1995, Brown et al. 1997).
Another study, in the upper Midwest (Beaver et al. 1999), compared constructed wetlands with
temporary wetlands and both altered and unaltered wetlands with more permanent hydroperiods.
Although species composition varied among the types, the differences in species richness were not
statistically significant. Rotifers accounted for 79% of total zooplankton abundance within the
constructed wetlands and were much less dominant in the non-impacted and temporary wetlands. In
contrast, other zooplankton (cladocerans and copepods) had low densities in the constructed and
impacted wetlands and wer more abundant in the non-impacted and temporary wetlands.
The success of using wetland restoration and creation to mitigate the loss of invertebrate functions
depends very importantly on proximity of the compensatory wetland to the altered wetland. This
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needs thoughtful consideration in wetland banking programs because often these programs, for
political and economic reasons, maintain broad flexibility in choosing locations for mitigation.
Also, on a number of occasions humans have wittingly or unwittingly introduced non-native fish and
invertebrates to wetlands. Native invertebrate communities seem ill-adapted to compete with or avoid
these alien species, but data on community-wide, long term effects are mostly lacking. Studies of the
northern Everglades have discovered some macroinvertebrate colonists from Central and South
America (Rader 1994). Zebra mussel (Dreissenapolymorpha) has invaded many aquatic systems
throughout North America (d'ltri 1997). This species can totally carpet substrates, displacing native
mussels (Tucker 1993, 1994), some midges (Chironomini and Tanypodinae), snails Polycentropus sp.,
Physella sp., Pleruocera sp., and the caddisfly Oecetis sp., while having minimal or positive effect on
amphipods and flatworms (Wisenden and Bailey 1995). They may also concentrate contaminants,
making them more available to invertebrate food chains (Bruner et al. 1994). Ironically, the rapid
spread of zebra mussels may have been partly attributable to the preceding decline of native mussels as
a result of pollution and habitat alteration (Roberts 1990, Nalepa and Schloesser 1991b, 1993, Hebert
et al. 1991, Mackie 1991, Haagetal. 1993, Whittier et al. 1995). Because unionid mussels in rivers
are relatively immobile and have long life spans (often over 10 years), they are particularly susceptible
to disruptions from introduced mussels as well as from impoundments and channelization (Mehlhop
and Vaughn 1994). Riverine wetlands with higher alkalinity tend to be more susceptible to invasions
by zebra mussels (Whittier et al. 1995), although wetlands generally tend to be less suitable than other
habitats for zebra mussels (Griffiths et al. 1991). Indeed, wetlands along rivers might serve as refuges
for native mussels otherwise impacted by zebra mussel expansion (Tucker and Atwood 1995).
In boreal regions, wetlands that contain fish (even native fish) had fewer macroinvertebrates than
wetlands without fish (Mallory et al. 1994). Stocking or accidental release offish into wetlands
unaccustomed to harboring fish can have a major impact on the invertebrate communities (Johnson et
al. 1995, Martin et al. 1991).
5.11 Wetland Monitoring
Spatial and Temporal Variability
Choice of appropriate sample sizes depends on measured variation in the target taxa and metrics. Such
coefficients of variation were calculated from previous invertebrate studies in prairie pothole wetlands
(Adamus 1996), and are summarized from various aquatic studies elsewhere at:
http://www.im.nbs.gov/powcase/powvariation.html
One major source of sampling variation is the variation among habitats within a wetland. Information
on within-wetland invertebrate variability is presented (at least qualitatively) in the book edited by
Batzer et al. (1999) and is available (at a minimum) for: Florida Everglades (Rader 1994, 1999),
Florida flatwoods wetlands, southern forested floodplain wetlands, forested limesink wetlands of
Georgia (Golladay et al. 1997), bay wetlands of the Carolinas, beaver ponds in the Southeast, beaver
ponds and constructed marshes in Pennsylvania, Canadian peatlands, depress!onal wetlands of Ohio,
Lake Michigan wetlands (King andBrazner 1999), prairie potholes (Euliss andMushet 1999),
lacustrine wetlands (Murkin et al. 1991) of Manitoba, ricelands and seasonal and semipermanent
wetlands of California (de Szallay and Resh 1997), and created wetlands (Cooper and Anderson 1996)
and High Plains wetlands of Wyoming. Some other significant sources of variation include geographic
region, season, and daily weather conditions (Anderson and Vondracek 1999).
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Techniques and Equipment
The growing number of sampler types used for sampling invertebrates are described in Adamus and
Brandt 1990, Cuffney et al. 1993a,b, Bookhout 1994, Adamus 1996, Merritt and Cummins 1996, and
other sources. They include, for example:
• artificial substrates (Magee etal. 1993, Benoit et al. 1998)
• sweep nets, to collect planktonic and epiphytic invertebrates (Cheal et al. 1993, Batzer et al. 1993)
• throw traps (Turner & Trexler 1997)
• light traps (Anderson and Vondracek 1999)
• bilge-pump samplers (Batzer et al. 1993)
• stovepipe samplers (Schwartz et al. 1994)
• activity traps (Murkin et al. 1991)
• Ekman dredge (Blomqvist 1990)
• coring devices (Leslie et al. 1997)
All of these have drawbacks that make some taxa less likely to be captured than others, so a
combination of methods should be used whenever possible (Brinkman & Duffy 1996, Turner &
Trexler 1997). If a specific type of organism is to be targeted (benthic, planktonic, epiphytic, etc.),
sampling should be designed to exclude influences from other community types. As an example,
Batzer et al. (1993) sampled both benthos and the water column of a marsh. Benthos was sampled
using a bilge pump sampler in which the opening of the sampler is planted firmly on the substrate,
which was drawn through the device into a sieve. Water column invertebrates were sampled using a 1-
mm mesh sweep-net kept 5 cm above the benthos.
Among the more novel (for wetland studies) approaches that have been used in the last decade to
sample invertebrates are sticky traps (Nordstrom and Ryan 1996, King andBrazner 1999); resting egg
counts (Euliss and Mushet 1999); and the various soil extraction and other methods used to sample
earthworms and nematodes in soils and sediments of seasonally dry wetlands (Lenz & Eisenbeis 1998,
New 1998) or sub-arctic wetlands (McElligott & Lewis 1994). Special techniques are often required to
accurately survey freshwater mussels (e.g., Miller 1991).
Identification of Taxa
Some of the more recent and commonly used keys for North American aquatic invertebrates include
Merritt and Cummins (1996), Thorp and Covich (1991), Pennak (1989), andPeckarsky etal. (1990).
Also, some recent keys address specific groups, e.g., damselflies (Westfall and Minter 1996) and
caddisflies (Wiggins 1996). Results are mixed regarding the necessity of identifying invertebrate taxa
to species. While generally desirable, this decision clearly depends on the available resources and the
objective of the study. Tests of alternative multimetric indices in the Seattle area showed that indices
based on identifications of emerging insects only to the family level failed to distinguish an urban land
use gradient among wetlands, whereas indices based on genus/species levels were much more sensitive
(Ludwa 1994, Ludwa & Richter 2000).
Metrics and Indices for Assessing Impacts to Wetland Invertebrate Communities
Multiple metrics (including indicators, indexes) have been used for assessing the condition of surface
waters and are described in Merritt and Cummins (1996) and Karr and Chu (1999). Following are
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some examples of recent studies that have examined wetlands that span a regional gradient of
anthropogenic (generally land use) disturbance. Most of these studies used a multimetric approach to
assessing wetland condition.
Montana (Apfelbeck 1998):
Macroinvertebrates were sampled in 80 wetlands statewide. The wetlands belonged to ten classes that
were defined by salinity, morphology, hydroperiod, and landscape position. The multimetric approach
did not work well for wetlands that lacked open water (partly because of the sampling methods used),
or for excessively saline or alkaline wetlands. The approach succeeded in indicating the relative level
of impairment (of a wetland's invertebrate communities) but was less useful in diagnosing a cause of
the impairment. Impairments were likely related to acidification, sedimentation, excessive
enrichment, contamination with heavy metals, and hydroperiod alteration. The following metrics were
judged to be too variable to distinguish water quality impacts in the study wetlands: % collector-
gatherers, % filterer-collectors, and ratio of Tanytarsini to total Chironomidae. Also, the metric, "%
amphipods" was not used because in Montana it was correlated naturally with pH (alkalinity).
The final "core" metrics that were used and combined into the multimetric index were:
number of taxa
percent dominance (cumulative total of %'s of 1, 2, and 5 most dominant taxa)
number of taxa that are stoneflies, mayflies, dragonflies, or caddisflies
total number of individuals per sample
number of taxa that are Crustacea or Mollusca
number of taxa that are leeches, sponges, or clams
mean tolerance values of component species
The range of multimetric values, and of the individual metrics, found in each wetland class is presented
in the report.
Washington (Ludwa 1994, Ludwa & Richter 2000):
Aquatic insects were sampled in 19 Seattle-area wetlands during 3 nonconsecutive years. The
wetlands were exposed to stormwater and urban runoff to varying degrees. Insects were sampled
solely with emergence traps, mostly checked on a monthly basis. The metrics found to be most useful
for distinguishing developed from undeveloped watersheds were:
Taxa richness
Richness of Ephemeroptera + Plecoptera + Odonata + Trichoptera (EPOT) taxa
Richness of Tanytarsini taxa
Richness of Chironomini taxa
Pdchness of Tanypodini taxa
Percent of individuals as EPOT
Percent of individuals as Tanytarsini
Percent of individuals as Chironomini
Percent of individuals as Tanypodini
Scraper and/or piercer taxa presence
Shredder taxa presence
Collector taxa presence
Presence of Thienemanniella
Presence of Endochironomus nigricans
Presence ofParachironomus
Presence of Polypedilum
Presence ofAblabesmyia
Presence ofAspectrotanypus algens
Presence ofParamerinasmithae
Presence of Psectrotanypus dyari
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Presence of Zavrelimyia thryptica
Presence of Tanytarsus
Minnesota (Gernes and Helgen 1999):
Invertebrates were sampled in 27 depressional wetlands in the Central Hardwood Forest ecoregion,
using standardized dipnetting and bottle traps. Ten metrics were tested (see the report for full details):
% Corixidae + % Coleoptera
% Erpobdella leeches
sum of %'s of 3 most abundant taxa
# of genera of caddisflies, mayflies, and sphaerid clams
# of genera of midges
# of intolerant taxa (Leucorrhinia, Libellula, Tanytarsus, Procladius, Triaenodes, Oecetis)
# of leech genera (Hirudinidae)
# of genera of dragonflies and damselflies
# of snail taxa (mostly to species)
total richness
The metrics based on leeches and snails were least effective in discerning the land use gradient. The
most discerning metrics were the # of intolerant taxa, the # of genera of dragonflies and damselflies,
and # of genera of caddisflies, mayflies, and sphaerid clams. Presence offish in some of the wetlands
apparently did not confound the use of the multimetric index to discern the disturbance gradient.
Another study in Minnesota sampled invertebrates in over 100 wetlands belonging to 8 types, with
sites representing each type selected to span a gradient of land cover. Results of testing over a dozen
candidate metrics for positive or negative correlations with land cover types at various distances
around each wetland are reported at:
http://www.hort. agri .umn. edu/mnwet/begin.htm
Prairie wetlands (Anderson and Vondracek 1999):
Aquatic insects were sampled in light traps placed near 126 wetlands during 2 years. Surrounding land
cover was mainly grassland or cropland. In general, the effects of surrounding land cover were
statistically overshadowed by geographic region and daily weather. However, the following metrics
showed some usefulness as indicators of surrounding land cover condition, especially when data were
separated by ecoregion:
abundance of aquatic insects
abundance of Caenidae mayflies
abundance of Scarabidae beetles
abundance of Lepidoptera moths
mayfly richness
Lake Huron wetlands (Burton et al. 1999):
In a comparison of three relatively pristine and three impacted Lake Huron wetlands, 24 potential
metrics were tested for their effectiveness at discriminating between altered habitats. Fourteen of the
metrics appeared to respond reliably to disturbance across a variety of wetland vegetation types,
although the sensitivity of the metrics differed (Table 5.4).
Table 5.4. Macroinvertebrate community metrics tested in 3 impacted and 3 unimpacted Lake Huron
wetlands (Burton et al. 1999).
Metrics that displayed the same response over all vegetation types are bolded.
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Direction of response with increasing disturbance
in each vegetation type1
METRIC
OS IS
TY WM ALL
# of Crustacea + Mollusca genera
# of Ephemeroptera + Trichoptera genera
# of Ephemeroptera genera
# of Odonata genera
# of Trichoptera genera
Total # of taxa
Total # of genera
Total # of families
% Amphipoda
% Midges
% Crustacea + Mollusca
% Ephemeroptera
% Gastropoda
% Isopoda
% Odonata
% Spheariidae
% Tanytarsini
% Trichoptera
% Diptera
% Crustacea
Evenness («F)
Shannon Index (H')
Simpson Index (D)
D
N
N
D
I
D
D
D
N
I
D
D
D
D
D
D
I
I
I
N
D
D
I
D
D
D
D
D
D
D
D
N
I
D
I
D
D
D
D
I
D
I
D
D
D
I
D
N
N
D
N
D
D
N
I
D
N
D
D
N
D
D
N
I
D
N
D
D
I
D
N
N
D
N
D
D
N
N
D
N
D
D
I
D
D
D
D
N
I
D
D
I
D
N
N
D
N
D
D
D
N
N
N
N
D
N
D
D
N
N
N
N
D
D
I
:OS = Outer Scirpus, IS = Inner Sc/rpMs/Pickerelweed, TY = Typha, WM = Wet meadow, ALL = All sampling stations
combined.
Reference Values
From a review of regional literature, one might begin to quantify the "expected" range of variation -
both natural and unnatural - in density and species richness, and also to specify taxa that characterize
least-altered (reference) wetlands of each type in the region. Knowing such reference conditions
provides a basis for comparison (benchmark) useful for interpreting data collected in future studies.
However, such a compilation of values and taxa would be confounded by the lack of commonality of
field methods, equipment, units of measurement, and schemes for classifying habitats within wetlands
and the wetlands themselves, as well as by unknown biases in selection of study sites. Nonetheless,
much useful information of this type is presented in the book edited by Batzer et al. (1999) and from
other sources, and is available (at a minimum) for: Florida Everglades, forested limesink wetlands of
Georgia, snowmelt ponds in Wisconsin, beaver ponds and constructed marshes in Pennsylvania,
riparian sedge meadows in Maine, pitcher plant bogs of eastern Canada, depress!onal wetlands of
Ohio, prairie potholes, riverine and prairie wetlands of Minnesota, lacustrine wetlands and prairie
wetlands (Wrubleski 1999) of Manitoba, High Plains wetlands of Wyoming, playas of the southern
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High Plains, wetlands of Yellowstone and Grand Teton parks, tidal freshwater marshes of Virginia,
seasonal and semipermanent wetlands of California, and urban wetlands of western Washington .
5.12 Literature Cited
Adamus, P.R. 1996. Bioindicators for Assessing Ecological Integrity of Prairie Wetlands. EPA/600/R-96/082. USEPA
Environmental Research Laboratory, Corvallis, OR.
Adamus, P.R. and A. Gonyaw. 2001. National Database of Wetland Invertebrate Tolerances. Prepared for the USEPA.
Internet address: http://www.epa.gov/owow/wetlands/bawwg/publicat.html (available in mid-2001).
Alikhan, M.A., Bagatto, G. and Zia, S. 1990. The crayfish as a "biological indicator" of aquatic contamination by heavy
metals. Water Research 24(9): 1069-1076.
Allard, M. and P.M. Stokes. 1989. Mercury in crayfish species from thirteen Ontario lakes in relation to water chemistry
and smallmouth bass (Micropterus dolomieui) mercury. Canadian Journal of Fisheries and Aquatic Sciences 46(6): 1040-
1046.
Anderson, DJ. and B. Vondracek. 1999. Insects as indicators of land use in three ecoregions in the Prairie Pothole Region.
Wetlands 19:648-664.
Ankley, G.T., D.A. Benoit, J.C. Balogh, T.G. Reynoldson, K.E. Day, and R.A Hoke. 1994. Evaluation of potential
confounding factors in sediment toxicity tests with three freshwater benthic invertebrates. Environ. Toxicol. Chem.
13(4):627-635.
Balding, T. 1992. Distribution, abundance, and diversity of mollusks (Bivalvia: Unionidae) from the Lower Chippewa
River, Wisconsin. Transactions of the Wisconsin Academy of Sciences, Arts, and Letters 80:163-268.
Barbour, M.T., J.B. Stribling, and J.R. Karr. 1995. Multimetric approach for establishing biocriteria and measuring
biological condition, pp. 63-80 IN: W.S. Davis, and T.P. Simon (Eds.) Biological Assessment and Criteria. Tools for
Water Resource Planning and Decision Making. Lewis Publishers, Boca Raton, FL.
Barton, D.R., and J.L. Metcalfe-Smith. 1993. A comparison of sampling techniques and summary indices for assessment
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Section 6. Fish
6.1 Use as Indicators
This section addresses fish that are closely associated with non-tidal wetlands. Much, perhaps
most, of the literature is from studies of lacustrine fringe wetlands (e.g., littoral vegetation). For
a general discussion of the topic based on pre-1990 scientific information, and for discussion of
advantages and disadvantages of using fish as indicators of wetland integrity, readers may refer
to Adamus and Brandt (1990).
In contrast to fish monitoring in streams (e.g., Bramblett & Fausch 1991), there have been few
attempts to develop fish IBIs (indices of biotic integrity) specifically in non-tidal wetlands of
North America. In Minnesota, wetland and stream fish were used together to represent the
condition of landscapes that contained a large wetland/ riparian component (Galatowitsch et al.
1998, Mensing et al. 1998). Sampling 15 wetlands belonging to each of 7 wetland types, the
investigators found positive or negative correlations of the following metrics with a site
disturbance score and/or various land cover types measured within 500, 1000, and 2500 m of
each wetland:
• In small-sized river floodplains: fish species richness, proportion of Cyprinids
• In medium-sized river floodplains: fish species richness
• In large river floodplains: the proportion of piscivores, total abundance offish, proportion of
Catostomids
• In non-calcareous littoral wetlands: total fish abundance
• In calcareous wetlands: species richness, total abundance, proportion of Cyprinids, number
of sunfish species
• In forest glacial marshes and prairie glacial marshes: total abundance, richness
Results are reported by wetland type at: http://www.hort.agri.umn.edu/mnwet/
Attempts to apply a fish IBI to lacustrine wetlands in 60 Florida lakes met with mixed success
(Schultz et al. 1999). Fish IBI scores increased with increasing nutrients (lake trophic status) and
lake surface area. Metrics used in the IBI were: total fish, native fish, Lepomis, piscivores,
generalists, insectivores, and intolerant and tolerant species. Attempts to develop and apply a
fish IBI to isolated wetlands near Lake Michigan are described by Simon (1998a,b).
The effects on North American fish of many human-associated factors are summarized by Miller
et al. (1989) and Hughes andNoss (1992).
6.2 Effects of Enrichment, Eutrophication, Reduced Dissolved Oxygen
As eutrophication increases plant and algal productivity, fish sometimes suffer from reduced
levels of dissolved oxygen, and feeding habits also may shift. Biomass and species richness may
increase or decrease, depending on the initial state of the wetland and the duration and magnitude
of the eutrophi cation. To some degree, fish families can be grouped according to decreasing
susceptibility to oxygen deficiencies: salmonids and coregonids require high levels of dissolved
oxygen, whereas Cyprinids often tolerate low dissolved oxygen levels (Harper 1992).
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In a study of an Everglades wetland, Gambusia holbrooki, Heterandria formosa and some other
small fish were 2 to 3 times more abundant in enriched wetlands than unenriched wetlands
(Rader and Richardson 1994) and biomass (standing stock) was greater (Turner et al. 1999). An
isolated Florida wetland receiving treated wastewater supported similar diversity and higher
abundance offish, compared with its condition prior to receiving the wastewater (Schwartz et al.
1994). In another wetland receiving advanced secondary treated wastewater in Central Florida,
fish populations were similar to those in a wetland not receiving wastewater (Best 1993). A
survey of 60 natural lakes in Florida found pygmy killifish, lined topminnow, chain pickerel, and
redfm pickerel in lakes with the least phosphorus and/or nitrate (Hoyer and Canfield 1994).
In an Alaskan tundra river system, fertilization with phosphorus nearly doubled the size of
young-of-the-year fish, suggesting that phosphorus limitation of algal production had limited the
fish previously (Deegan & Peterson 1992).
6.3 Effects of Contaminant Toxicity
Smaller fish may be the first members of a fish community to elicit a response to contaminants
(Matuszek et al. 1990) due to their high metabolic rate relative to larger fish (Shuter and Post
1990). The body burden of copper, cadmium, and silver in minnows can be a strong indicator of
biologically available heavy metals and thus of ecological impact (Birge et al. 2000). The
toxicity of copper and zinc to some fish species depends on other chemical characteristics of
waters (Munkittrick & Dixon 1992, Welsh et al. 1993, Erickson etal. 1996), as well as fish
behavior (Pourang 1995). For example, presence of 5 mg C/L as dissolved organic matter from a
marsh kept copper from binding to the gills of small Oncorhynchus mykiss. This occurred due to
the complexing of copper with dissolved organic carbon, making the copper unavailable (Hollis
etal. 1997). Some fish species appear capable of becoming acclimatized to moderately elevated
levels of some metals (Klerks & Lentz 1998).
The impact of combined contamination from heavy metals and acidification was investigated in
a Canadian beaver pond receiving ore smelting effluent from an abandoned industrial center
(Rutherford and Mellow 1994). Besides contributing nickel, copper, aluminum, iron, lead, zinc,
arsenic, cadmium and chromium, the effluent acidified the water. Immediately downstream
from the source, no living fish were captured during 4 months of seining effort, implying either
extreme avoidance of the area by fish or severe mortality of resident fish. Fish captured within
0.1 km upstream of the contamination included Culaea inconstans, Phoxinus neogaeus,
Phoxinus oes, andPimephalespromelas.
Bioaccumulation of mercury in fish is a growing concern in many North American lakes and
wetlands. Properties of individual lakes appear more important for determining fish tissue
mercury concentrations than do small-scale ecoregional differences. In a survey of 24
Massachusetts lakes relatively unimpacted by mercury, lake trophic state did not influence
mercury levels in fish tissues, whereas pH was highly (inversely) correlated with mercury in fish
tissue, at least in yellow perch and brown bullhead (Rose et al. 1999). Mercury accumulation by
fish is especially great during the first 2 years an area is flooded (Kelly et al. 1997, Bodaly &
Fudge 1999) but does not always occur (Miles & Fink 1998). Tissue mercury levels typically
vary greatly by fish size and trophic level. In 13 Ontario lakes, mercury in tissues of smallmouth
bass was positively correlated with crayfish mercury levels (Allard and Stokes 1989). In Florida,
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mercury found in mosquitofish was lower than in bass and was lower in the wetland interior than
in the inflow and outflow channels (Miles & Fink 1998).
Selenium is not directly toxic to fish at usual concentrations, but can become toxic once
concentrated in fish food chains, especially in some wetlands that receive effluents from irrigated
fields or power plant reservoirs in some regions (Zilberman 1991, Lemly 1996).
Synthetic organics, including pesticides, can accumulate in wetland fish (Cooper 1991), often
with adverse effects. In a Canadian wetland receiving oil sand effluent, fish had altered blood
chemistry and died within 14 days (Bendell-Young et al. 2000). Two herbicides used to control
the invasive wetland plant, purple loosestrife (Lythrum salicaria), were not toxic to rainbow trout
(Gardner and Grue 1996). Sensitivity of smaller non-game wetland fish was not investigated.
Fish exposed to pesticides, PCB's, and other synthetic organics are often more vulnerable to
disease, and the literature on this subject was reviewed by Dunier & Siwicki (1993).
6.4 Effects of Acidification
Surveys of literature on effects of acidification on fish, including pH values critical to fish
response, are provided by Baker and Christensen (1991), Minns et al. (1990), Carline et al.
(1992), and others. Acidity can be directly toxic to fish, inhibit reproductive maturation, inhibit
spawning behavior, induce emigration, and alter food availability (Baker and Christensen 1991).
Acidity induces aluminum toxicity in fish in many lakes and wetlands, although in Florida, soil
characteristics made this less problematic (Keller and Crisman 1990). Acidities in the range of
5.0 to 6.0 pH are critical for several species. In an Ontario beaver pond, a pH of 3.7 resulted in
complete mortality due to acid-related release of toxic metals from sediments (Rutherford and
Mellow 1994). Models for fish response to acidification are provided by Charles (1991). Ten
species of the Upper Midwest - four cyprinids, three percids, Coregonus artedii, Percopsis
omiscomaycus, and Lota lota - were not caught in waters with pH < 6.0 (Cook and Jager 1991,
Cusimano etal. 1990). Percaflavescens and Umbra limi were the most acid-tolerant in the
studies of lakes in Michigan and Wisconsin. Other acid-tolerant species included Culaea
inconstansjctalurus nebulosus, Lepomis macrochims, andL. gibbosus (Cook and Jager 1991).
However, factors such as recreational fishing, lake size, and predation confound attempts to
attribute fish absence to a particular contaminant (Cook and Jager 1991). Table 6.1 classifies
fish by acid sensitivity in waters of upstate New York (Schofield and Driscol 1987 as cited in
Charles 1991).
Table 6.1. pH associations offish of the north branch of the Moose River, New York (from
Schofield and Driscol 1987).
Sensitive - found only in Intermediate - found in both Tolerant - found in waters
waters with pH > 6.0 types of waters withpH<5.0
Blacknose dace Brook trout Mudminnow
Redbellied dace White sucker Killifish
Common shiner Creek chub Brown Bullhead
Smallmouth bass Pumpkinseed sunfish Golden Shiner
Rock bass Finescale dace Yellow Perch
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Among 426 Ontario lakes, cyprinid species declined when pH was below 6.0, and no cyprinid
species were found below pH 4.7. Lakes below pH 5 tended to be dominated by Perca
flavenscens, Lepomis gibbosus and Ambloplites rupestris. Above pH 6, lake size (across a range
of 10-1585 ha) influenced fish species richness more than pH, or neither pH nor lake size had
any discernable effect (Matuszek etal. 1990) Regional environmental differences can mediate
the impact of water quality characteristics. For example, subtropical Florida lakes had higher
numbers offish species than similar temperate lakes despite having approximately the same pH
(Keller and Crisman 1990). In Florida, 11 fish species were found in lakes with a minimum
measured pH of 4.3 (Hoyer and Canfield 1994). These included the lined topminnow,
Everglades pygmy sunfish, pygmy killifish and redfm pickerel.
6.5 Effects of Salinization
The age structure and growth rate of Micropterus salmoides from a brackish marsh and a
freshwater oxbow lake system was investigated in south-central Louisiana (Meador and Kelso
1990a, b). Marsh fish exhibited small size and reduced length at age. However, growth rates of
older marsh fish equaled or exceeded those of freshwater largemouth bass. Brackish water fish
maintained better body condition throughout the year whereas freshwater fish exhibited reduced
condition during early spring and fall. Laboratory trials consisting of 120-day exposure of marsh
and freshwater largemouth bass to four salinity levels (0, 4, 8, and 12 ppm) indicated a
significant decrease in growth rate of freshwater largemouth bass with increasing salinity level
up to 8 ppm. No such decrease was observed for marsh largemouth bass. All fish held at 12 ppm
stopped feeding within one week and died before the end of the experiment.
In another experiment, juvenile bluegill (Lepomis macrochirus) from a freshwater pond in
northeastern Mississippi and a brackish bayou in coastal Mississippi were held in a chamber with
0 ppm salinity but given access to chambers containing 0, 2, 4, 6, 8, and 10 ppm salinity
(Peterson et al. 1993). Fish from neither habitat showed clear preference for any of the salinity
options. These data and data from previous studies suggest bluegill are better able to
physiologically and behaviorally tolerate elevated salinity relative to other centrarchids,
particularly bass (Micropterus) (Peterson et al. 1993).
A model for assessing effects on wetland fish of major cations from irrigation runoff was tested
and described by Dickerson et al. (1996).
6.6 Effects of Sedimentation/Burial
No recent studies specific to North American wetlands or lakes were found.
6.7 Effects of Thermal Alteration
Laboratory experiments and field observations of centrarchid, poeciliid, and cyprinodontid fishes
in the Florida Everglades demonstrated that suddenly cooler temperatures (15 °C minimum daily
temperature) led to fish behavioral changes that might affect higher trophic levels by reducing
prey availability (Frederick and Loftus 1993). Throughout Ontario, lake elevation (a surrogate
for temperature and possible absence of predatory fish) had a statistically strong and positive
influence on abundances of yellow perch and white sucker (Hinch et al. 1994).
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6.8 Effects of Vegetation Removal
Removal of vegetation from within or alongside wetlands affects wetland fish largely by
increasing water temperature and susceptibility to predation, and by altering foods and their
availability. Woody material is especially important as a source of cover for fish in off-channel
wetlands (e.g., oxbows, sloughs) and in lakes (Leitman et al. 1991, Dewey & Jennings 1992,
Fausch and Northcote 1992, Mclntosh et al. 1994), but shoreline trees are often cut to improve
landowner views or for firewood (Christensen et al. 1996). Larval fish prefer structurally
complex habitats in Mississippi oxbow lakes (Killgore and Miller 1995, Killgore and Baker
1996) and probably in other systems.
In lacustrine fringe wetlands, submerged macrophytes are particularly important. For example,
declines in macrophytes (resulting from grass carp introductions, Bain 1993) have been linked to
an increase in the proportional abundance of limnetic (open water) fish species (Bettoli etal.
1991, Maceina et al. 1991, Martin et al. 1992). Macrophyte diversity and coverage also were
identified as key determinants of species composition offish communities in Lake Michigan
wetlands (Brazner and Beals 1997). Nonetheless, intentional thinning of extensive macrophyte
beds can result in higher growth rates of some age classes of lake fish, presumably due to
improved access offish to invertebrate foods (Olson et al. 1998).
At least for some fish, there is little difference in their propensity to use aquatic beds of native
plants as opposed to aquatic beds of non-native, invasive plant species (Conrow et al. 1990,
Duffy &Baltz 1998).
6.9 Effects of Turbidity, Shade
Turbidity, partly through its adverse impacts on submersed macrophytes and algae, affects the
structure of many wetland fish communities (e.g., Lake Michigan wetlands, Brazner and Beals
1997). Bottom-feeding carp (Cyprinus carpid) often are a major contributor to turbidity in
shallow lakes and wetlands, due to their large size and regular stirring of sediments as they feed.
Turbidity, total phosphorus, and total ammonia concentrations increased predictably with total
biomass of carp stocked in experimental enclosures within an Ontario marsh (Lougheed et al.
(1998) Although carp had no direct effect on zooplankton community structure, zooplankton
biomass declined due to increased turbidity and altered nutrient availability.
Shade produced by increased cover of duckweed after the addition of wastewater to a blackwater
wetland in Central Florida caused a reduction in Gambusia (Smith 1992).
6.10 Effects of Dehydration/Inundation
Spring-fed isolated wetlands in the western United States provide the only habitat for several rare
fish (Meffe 1989). Among isolated wetlands of the upper Atlantic Coastal Plain, fish (as
expected) inhabit only those wetlands that dry infrequently and/or have elevations similar to
nearby permanent waters that periodically serve as a source of colonists (Snodgrass etal. 1996).
Along the lower Missouri River, twice as many fish species were found in sloughs regularly
connected to the main channel as in isolated sloughs that were not. The fish community in the
connected sloughs also contained a larger component of native species (Galat et al. 1998). In the
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northern United States, isolation of small lakes from other water bodies also influences fish
species composition (Magnuson et al. 1998).
In riverine systems, the availability of off-channel or in-channel pools that hold water through
dry periods and provide calm-water feeding areas during floods is a critical feature in the
stability offish communities in both riverine and non-riverine wetlands (Gelwick 1990, Capone
and Kushlan 1991, DeAngelis et al. 1997). Floods that create such areas, by periodically
rearranging the geomorphic structure of floodplains and importing large woody debris from
normally unflooded areas are essential to maintaining fish habitat in large rivers (Galat et al.
1998, Michener et al. 1998, Shields et al. 2000). Floods also are important because they provide
adult fish with temporary access to a rich supply of relatively unexploited floodplain foods
(Leitman et al. 1991, Killgore & Baker 1996, Jordan et al. 1998).
Declines in water level, whether drought-related or human-caused, alter community structure of
wetland fish, partly by creating greater overlap of resource utilization and increasing predation
risk. In a North Dakota stream that experienced 6 years of drought, species richness declined
from 23 to 13 species (Kelsch 1994). Low water also increases chances offish freezing in winter
and being subjected to lethal thermal stress in summer. Sustained drawdowns, by temporarily
eliminating larval dragonflies and other large invertebrates that normally compete with or prey
on larval fish, also can reduce competition among fish that return to wetlands when water levels
rise again (Travnichek & Maceina 1994, Jordan et al. 1996). Isolated wetlands that experience
frequent near-drought conditions tend to support mainly smaller species, and small individuals of
normally larger species (Loftus and Eklund 1994).
6.11 Effects of Other Stressors
Non-native fish have been widely introduced into waters of the United States, both on purpose
and by accident. Pollution and alteration of water regimes sometimes accelerate invasion of
natural habitats by non-native fish species, which tend to be broadly tolerant. Effects of invaders
on native fish communities are usually adverse (Baltz and Moyle 1993), especially when coupled
with simultaneous impacts from other factors (Larimore & Bayley 1996, Marschall & Crowder
1996). Consequently, dominance by non-native species is often used as a measure of low
biological integrity (Fausch etal. 1990, Farr & Ward 1993), although native fish communities
may sometimes be impacted by watershed urbanization even in the absence of invasion by non-
native species (Weaver and Garman 1994). A survey of the St. Louis River estuary revealed the
invasive Gymnocephalus cernuus was the fifth most abundant species in areas impacted by
human disturbance and the 15th most abundant species in the relatively unimpacted inner marsh.
Invasion by this species appeared to be inhibited by the presence of dense wetland vegetation
(Brazner et al. 1998). Careful sampling of New England lakes indicates that assemblages of
minnows have been devastated by introduction of non-native predatory fish (present in 69% of
the lakes) and by often-accompanying development of shorelines and watersheds (Whitter et al.
1997).
Like larval amphibians, young fish of some species have been found to be highly sensitive to
ultraviolet-B radiation (Ewing et al. 1999).
A diked wetland on Lake Erie was found to support 23 fish species, whereas an undiked wetland
nearby supported 40. The five most abundant fish in the diked wetland were Pomoxis annularis,
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Amieurus nebulosus, Amieurus melas, Carassius auratus and Cyprinus carpio. Comparatively,
the five most abundant fish in the undiked wetland wereMorone americana, Dorosoma
cepedianum, Pomoxis annularis, Amieurus nebulosus., and Percaflavescens. Mean summertime
lengths of several fish species were significantly greater in the undiked wetlands (Johnson et al.
1997).
Suitably-designed constructed wetlands have been reported as usually supporting a richness and
abundance offish comparable to or greater than natural wetlands (e.g., Langston and Kent 1997,
Morrow 1999). However, mean abundance and biomass of G. holbrooki andE1. evergladei in
some constructed Florida wetlands differed from natural wetlands there (Streever and Crisman
1993). The capacity of stream fish communities to recover from disturbance is discussed by
Meffe & Sheldon (1990), Detenbeck et al. (1992), and Kinsolving & Bain (1993). Wetland fish
populations in the Florida Everglades may require up to 1 year to recover from drought
(DeAngelis et al. 1997).
6.12 Wetland Monitoring
Spatial and Temporal Variability
In six Wisconsin lakes, annual variation in fish community structure was minor, indicating that
each lake was fairly stable in time. In contrast, variation in fish communities among lakes was
much greater, a consequence of substrate, macrophyte and depth variations (Benson and
Magnuson 1992). Low species richness in one of the lakes was accompanied by overwhelming
dominance of a few species (Magnuson and Lathrop 1992). Within a lacustrine fringe wetland of
Lake Huron, fish richness and abundance decline 40-70% with distance from the open water
edge (Bouchard 1998). In a Texas reservoir, a single habitat type hosted a variety offish species,
depending on the season (Gelwick and Matthews 1990). Fringing wetlands provide spawning
and rearing areas in most lakes, and in large lakes may be much more important than their
proportionate size alone would suggest (Jude and Pappas 1992). Lake fish communities tend to
be the result of multiple structuring factors, events, and processes that produce communities that
are nearly unique to each water body (Tonn 1990, Tonn etal. 1990). The spatial distribution of
fish in Florida wetlands also changes with season (Jordan et al. 1998).
Techniques and Equipment
Fish are captured in wetlands and lakes using a variety of methods. Information on equipment
and guidance for sampling fish in streams and/or lakes is provided by Meador et al. (1993),
Murphy and Willis (1996), and many reports that pre-date the review period covered by this
document. Equipment includes but is not limited to:
• seines (Pierce et al. 1990, Rutherford and Mellow 1994)
• fyke nets (Weaver & Magnuson 1993, Brazner et al. 1998)
• dip nets
• throw trap (Chick et al. 1992, Jordan et al. 1997)
• pop nets (Petering and Johnson 1991, Dewey 1992)
• minnow traps (He & Lodge 1990)
• electroshocking (Dewey 1992)
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Although more difficult and potentially dangerous, sampling at night is often much more
effective than daytime sampling, and frequently reveals species assemblages and habitat use
patterns much different than found during the day.
Very different kinds of gear are needed for sampling fish larvae. This may include plankton tow
nets (Conrow etal. 1990) and floating light traps (Killgore & Baker 1996)
A variety of biochemical and other methods are of potential use for assessing health of individual
fish in relation to toxic chemicals. A review was published by Niemi (1990).
Metrics for Assessing Impacts to Wetland Fish Communities
Information on this topic as well as sampling equipment is reviewed on EPA's wetland
biomonitoring web page: http://www.epa.gov/owow/wetlands/bawwg
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Section 7. Amphibians And Reptiles
7.1 Use as Indicators
This section addresses the monitoring of turtles, frogs, toads, salamanders, newts, snakes,
crocodilians, and lizards that occur in wetlands. Because most amphibians and many reptiles
require aquatic habitats, they are especially vulnerable to alteration or contamination of wetlands
(Dodd and Cade 1998, Stebbins and Cohen 1995, Lannoo 1998, Pough etal. 1998, Richter and
Azous 1995, 2000, Olson and Leonard 1997). Although amphibians have shown some promise
as indicators of wetland and/or landscape integrity, no "indices of biotic integrity" based solely
on amphibian community composition have yet been developed and validated successfully.
Much information on the ecology of tadpoles is summarized by McDiarmid & Altig (1999).
In the decade since our last review, scientific and public concern has increased over the decline
of amphibian and reptile populations in North American wetlands (Blaustein and Wake 1995,
Cohn 1994, Halliday 1993, Livermore 1992, Wake 1991, Wyman 1990, Phillips 1990,
Pechmann etal. 1991). Amphibian decline has been well documented (Phillips 1990, Wyman
1990, Wake 1991, Crump etal. 1992, Barinaga 1990, Blaustein and Wake 1990), although
causes are not yet well understood definitively. A near-absence of long-term studies hinders
conclusions regarding the hypothesis that amphibians are experiencing an overall decline due to
human activities (Blaustein 1994, Pechmann and Wilbur 1994). Declines have been attirbuted to
multiple factors acting singly or in combination (Blaustein and Wake 1990, Sarkar 1996), with
diseases and parasites being suggested most often as direct or indirect causes of decline (Carey
and Cohen 1999). In the Pacific Northwest alone, amphibian decline or changes in species
composition have been linked partly to pathenogenic fungi (Blaustein et al. 1994a, Kiesecker &
Blaustein 1999), ultraviolet-B radiation (Blaustein etal. 1994b, 1995), agricultural runoff (Boy er
1993, Boyer and Grue 1995, Marco et al. 1999), and introduced species (Hayes and Jennings
1986). Investigators also have found some amphibian life stages to be negatively affected by pH
and a variety of chemical contaminants (Beattie and Tyler-Jones 1992, Rowe et al. 1992,
Sadinski andDunson 1992, Rowe andDunson 1993, 1995, and see section 7.3 below).
In the past decade a few studies have begun to use amphibian assemblages specifically to
indicate the ecological condition of a large series of wetlands, e.g., Richter & Azous (2000). In
Minnesota, amphibians were used to represent the condition of landscapes that contained a large
wetland/ riparian component (Galatowitsch et al. 1998, Mensing et al. 1998). Sampling 15
wetlands belonging to each of 4 wetland types, the investigators found positive or negative
correlations of the following metrics with a site disturbance score and/or various land cover types
measured within 500, 1000, and 2500 m of each wetland:
• In small-sized river floodplains: total abundance, abundance of leopard frog
• In medium-sized river floodplains: species richness
• In forest glacial marshes: total abundance, richness
• In prairie glacial marshes: total abundance, richness, abundance of leopard frog
Results are reported by wetland type at: http://www.hort.agri.umn.edu/mnwet/
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Other efforts to develop wetland indices of biotic integrity (IBI's) using amphibians are
underway in Ohio, Maryland, Maine, and elsewhere.
7.2 Effects of Enrichment/ Eutrophication/ Reduced Dissolved Oxygen
Sublethal responses and mortality of 5 Oregon amphibian species following laboratory
applications of nitrate indicated that EPA nitrate criteria for drinking water and/or for protection
of warm water fish are inadequate to protect these amphibians (Marco et al. 1999). In Texas,
playa wetlands receiving nutrient-laden feedlot effluent were devoid of amphibians found in
natural playas (Chavez et al. 1999). Experiments indicated that effluent had to be diluted to less
than 3% strength in order to minimize adverse effects on leopard frog (Ranapipiens).
Indirect effects of excessive nutrients can also be important. Shifts in seasonal timing and
amount of nutrients that enter a wetland can, over a period of years, increase the relative
dominance of algae and/or emergent plants at the expense of submersed plants (see Sections 2
and 3). This in turn can reduce the availability of submersed plants as attachment substrates for
amphibian eggs and as cover for larvae (Beebee 1996). It can also diminish dissolved oxygen
levels (Tattersall and Boutilier 1999), alter the abundance of aquatic predators, and shift the algal
and invertebrate foods available to amphibians (Home andDunson 1995b). As a result, species
composition and sometimes species richness of amphibian communities can decline as
eutrophication becomes severe, but well-designed studies of such effects are few. In
Pennsylvania, 10 duckweed-covered wastewater ponds were compared with 10 naturally-
occurring ponds over a 19-week period. Egg hatching and survival of all amphibian species was
lower in the wastewater ponds (Laposata & Dunson 2000). In the southeastern United States,
flatwoods salamanders (Ambystoma cingulatum) are not found in wetlands with excessive
amounts of algae (Palis 1996).
7.3 Effects of Contaminant Toxicity
Studies of the effects of heavy metals, pesticides, and other toxins on reptile and amphibian
communities have mainly been conducted at the species (not community) level of organization.
A review of much of the relevant literature was published by Sparling et al. (2000). As a partial
starting point for formulating indices of biotic integrity, Schuytema & Nebeker (1996) compiled
a database of toxicity information from published literature, for 58 amphibian species as related
to 135 chemicals. A similar toxicological database was compiled by Pauli et al. (2000).
Toxicity of aluminum and other metals has been the focus of studies of the embryos and tadpoles
ofBufo americanus (Birge et al. 1992, Freda 1990, Freda and McDonald 1993), B. canorus
(Bradford et al. 1991), Ambystoma maculatum (Freda and McDonald 1993), Hyla crucifer
(Glooschenko etal. 1992), Rana sylvatica (Freda and McDonald 1993), R. muscosa (Bradford et
al. 1991), and R. pipiens (Freda and McDonald 1990, Freda etal. 1990, Freda 1991, Freda
1989). Significant variation exists in the suceptability of amphibians to aluminum and pH.
Aluminum toxicity was correlated with the mortality of R. sylvatica tadpoles but not with
mortality of R. sylvatica., Bufo americanus., and Ambystoma maculatum embryos, which were
influenced more by pH (Freda and McDonald 1993, Freda et al. 1991). Aluminum treatment at
pH 5.0-6.0 resulted in reduced survivorship in B. canorus embryos but notR. muscosa embryos.
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Although not lethal, moderately reduced pH and elevated aluminum was associated with
sublethal effects to B. canorus and R muscosa embryos, such as reduced growth rates, increased
frequencies of developmental abnormalities, and earlier hatching (Bradford etal. 1994).
Aluminum can ameliorate the toxicity of acidic conditions at some pH levels while becoming
toxic at other pH levels. The difference between toxicity and non-toxicity can be quite narrow.
Aluminum buffered the toxic effects of acidity on Rcmapipiens embryos and pre-stage tadpoles
in the 4.2- 4.4 pH range but became toxic in the 4.6-4.8 pH range (Freda and McDonald 1990).
The frog, Hyla crucifer, was mostly absent in Ontario ponds downwind of a smelting operation,
which had higher levels of cadmium, nickel, and sulfate (Glooschenko etal. 1992). Cadmium
and lead readily accumulated in the frog, Rana ridibunda (Vogiatzis and Loumbourdis 1998,
1999) and other amphibian species (Herkovits & Perez 1993, Steele et al. 1999). The
fertilization success of Rana heckscheri declined markedly when exposed to mercuric chlorides
(Punzo 1993a,b). Development was completely blocked at concentrations of 5.0 mg/L of
mercuric chloride. Alligators readily concentrate mercury (Heaton et al. 1997, Yanochko et al.
1997) and concentrations tend to be unrelated to the length of sampled individuals, and thus
presumably their age (Jagoe et al. 1998). Some studies have linked local declines of wetland
amphibians with presumed applications of agricultural chemicals in adjoining areas (Berrill etal.
1997, Howe etal. 1998).
Many synthetic organic compounds affect amphibians and aquatic reptiles. Petroleum
derivatives have been noted to stunt tadpole growth of Hyla cinerea (Mahaney 1994) as well as
reduce development time, growth, and survival in frogs and toads (Pollet et al. 2000). In areas
with high oil concentrations, no tadpoles successfully metamorphosed. Northwestern
Salamander (Ambystoma gracile) egg mortality corresponded with levels of total petroleum
hydrocarbons in western Washington (Platin 1994, Platin and Richter 1995). Snapping turtles
(Chelydra serpentina serpentina) with higher exposures to polychlorinated aromatic
hydrocarbons (PAH's) in Ontario and New York had a higher incidence of abnormal
development (Bishop et al. 1990, 1995). DNA damage was found in slider turtles (Trachemys
scripta) and Chelydra serpentina that had been exposed to radionuclides and chemical
contamination (Lamb et al. 1991, Meyers-Schone et al. 1993). Eggs of the turtle Chelydra
serpentina from sites with the greatest pollution had the highest rates of abnormalities (Bishop et
al. 1990). The pesticide, esfenvalerate, caused damaging sublethal effects on tadpoles of Rana
pipiens (Materna et al. 1995). Eisler tested the toxicity to frog tadpoles of paraquat (1990a,b),
cyanide (1991a), diflubenzuron (1991b), fenvalerate (1992b), zinc (1993), and acrolein (1994).
Laboratory exposure of American toad (Bufo americanus americanus) and green frog (Rana
clamitans melanota) to water from a vegetable-growing area in Ontario resulted in a higher rate
of tadpole deformities and lower egg hatching rates than exposure to water obtained from
upstream of the agricultural area; the water contained high nutrient levels and a mix of pesticides
typical of ambient field concentrations (Bishop et al. 1999).
Tests of three forest insecticides (fenitrothion, triclopyr, and hexazinone) on three frog species
(Rana catesbeiana, R. clamitans melanota, R. pipiens) in Ontario suggested that none of the
species were adversely affected by hexazinone exposure, at least not immediately or observably.
Tadpoles of all were sensitive to triclopyr and fenitrothion, with R. pipiens being less sensitive
than,/?, clamitans melanota andR. catesbeiana (Berrill et al. 1991). Embryos and larvae of five
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amphibians (the frogs Rana sylvatica, R. pipiens, R clamitans melanota; the toad Bufo
americanus; the salamander Ambystoma maculatum) were exposed to one or both of the
pyrethroid pesticides permethrin and fenvalerate, and Ambystoma maculatum was found to be
particularly sensitive (Berrill et al. 1993). Densities of mink frogs (Rana septentrionalis) in New
Brunswick, Canada, were lower in ponds with greatest exposure to the insecticide fenitrothion
(McAlpine et al. 1998). Exposure to sublethal levels of the insecticide carbaryl caused
significant and permanent behavioral disruptions in plains leopard frog tadpoles (Rana blairi)
(Bridges 1997). However, considerable variation was noted in the response of various
individuals to this insecticide (Bridges & Semlitsch 2000).
In a study of accumulation of organochlorine chemicals in embryonic turtles, tissue
concentrations peaked at or just before hatching and then declined, which is consistent with
trends reported in developing sea turtles, fish, and birds (Bishop etal. 1995). Morphological
abnormalities in Ontario turtles coincided with accumulation of organochlorines in turtle blood
(Solla et al. 1998). Deformities in Florida alligators also coincided with organochlorine and PCB
contamination, but could not be causally linked to those substances (Guillette et al. 1999).
PCB's were found to accumulate to toxic levels in turtles in Mississippi wetlands (Kannan et al.
2000). In Ontario, PCB accumulation also was noted among green frogs (Rana clamitans) and
snapping turtles (Russell et al. 1997, Bishop et al. 1996). In Illinois, reproductive abnormalities
in cricket frogs (Acris crepitans) were linked to contamination with PCB's and possibly the
herbicide atrazine (Reeder et al. 1998). Atrazine exposure appeared to influence the size and
weight of tiger salamanders in North Dakota (Larson et al. 1998) and some reproductive
hormones in Florida alligators (Grain et al. 1997).
At times, adult amphibians seem unaffected by ambient concentrations of pesticides and other
synthetic organics. In a comparison of frogs in wetlands within apple orchards receiving
pesticide (dichlorodiphenyltrichloroethane [DDT]- or endosulfan-related) treatment and those
outside of the orchard, no significant effects were observed consistently, despite accumulation of
these substances in frog tissue (Harris et al. 1998a,b). When the insecticide endosulfan was
tested for its toxicity to Rana sylvatica, Bufo americanus, and Rana clamitans embryos and
tadpoles in a laboratory setting, the Rana sylvatica embryos hatched successfully and displayed
no adverse morphological effects during the following 10 days, although nearly all tadpoles had
shown paralysis from the insecticide at some point in their development (Berrill et al. 1998). In
another experiment involving endosulfan, pre-metamorphic mortality of Bufo americanus
tadpole mortality was high and occurred at the lowest exposure concentrations (0.041-0.053
mg/L) for each species exposed as 2-week-old tadpoles. Newly hatched Bufo americanus
tadpoles were more tolerant than tadpoles of the other two species, but no clear species
differences in sensitivity of 2-week-old tadpoles were apparent. Bufo americanus pre-
metamorphs exposed to endosulfan did not recover from adverse exposure effects (Berrill et al.
1998).
7.4 Effects of Acidification
Excessive acidity damages amphibians both directly (Home and Dunson 1994b) and as a result
of its capacity to mobilize toxic metals and perhaps, by making sodium less available in some
soil types (Wyman and Jancola 1991).
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Most adult salamander species choose less acidic pools for breeding (Kutka and Bachmann 1990,
Whiteman etal. 1995). In Ontario, the acid-neutralizing capacity (alkalinity) of 38 wetlands
positively influenced the probability ofRanapipiens, R. clamitans, and Hyla crucifer being
present (Glooschenko etal. 1992). Many researchers have expressed concern that temporary
ponds could be the most sensitive freshwater bodies to atmospheric acidification because they
are more isolated from inorganic soil buffers, have less dilution, and are directly exposed to
acidic rainfall and unbuffered snowmelt (Harvey 1989). In the most acidic spawning ponds on
Cape Cod (pH 4.3 to 4.5), spotted salamanders (Ambystoma maculatum) suffered complete
mortality, which was preceded by gross abnormalities (Portnoy 1990). Similar results were
found elsewhere in Massachusetts, in vernal pools (Shortelle et al. 1989). Naturally acidic
streams in the mountains of North Carolina are virtually devoid of salamander larvae (Kucken et
al. 1994). When R. sylvatica, Ambystoma maculatum and A. jeffersonianum were exposed to
acidic conditions (pH 4.2) in temporary wetlands, A. jeffersonianum suffered high mortality for
the acute tests and very high mortality for the chronic study. R sylvatica survived the low pH
but had reduced growth rates. A. maculatum had reduced survival rates in the chronic study
(Rowe et al. 1992). The trend in acid tolerance seems to be : R sylvatica>A. maculatum>A.
jeffersonianum (Rowe and Dunson 1993, Rowe et al. 1992). Embryos of Ambystoma tigrinum
tigrinum had more than 70% survival at pH 4.5 and above, but suffered much greater mortality at
lower pH levels (Whiteman et al. 1995). Tables 7.1 and 7.2 provide critical pH values for some
common amphibians.
Table 7.1. Lethal and critical pH for embryos of eight amphibian species
(adapted from Glooschenko etal. 1992).
Species Lethal pH1 Critical pH2
Hyla crucifer 3.8-4.2 4.2
Hyla versicolor 3.8 4.3
Rana sylvatica 3.5-4.0 3.9-4.25
Ranapipiens 4.2-4.5 4.6
Rana clamitans 3.7-3.8 4.1
Rana catesbeiana 3.9 4.3
Bufo americanus 3.8-4.2 4.0-4.2
Ambystoma maculatum 4.0-4.5 4.5-5.0
1 the pH that kills 85% or more of the embryos within a few hours
2 the pH that causes high embryonic mortality, with effects that are more complex and less direct than lethal effects
(Gosner and Black 1957).
TABLE 7.2. Response of several western amphibian embryos to acid water.
The last column is the lowest tested pH without mortality significantly different from controls
(adapted from Corn and Vertucci 1992 ,with data from Harte and Hoffman 1989, and Corn et al
1989).
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Species LC50 pH level No Mortality pH level
Ambystoma tigrinum 5.3 6.1
Bufo boreas 4.5 4.9
Pseudacristriseriata 4.8 5.2
Ranapipiens 4.5 5.0
sylvatica 4.3 4.6
A number of amphibians inhabit peatlands, e.g., Ranapipiens, R. sylvatica., Bufo americanus,
Ambystoma laterale, Pseudacris triseriata, P. crucifer, andHyla versicolor. However, acidic
water of bogs and marginal fens inhibits embryonic development and hatching of many. Perhaps
least affected is Rana sylvatica, but even that species is affected in some situations (Karns 1992).
The variability in that species' tolerance may depend on genetic or non-genetic maternal factors
(Pierce and Wooten 1991).
Concerns have been raised regarding the vulnerability to acidification of western montane
wetlands. Acidification makes aluminum and cadmium more mobile and increases their
concentration in surface waters. Acidification effects on mobility of lead and inorganic mercury
is less, due to the high binding affinities of these metals for humic substances in soils and
sediments (Scheuhammer 1991). Amphibians (e.g., Jefferson and spotted salamanders) are
known to be sensitive to acidity and elevated concentrations of aluminum found in some acidic
ponds (Blancher 1991, Huckabee etal. 1989, Ireland 1991, Home andDunson 1995b).
However, vernal pools in the mountains of northern Colorado and southern Wyoming rarely had
pH levels below 6.0 during the amphibian breeding season (Corn et al. 1989). In the California
Sierra Nevada mountains, snowmelt water was found to have insufficient acidity to adversely
impact Bufo canorus and Rana muscosa embryos and hatchlings (Bradford et al. 1991, 1992).
Also, the pulse of snowmelt water probably occurs before breeding begins (Corn and Vertucci
1991). Moreover, no evidence was found to suggest that acidification was affecting Pacific
chorus frogs (Pseudacris regilla) in the Emerald Lake watershed of the southern Sierra Nevada
(Soiseth 1992).
Nonetheless, aluminum released into montane pools as a result of acidification sometimes has
harmed embryos, reduced growth rates, and/or caused deformities and premature hatching of
native amphibians (Bradford et al. 1991, Corn and Vertucci 1991, Lamnicky 1990). Survival of
wood frog embryos declined when exposed to aluminum concentrations of 100 (ig/L or greater.
Boreal toad embryos survived exposure to aluminum concentrations of 400 (ig/L (Corn etal
1989). Concentrations of aluminum, sulfate, and zinc were higher in 40 ponds where breeding of
Jefferson salamander (Ambystoma jeffersonianum) was unsuccessful, whereas alkalinity, copper,
dissolved organic carbon, magnesium, sodium, and nitrate were significantly higher in ponds that
supported successful breeding. Low pH increased the time until eggs hatched, decreased
hatching success, and slowed amphibian development rates. Copper and low pH negatively
affected the rate of larval development, whereas aluminum, lead, and sodium did not (Home and
Dunson 1994a). Bullfrog (Rana catesbeiana) tadpoles collected from coal ash deposition basins
contaminated with As, Cd, Cr, Cu, Se and other elements had reduced number of labial teeth and
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deformed labial papillae (Rowe et al. 1996). Deformed tadpoles were less able to graze algae,
which resulted in lower growth rates.
Amphibian vulnerability to acidification of temporary ponds is largely species-specific (Home
and Dunson 1995a). In temporary ponds of central Pennsylvania, the number of egg masses
deposited by Ambystomajeffersonianum was correlated positively with pH and alkalinity and
negatively with Al levels. Egg deposition by A. maculatum was correlated positively with pH
and pond volume and negatively with total cations (Na, K, Mg, Ca, specific conductance) and
silicon. Egg deposition by R. sylvatica was correlated positively with pond volume and dissolved
organic carbon. The results suggest that the breeding success of these amphibians can be
predicted in part by abiotic conditions such as pH, aluminum total cations, dissolved organic
carbon, silicon, and pond volume (Rowe and Dunson 1993). Mortalities ofBufo americanus,
Rana sylvatica, andAmbystoma maculatum embryos were correlated with pH and not to
aluminum (Freda and McDonald 1993). The mortality of Rana sylvatica tadpoles was correlated
both with aluminum and pH. Naturally occurring metals (Al, Cu, Fe, Pb, and Zn) at
concentrations analogous to those observed in prior field studies had variable effects on acute
exposure and survival for both the Jefferson salamander and wood frog (Rana sylvatica) (Home
and Dunson 1995a). Acute exposure to aluminum and copper significantly reduced wood frog
survival. Increased water hardness significantly increased acute exposure wood frog survival.
Acute exposure mortality of the Jefferson salamander was significantly higher in the aluminum
and copper treatments; in toxic metal treatments, survival was higher at the low pH level.
Chronic exposure of wood frogs to aluminum and copper at a higher pH level, and lower water
hardness level greatly reduced survival. Similarly, chronic exposure of Jefferson salamander
larvae to aluminum and copper significantly reduced survival. Newly hatched frog embryos that
were exposed to a combination of pH 4.5 and increasing aluminum concentrations experienced
increased mortality, reduced body size and swimming speed of tadpoles, and increased predation
of tadpoles by dragonfly larvae (Jung and Jagoe 1995). Slower tadpoles are more susceptible to
predation by dragonfly larvae (Richards and Bull 1990a).
Excessively high pH, as well as low pH, is detrimental to amphibians. Experiments suggested
that the current upper regulatory limit of pH 9 may be inadequate to protect tiger salamanders
from detrimental effects of some irrigation and urban wastewater (Abbasi et al. 1989).
7.5 Effects of Salinization
Three studies reported a statistically significant negative correlation between water column
conductivity and amphibian species richness (Azous 1991, Platin 1994, Platin and Richter 1995).
In general, relatively little is known about thresholds of amphibian tolerance to salinity.
7.6 Effects of Sedimentation/Burial, Turbidity, Shade
A single large pulse of sediments into a northern California stream during construction
apparently reduced populations ofAscaphus truei (larvae), Dicamptodon tenebrosus
(paedomorphs and larvae) and Rhyacotriton variegatus (adults and larvae) (Welsh and Ollivier
1998). Deposition of silt, especially in combination with motor oil, resulted in reduced growth
and earlier metamorphosis of larval mole (A. opacum) and tiger (A. tigrinum tigrinum)
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salamanders, as well as increased susceptibility of these species to Saprolegnia fungus (Lefcort
etal. 1997).
Few studies of sedimentation effects on amphibians have been conducted in wetlands. On one
hand, many species require soft sediments as hibernation sites, e.g., painted turtle (Chrysemys
picta) in sediments 0.50 to 0.95 meters thick in an Ontario pond (Taylor andNol 1989). On the
other hand, excessive sediments (when stirred) impair light penetration of the water column and
thus can inhibit growth of algae and especially submersed aquatic plants, which provide cover
and attachment sites for amphibian eggs.
7.7 Effects of Vegetation Removal
Gross classifications of vegetation form (e.g., Cowardin class) are only moderately useful for
describing habitat of some amphibians (Munger et al. 1998). In oligotrophic lakes of Quebec,
vegetation structure had a low capacity to predict the occurrence of bullfrogs (Rcma
catesbeiana\ mink frog (R. septentrionalis\ and green frog (R. clamitans melanota) (Courtois et
al. 1995), but was important to amphibians using a wetland constructed to treat mining wastes
(Lacki et al. 1992). In the Puget Sound Basin of Washington, surveys of 19 wetlands found no
statistically significant correlations between amphibian richness and vegetation form (Richter &
Azous 1995). Aquatic amphibians appear to require particular types of submersed plants as
deposition sites for their eggs. Plant stem diameter (e.g., stems <3 mm preferred by
northwestern salamander)(Richter and Roughgarden 1995) is apparently more important than
plant species (Richter 1997). Density of submerged plants also is important. A survey of 40
wetlands in the Puget Sound found more native species among wetlands containing dense
emergent vegetation (Adams and Bury 1998). Dense vegetation may help protect larval young
of native aquatic amphibians from larger predators.
Vegetation condition in surrounding buffer zones and watersheds can be at least as important to
amphibians as vegetation condition within wetlands. Amphibians were less common in
Pennsylvania stream corridors where vegetation had been removed from surrounding areas
(Croonquist 1990, Brooks & Croonquist 1990). After a clearcut of a bottomland forest wetland
in Louisiana, mole salamanders (Ambystoma talpoideum) in the vicinity had lower survival rates
(Raymond and Hardy 1991). Clearcutting in a southern Alabama bottomland hardwood wetland
resulted in only brief depression of species richness, but salamander diversity and abundance
were greatly reduced whereas frog and toad species increased (Clawson et al. 1997). In a
bottomland forest wetland in South Carolina, salamanders were much more common in mature
stands than in clearcut areas (Phelps & Lancia 1995). Uncut sites also had more gray treefrogs
(Hyla chrysoscelis), bronze frogs (Rana clamitans)., and box turtles (Terrapene Carolina).
Among 16 ephemeral ponds situated within tree plantations in New Brunswick, Canada, higher
densities and rates of recruitment of several amphibian species were found in the ponds situated
closer to natural forest (Waldick et al. 1999). The juxtapositioning of ponds and forest also was
found to be important to amphibian diversity in Shenandoah Mountains of Virginia (Mitchell et
al. 1997). Salamander populations in the Appalachians may require 50-70 years to recover from
clear-cutting (Petranka et al. 1993). Recolonization of wetlands denuded by the Mount St.
Helens volcano in Washington is projected to take up to 100 years due to slow dispersal rates of
amphibians and the distance from source habitats (Hawkins and Sedell 1990).
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In Maine forests, the abundance of frogs and salamanders (Rana sylvatica andAmbystoma
maculatum) declined along a gradient from mature forest-interior habitat (70-90 years old) to
recently clearcut habitat (2-11 years old) (deMaynadier and Hunter 1999). In one study in
northern California, streams flowing through uncut forests contained all 4 common amphibians
whereas only one of the streams flowing through cut forests contained all 4 amphibians. Only 2
uncut sites had fewer than 3 amphibians whereas 11 of the logged sites had no amphibians
(Welsh 1990). In Oregon, amphibian species richness was higher in streams flowing through
undisturbed forests than in streams flowing through logged forests (Corn and Bury 1989).
Undeveloped vegetated buffers of 30-95 meters have been suggested to help maintain diversity
of amphibians (Rudolph and Dickson 1990) and are at least equally important to turtles (Burke &
Gibbons 1995).
Effects on amphibians and reptiles of removing surrounding wetland vegetation are not always
negative. The affinity of many reptiles for warm microclimates led to an overall increase in
herptefauna diversity in clearcut plots in a South Carolina bottomland wetland landscape (Phelps
and Lancia 1993, Perison et al. 1997). Among 37 Michigan wetlands studied over 20 years, two-
thirds of the local species extinctions occurred in wetlands where forests had grown up in the
surrounding area during that period, perhaps shortening the annual duration of inundation of
understory vernal pools and cooling the substrate (Skelly et al. 1999).
7.8 Effects of Thermal Alteration
Water and air temperatures can have different effects on larval vs. adult amphibians.
Excessively warm temperatures can dessicate amphibians once pools dry up (Shoemaker etal.
1992). In winter, painted turtles (Chrysemyspicta) sometimes move about so they can maintain
a temperature of 4-6°C, but must restrict such movements if ice cover causes anoxia in pond
sediments (Taylor and Nol 1989). Some aquatic salamanders in the Pacific Northwest appear to
choose northern shores of ponds and wetlands for egg-laying, presumably due to greater solar
exposure and warmer microclimate (Richter 1997). Northern water snake (Nerodia sipedori) is a
wetland-dependent species that may be especially sensitive to temperature (Robertson and
Weatherhead 1992).
7.9 Effects of Dehydration/Inundation
Most amphibians require moist conditions and cannot tolerate prolonged dry periods.
Amphibians also can be extremely specific in their water depth requirements, especially for
oviposition (Miaud 1995). Desiccation of seasonal pools, especially when it occurs ahead of
normal seasonal schedules, can ruin breeding success of amphibians (Rowe andDunson 1993).
This is partly because many amphibian species disperse only short distances (Berven and
Grudzien 1990). Many amphibian species survive long-term droughts or floods by maintaining
populations scattered across a variety of wetlands of different depth and water permanence.
The availability of numerous scattered wetlands can serve as a "cushion" against effects of
localized drought. Indeed, some frog and toad species living in relatively intact landscapes seem
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mostly unaffected at a population level by significant periods of drought (Dodd 1995). In
contrast, when wetland alteration increases the distances between the remaining fishless wetlands
most suitable for amphibians, and when terrestrial vegetation along dispersal routes is replaced
by roads or other unsuitable habitats, amphibian populations recover slowly or not at all from
droughts they might otherwise survive (Pounds and Crump 1994). Some amphibians in the
Florida Everglades survive the dry season in limestone solution cavities beneath the land surface,
and may live up to several months in these waters. Increased extraction of this water for human
use has made many of these potential dry-season refuges unavailable to aquatic animals (Loftus
etal. 1992).
Both prolonged desiccation and extreme floods can also increase opportunities for invasion of
wetlands by exotic plant species. Consequently, patterns of vegetation typically become more
homogeneous, and suitability of amphibian habitat as well as prey abundance may decline
(Munn and Brusven 1991, Ludwa 1994). Currents associated with floods can hinder breeding
and egg deposition for many amphibian species that require still water (Lind et al. 1996, Richter
1997), but also distribute woody debris and coarse sediments that are important components of
amphibian and reptile habitat. Naturally-occurring floods in larger rivers create a
geomorphically diverse mosaic of floodplain wetlands that cumulatively can support a similarly
diverse array of amphibian and reptile species (Galat et al. 1998).
A virtual absence of water level variation also can be indirectly detrimental to some amphibians,
because without occasional drying of substrates, nutrient cycles in some wetlands stagnate,
vegetation patterns become more homogeneous, and suitability of habitat may decline.
Effects on Species Richness
In South Carolina wetlands, amphibian species richness was statistically correlated with seasonal
permanence (Snodgrass et al. 2000). In Indiana, a survey of 30 forested wetlands found the
greatest amphibian species richness in wetlands of intermediate permanency (Kolozsvary and
Swihart 1999), although another Indiana study found amphibian richness to be greatest when
wetlands were located near permanent water bodies (Brodman and Kilmurry 1998). The number
of amphibian species in wetlands of the Puget Sound Basin of Washington was related more to
water level fluctuations than to vegetation form, with lowest richness occurring when springtime
fluctuations exceeded 20 cm (Richter 1997). Amphibian richness among 12 temporary wetlands
in Florida was related more to the presence offish (Lepomis sp.) than to water source (Babbitt
and Tanner 2000). Amphibian breeding success appeared to depend largely on the timing of
inundation, not as much on its duration. Wetlands connected to fish bearing waters had fewer
amphibian species, a phenomenon that has been noted elsewhere and is attributable to significant
fish predation on amphibian eggs and larvae (Hecnar & M'Closkey 1997, 1998).
Effects on Species Composition
Although some amphibian species are adapted to short (less than 3 months) inundation of
wetlands, many require longer permanence to produce adequate offspring (Pfingsten and Downs
1989, Tyning 1990, Conant and Collins 1991). For example, in Michigan, the distribution of 3
of 14 species were significantly affected by pond permanence (Skelly and Meir 1997). Among
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depressional wetlands of the upper Atlantic Coastal Plain, sirens (Siren intermedia and S.
lacertina) and amphiumas (Amphiuma means) were mostly found in wetlands with longer
durations of inundation. Biological interactions between these species also was suspected of
limiting their geographic distribution (Snodgrass et al. 1999).
In northwestern Nevada, severe drought and accompanying high temperatures killed many
Pacific chorus frogs (Pseudocris regallis) (Weitzel and Panik 1993). During a 2-year drought in
the state of Washington, a local population of painted turtle (Chrysemys picta belli) suffered a
70% decline (Lindenman and Rabe 1990). This appeared to be due to both mortality and
emigration from the wetland. Growth also was suppressed, but recovered following improved
conditions. The average sizes of cohorts was not affected. However, in temporary forest pools
in Mississippi, only 47% of the amphibian cohorts inhabiting the pools appeared metamorphosed
before the pools dried out (Bonner et al. 1997). In central Pennsylvania, spotted salamanders
(Ambystoma maculatum) -- which are among the last species to breed in the spring — were noted
as being especially sensitive to drought (Rowe and Dunson 1993). In a Florida population of
striped newt (Notophthalmusperstriatus) larger individuals became proportionally more
dominant after a severe drought (Dodd 1993b). Also in Florida, natural hydroperiod alterations
did not influence the activity of swamp snakes (Seminatrixpygaea) in wetlands. Drought had
little direct effect on overland migration or body condition, and only caused snakes to leave or to
shorten the amount of time they spent within a wetland (Dodd 1993a). Water level drawdowns
conducted in the autumn for wetland management, flood control, or other reasons can cause high
mortality among juvenile overwintering turtles (due to freezing) if the drawdowns follow
abnormally high late-summer water levels that attracted turtles (Galat et al. 1998).
Relative dominance in a wetland of amphibian species that metamorphose quickly into a
terrestrial stage may indicate short duration of inundation. In contrast, relative dominance of
slow-growing species can indicate significant co-occurrence of predatory fish and invertebrates,
which are typically associated with long duration of inundation (Wellborn et al. 1996, Schneider
and Frost 1996, Schneider 1997). Some amphibians seem especially susceptible to predation as
they move from drying fringes to deeper waters of the same wetland, which often tend to have
less protective vegetation. For example, Pacific tree frog (Pseudacris regilla), red-legged frog
(Rana aurora), spotted frog (R. pretiosa), northern leopard frog (R pipiens), and western pond
turtle (Clemmys marmorata) appear to be more susceptible to predation by bullfrog (R
catesbeiana) as they move to more open water (Leonard et al. 1993, Hallond et al. 1995).
Extended floods, by inundating shoreline turtle nests or alligator nest mounds, can diminish or
ruin a season's recruitment of young (Kushlan and Jacobsen 1990, Tucker et al. 1997). In
Oregon, numbers of a mostly terrestrial salamander (ensatina, Ensatina eschscholtzii) were less
in frequently-flooded riparian habitats of red alder (McComb et al. 1993a, McComb etal.
1993b), second growth conifer (Gomez and Anthony 1996), and unmanaged Douglas fir (Aubry
and Hall 1991, Gilbert and Allwine 1991).
7.10 Effects of Habitat Fragmentation and Other Stressors
Amphibian and turtle diversity, measured at a local or regional scale, is often severely affected
by filling of— or intentional connecting of (with ditches) — small, somewhat isolated,
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temporarily or seasonally inundated wetlands. Such wetlands are often the first to be ignored by
state and federal regulatory programs, or altered in exchange for protection/restoration of larger
wetlands as part of mitigation banking agreements. Large wetlands do not necessarily support a
wider variety of amphibians, as indicated by surveys of 97 Ontario ponds (Hecnar and
M'Closkey 1996), 19 Washington wetlands (Richter and Azous 1995), and 22 South Carolina
wetlands (Snodgrass etal. 2000).
An analysis of wetland spatial data in Maine, using demographic models for amphibians and
turtles, indicated that cumulative loss of many small wetlands, by increasing the distance
between wetlands and reducing dispersal success of several species, would result in eventual
extinction of many species, especially turtle species (Gibbs 1993). A similar simulation based on
landscapes in the Atlantic coastal plain concluded that isolated wetlands as small as (0.2 hectare)
need to be protected, and spatial patterns of wetland distribution taken into account, if regional
biodiversity of amphibians is to be maintained (Semlitsch & Bodie 1998). Connecting isolated
wetlands with ditches often introduces predatory fish or carnivorous amphibians to somewhat
isolated amphibian populations, with consequent shifts in species composition and sometimes a
general decline in abundance of native species (Babbitt and Tanner 2000).
Loss of vegetated uplands that connect isolated wetlands can hinder an amphibian population's
ability to recover from drought (Pounds and Crump 1994), disease (Bradford 1991) low
reproductive rates (Sinsch 1992), and other wetland alterations (Dodd and Cade 1998).
Probability of occurrence of several amphibian species in Indiana was strongly associated with
proximity to wetlands (Kolozsvary & Swihart 1999). In another Indiana study, amphibian
richness was greater in wetlands located close to forested areas (Brodman and Kilmurry 1998).
A similar result was found in Washington (Richter & Azous 2000). In New Mexico, 3 wetland-
associated toad species fed extensively on carabid beetles in surrounding terrestrial habitats
(Anderson et al. 1999). A lack of suitable upland habitat adjoining some southern Illinois
wetlands was partly blamed for absence of some species (Burbrink et al. 1998). The width of the
riparian corridor connecting the wetlands appeared to be less important to supporting a diverse
amphibian community than local habitat heterogeneity and distances between source wetlands.
Changes in land cover leading to increased isolation of wetland breeding habitats from each
other, or from essential upland habitats, have been suggested as a cause of decline of many
amphibian species (Blaustein etal. 1994c, Sjogren-Gulve and Ray 1996), including mountain
yellow-legged frog (Rana muscosa) (Bradford, Tabatabai, and Graber 1993), Oregon spotted
frog (Ranapretiosa) (Orchard 1992, Azous & Richter 1995, McAllister and Leonard 1990,
1997) and western pond turtles (Clemmys marmorata) (Gray 1995). Among 21 Minnesota
glacial marshes, amphibian richness was lower in marshes that were more isolated (Lehtinen et
al. 1999). Among 37 Michigan wetlands studied over 20 years, the number of amphibian species
extinctions and colonizations was greater among the more isolated wetlands (Skelly et al. 1999).
However, among 19 Puget Sound Basin wetlands of Washington, amphibian richness was not
necessarily greater in wetlands that were less isolated from (closer to) other water bodies
(Richter & Azous 2000).
Even when they are connected by water, wetlands might be considered "isolated" when the
distances between them exceed the typical dispersal distance of amphibian and turtle species
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they otherwise are capable of supporting. Adults of 6 salamander species used habitats an
average of 125 meters from the ponded edge of a wetland in Missouri (Semlitsch 1998).
Dispersal distance for many wetland amphibians is <0.3 km (Gibbs 1993; Semlitsch 1998,
Semlitsch and Bodie 1998). Semlitsch (1998) recommended that natural land cover be protected
within an average distance of 164 m from the edge of wetlands to protect 95% of the
ambystomid salamander populations. Non-ambystomid salamanders (newts), frogs, and toads
may have greater requirements because of typically greater dispersal distances. Even when
habitat or water quality within or around a wetland diminishes, amphibians may still attempt to
breed there due to strong behavioral attachment to natal sites (Karns 1992).
Although more protective than developed land uses, it is uncertain whether vegetated upland
buffers and/or upland corridors that link wetlands are sufficient to protect some amphibians and
turtles from becoming locally extinct. Impacts to amphibians occur because developed land is
characterized not only by less habitat space and greater wetland isolation, but also by alteration
of wetland water quality/hydrology and an increased number of edges between forest and
openland, which some species avoid (Gibbs 1998). Among 21 Minnesota glacial marshes,
amphibian richness declined as the proportion of urban land cover increased at all spatial scales
(Lehtinen et al. 1999). Similar results were found in surveys of frogs or amphibians generally in
wetlands of the Seattle area (Azous and Richter 1995, Richter and Azous 1995), Connecticut
(Gibbs 1998), and Wisconsin and Iowa (Knutson et al. 1999). Specifically, land cover alteration
is believed to have negatively impacted spotted frogs (Ranapretiosa) in Washington (McAllister
and Leonard 1991) and red-legged frogs (R. aurora draytonii) in California (Hayes and Jennings
1988). Even outside of urban areas, other areas of relatively homogeneous land cover, such as
grazed rangeland and agricultural land, have been shown by some studies to support lower
amphibian abundance and/or richness (Mensing et al. 1998, Hecnar 1997, Bonin et al. 1997,
Delis etal. 1996, Bishop et al. 1999).
However, land cover alteration does not inevitably diminish richness at a local scale. In south-
central Florida, an area that was comprised of ditches, pastures, ponds, and orange groves
contained as many native reptile and amphibian species as more natural reference areas
(Meshaka 1997). In playa wetlands of Texas, frog occurrence did not appear to correlate with
land cover or the presence of irrigation pits (Anderson et al. 1999). In Wisconsin, amphibian
abundance was associated positively with both agricultural lands and with forested lands
(Knutson et al. 1999). Thus, the configuration of suitable habitat, which reflects the ability of
individuals to safely disperse, may be equally or more important than total habitat area. Effects
also may depend on the particular species, with some being more dependent on wetlands and
other natural environments.
Many of the impacts found in developed landscapes may be attributed to roads (Langton 1989,
Fahrig et al. 1995, Gibbs 1998). Vehicular traffic can lead to amphibian declines either through
direct mortality and increased exposure to predators (Ashley and Robinson 1996), or as
amphibians avoid crossing roads, thus reducing gene flow among populations as evidenced by
greater genetic distancing (Reh 1989). Amphibian species richness in 21 Minnesota glacial
marshes was less at all spatial scales having greater road density, in 2 ecoregions (Lehtinen et al.
1999). In Massachusetts, suburban highways affected amphibian populations more than 100 m
from the road (Forman et al. 2000).
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Constructed and restored wetlands have been surveyed for amphibians only occasionally. In one
instance, a constructed wetland being used to treat acid-mine drainage also supported more
amphibian species than nearby natural wetlands, and also had greater abundance of green frogs
(Rana clamitans) and pickerel frogs (Ranapalustris) as well as a variety of snakes attracted by
the large frog prey base and the availability of den sites amid mining rock debris. Apparently the
acidity did not reach toxic levels.
Several studies documented the avoidance of some breeding ponds that contained potential
predators. American toads avoided laying eggs in ponds that contained wood frog tadpoles,
which feed on the eggs and larvae of toads (Petranka et al. 1994). Wood frogs (Rana sylvatica\
spotted salamanders (Ambystoma maculatum), and Jefferson salamanders (A. jeffersonianum)
seem to be almost entirely dependent on fish-free wetlands for breeding (Rowe and Dunson
1993).
Introduced reptiles and amphibians have caused the decline of some native amphibians. The
bullfrog (Rana catesbeiana), a large anuran that regularly consumes smaller frogs, was
introduced to parts of the western U.S. decades ago and is suspected of contributing to the
decline of northern leopard frogs (R. pipiens) and red-legged frogs (R. aurora) (Panik and Barrett
1994, Lawler et al. 1999). However, other evidence suggests the distribution of native
amphibians among Washington wetlands is not linked tightly to bullfrog presence (Adams et al.
1998, Adams 1999), and predatory fish can be at least as important as predators. Introduced fish
eliminated populations of mountain yellow-legged frog (Rana muscosa) a century ago in many
lakes and streams (Bradford et al. 1991, Graber 1993, Knapp & Matthews 2000). However,
among 21 Minnesota glacial marshes, species richness was greater in wetlands that contained
fish and tiger salamander (Ambystoma tigrinum) (Lehtinen et al. 1999).
Native turtles possibly are being affected by introduced turtles that compete for habitat and food,
as well as spread diseases and parasites (Bury 1994). Decline of western toad (Bufo boreas) has
been hypothesized to be the result of increased incidence of the pathogenic fungus, Saprolegnia
Blaustein et al. (1994a). Another hypothesis is that the decline of both western toad and Cascade
frog (Rana cascadia) is attributable to ultraviolet-B radiation (Blaustein et al. 1994b). Field
experiments revealed that fungus reduced the hatching success of three Oregon anurans most
dramatically when the eggs were also exposed to elevated levels of ultraviolet-B radiation
(Kiesecker & Blaustein 1995).
Human recreation negatively affected populations of North American wood turtles (Clemmys
insculpta) in part of Connecticut (Garber and Burger 1995). Wood turtle populations declined
where habitat was opened to recreation (i.e., hiking and fishing), perhaps partly because of illegal
collection of individuals. Populations remained stable in habitats where recreation was
restricted. Legal harvest of 50% of the annual production of alligators in 2 Florida lakes did not
significantly affect distribution of age classes within the alligator population (Rice et al. 1999).
7.11 Wetland Monitoring
Spatial and Temporal Variability
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Choice of appropriate sample sizes depends on measured variation in the target taxa and metrics.
Such coefficients of variation are summarized from various amphibian studies elsewhere at:
http://www.mp2-pwrc.usgs.gov/ampCV/ampdb.cfm
At a landscape scale in natural landscapes, the distribution of many amphibians appears to be
influenced mainly by suitability of adjoining habitats, water persistence, predator presence, and
their interaction (Wellborn et al. 1996, Skelly and Meir 1997, Pfmgsten and Downs 1989,
Tyning 1990, Conant and Collins 1991). Wetlands with greater persistence of water tend to
support more animals that prey on amphibians, and often have fewer amphibian species.
Temporally, the populations of most amphibian and reptile species in natural landscapes
fluctuate greatly from year to year, sometimes with no obvious cause (Pechmann et al. 1991,
Dodd 1992, Stone et al. 1993, Cohn 1994). However, substantial turnover may occur in the
species composition of individual wetlands, even in the absence of strong human influence. For
example, in a study of 14 amphibian species in 37 Michigan wetlands, 2 surveys conducted
about 20 years apart recorded 40 colonizations and 34 extinctions, with little overall net change
in the breeding populations of most species (Skelly et al. 1999). This natural temporal variability
needs to be accounted for when attempting to interpret trends from monitoring data.
Techniques and Equipment
Equipment and methods used to sample amphibians and/or turtles in wetlands include, but are
not limited to:
• call surveys (see: http://www.im.nbs.gov/amphib/naampappl.html)
• timed searches (Croonquist and Brooks 1990, Petranka et al. 1993)
• dip net sweeps (Anderson et al. (1999)
• drift fences (Dodd 1991, Brenner et al. 1992, Buhlmann. 1999)
• pitfall traps (McComb et al. 1991, Mitchell et al. 1993)
• box traps (Babbit and Tanner 2000)
• funnel traps (Richter 1995)
• minnow traps (Kolozsvary and Swihart (1999)
• shelter boards (Grant et al. 1992)
Methods for designing amphibian studies, as well as sampling methods, equipment, and data
interpretation, are described by Heyer etal. (1994), Fellers & Freel (1995), and Olson et al.
(1997). Much of this information was also summarized in Adamus and Brandt (1990), by
EPA's amphibian assessment group (http://www.epa.gov/owow/wetlands/bawwg), and by the
USGS's NAAMP program: http://www.mpl-pwrc.usgs.gov/amphibs.html
Metrics for Assessing Impacts to Amphibian Communities
Few published studies have examined specific metrics applicable to using amphibians and
reptiles for monitoring wetland condition. Because most individual wetlands have few
amphibian and turtle species, richness and species composition are best employed as metrics of
ecological condition at landscape scales, e.g., to wetland "complexes," rather than to individual
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wetlands. Nonetheless, within individual wetlands, rates of deformities and total abundance of
tadpoles, turtles, and aquatic salamanders can often be used as indicators of wetland condition,
provided natural reference conditions have first been adequately measured.
7.14 Literature Cited
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Section 8: Birds
8.1 Use As Indicators
This section addresses birds that are closely associated with inland wetlands and riparian areas.
This includes waterfowl, wading birds, shorebirds, and many songbirds. For a general
discussion of the topic based on pre-1990 scientific information, and for discussion of
advantages and disadvantages of using birds as indicators of wetland integrity, readers should
refer to Adamus and Brandt (1990). A recent book by Weller (1999) provides a good overview
of wetland bird ecology and behavior.
In contrast to most other flora and fauna, there have been no recent publications demonstrating
use of observation-based bird IBIs (indices of biotic integrity) specifically in wetland or riparian
settings. However, several studies have contrasted bird communities in urban/developed settings
with similar habitats in rural/undeveloped settings, both at individual paired sites (e.g., Craig &
Barclay 1992, Dowd 1992) and at a regional scale that encompasses gradients of human
influence (e.g., Croonquist & Brooks 199, Blair 1996, Flather & Sauer 1996, Miller et al. 1997,
Galatowitsch et al. 1998, O'Connell et al. 1998, 2000, Whited et al. 2000, Cam et al. 2000). As
summarized by Adamus (2001), these studies have clearly supported the utility of employing
bird species composition - and wetland birds in particular - as an indicator of land cover
alteration, habitat fragmentation, and other human influences at multiple scales.
8.2 Effects of Enrichment/ Eutrophication/ Reduced Dissolved Oxygen
Eutrophication can indirectly affect wetland bird community composition by altering the
vegetation structure and availability of prey items. For example, fish production is generally
greater in Canadian lakes that have at least moderate nutrient levels, and distribution offish
among lakes largely determines local distributions of Common Loon (Kerekes 1990). However,
many waterfowl that feed on aquatic invertebrates avoid lakes with fish, because fish can
decimate populations of invertebrates most important to ducks (McNicol and Wayland 1992).
Moderately elevated nutrient levels also spur the growth of submersed macrophytes important as
food for ducks, as well as supporting more aquatic insects that are especially important as food
for ducklings and for aerial foragers like swallows. However, excessive nutrients cause algal
blooms that can kill fish, decimate macrophytes by blocking light, and reduce visibility to birds
of food items located under the water surface. Such a situation has been documented in the
Chesapeake Bay (Perry & Deller 1996). Excessive nitrates have been implicated in deaths of
some frogs, which are significant prey for many wetland birds (see Amphibians chapter).
Northern Shoveler and Eared Grebe were positively associated with phosphorus in a survey of
wetlands in interior British Columbia (Savard et al. 1994).
Waterbird abundance and biomass were positively correlated in 46 Florida lakes with levels of
phosphorus, nitrogen, and chlorophyll. There also was a positive correlation of waterbird
richness with phosphorus, after accounting for nutrients contributed to the lakes by the birds
themselves (Hoyer and Canfield 1994). Wetlands constructed for wastewater treatment are often
heavily used by waterbirds (Frederick and McGhee 1994). Surveys of 92 British Columbia
ponds reported that densities of most of the 17 breeding duck and grebe species were associated
positively with total dissolved nitrogen. Total dabbling duck density was correlated positively
with total dissolved nitrogen (Savard et al. 1994). Surveys of 837 river corridors in England and
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Scotland identified 23 (of 29) waterbird species whose breeding abundance was associated with
unspecified water quality variables (Rushton et al. 1994). British surveys of 109 estuaries also
identified trophic status as a factor possibly influencing species composition of shorebird
assemblages (Hill et al. 1993). A survey of 95 inland wetlands in Australia identified 15 (of 61)
waterbird species whose occurrence was somewhat related (negatively) to higher phosphorus
levels (Halse 1993).
The parasitic nematode, Eustrongylides ignotus, which has only been found in disturbed and
enriched wetlands (Spaulding and Forester 1993), negatively affects the health of adult wading
birds and the survival of nestlings (Spaulding et al. 1993).
8.3 Effects of Contaminant Toxicity
Several new studies examined effects of contaminants on individual bird species, but seldom on
entire bird communities within wetlands.
Several instances have been documented of wetland birds being directly poisoned by insecticides
applied at recommended rates (e.g., Flickinger et al. 1991 - parathion). In North America,
recent attention has focused on effects on birds of fenthion, diazinon, phorate, chlorfenapyr, and
chlorpyrifos. Many pesticides are more likely to affect birds by altering their habitat and foods
than by being directly toxic. Insecticide-related reductions in invertebrate foods of waterbirds
have been documented in prairie wetlands (Tome et al. 1990, 1991, McCarthy and Henry 1993,
Martin and Solomon 1990). In contrast, application of the non-chemical insecticide Bacillus
thuringiensis subsp. israelensis (Bti, applied as Vectobac-G granules) and methoprene had no
detectable effect on breeding birds in Minnesota (Hanowski et al. 1997). Herbicides have been
applied to wetlands to alter vegetation structure and species composition, with consequent shifts
in bird species composition (Solberg & Higgins 1993, Linz et al. 1997). Information on
pesticides in prairie wetlands was compiled by Facemire (1992).
Much research has continued to focus on the effects of selenium on waterfowl in western states
(e.g., Hoffman et al. 1996). Biogeochemical conditions favoring the release of selenium into
wetlands are found throughout the arid regions of the western states and threaten bird
communities in many wetlands that are part of the Pacific and Central Fly ways (Lemly et al.
1993, Stephens etal. 1992, Paveglio et al. 1992). Agricultural drainage, irrigation, and natural
waters can leach selenium from many western soils. Subsurface irrigation is the most
widespread and biologically important source of selenium toxicity for waterfowl, including the
waterfowl in six national refuges (Lemly et al. 1993, Naftz etal. 1993, Ohlendorf etal. 1990,
Feltz et al. 1991). However, Barn Swallows that aerially forage for insects over contaminated
areas apparently suffered no adverse reproductive effects (King et al. 1994). Selenium is often
accompanied by boron, which also is toxic to ducklings (Stanley et al. 1996).
Contamination offish-eating terns by selenium and mercury was documented by one study that
found significantly higher levels of these metals in breast feathers of birds in North Temperate
breeding areas than in feathers of birds just returning from South Temperate wintering grounds
(Burger et al. 1992). Concentrations of mercury in feathers of 92 southern Florida wading birds
indicated that body concentrations were probably high enough to interfere with reproductive
success (Sundlof et al. 1994, Beyer et al. 1997). Mercury concentrations vary with geographic
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locations but also with diet and age of the bird (Sundlof et al. 1994, Beyer et al. 1997). Species
that eat larger fish, and older birds, tend to have the highest Hg concentrations.
Breeding songbird richness near zinc- and copper-contaminated Montana wetlands was not
significantly less than in uncontaminated reference areas (Linder et al. 1994, Pascoe et al. 1994),
but reproductive success of most species was not measured. Another metal - lead (Pb) - has
been documented to accumulate in Tundra Swans feeding from wetland sediments contaminated
by mining and smelting waste (Beyer et al. 1998), reaching levels capable of killing birds (Blus
et al. 1991). Kendall et al. (1996) review the effects of Pb on raptors and upland, nongame
birds. Winter survival rates for immature Canvasbacks exposed to Pb in Louisiana were lower
than those not exposed to Pb. In all, 16% of the immature canvasbacks that were examined had
died from Pb they absorbed by ingesting lead shot in the water (Hohman et al. 1995). Other
evidence of bird toxicity from lead shot was reported by Havera et al. (1992), Hohman et al.
(1993), and Peters and Afton (1993).
Oil spills can, of course, severely impact wetland bird populations (e.g., Wiens et al. 1996,
Burger 1997). The effects of synthetic organic compounds on birds also are receiving increased
attention. Past exposure to organochlorine compounds, PCB, dioxin, and other contaminants was
documented in Black-crowned Night Herons by measurement of cytochrome P450 (Rattner et al.
1994, 1996, 1997). Using the same biomarker, chronic exposure to petroleum hydrocarbons in
oil field brines was documented in Western Sandpipers in Texas, but no acute effects were
apparent (Rattner et al. 1995). Planar halogenated hydrocarbons (PHH) have caused
embryotoxicity, congenital deformities, and poor hatching success in Forster's Terns (Tillitt et
al. 1993). PCBs, dibenzodioxins, and poly chlorinated dibenzofurans have been implicated as
contributors to the continued decline of Common and Forster's Terns in the Great Lakes,
whereas some populations of Double-crested Cormorants, Herring Gulls, and other colonial,
fish-eating bird populations appear to be increasing as contamination of their particular food
chains diminishes (Giesly et al. 1994). Detrimental reproductive effects have been documented
of organochlorines on Herring Gulls (Ewins et al. 1992), dioxins on Great Blue Herons (Hart et
al. 1991), dioxins and furans on Wood Ducks (White and Seginack 1994, 1995), PCB's in
American Kestrels, and petroleum in Mallards (Holmes and Cavannaugh 1990). Advantages and
disadvantages of using particular biomarkers of chemical exposure in birds are discussed by Fox
(1993).
8.4 Effects of Acidification
Acidification of wetlands affects birds primarily because it reduces calcium availability
(important for egg development), potentially increases toxic metal availability, and alters the
species composition and occasionally the abundance of aquatic insects, submersed plants, and
fish that are important foods for waterfowl.
Reduced availability of calcium-rich foods and the change in types of available food can
diminish egg shell thickness and generally reduce the reproductive success of waterbirds in
wetlands (Sparling 1990, 1991, Blancher andMcNicol 1991, St. Louis et al. 1990, Albers and
Camardese 1993) and in streams (e.g., dippers in England: Tyler and Ormerod 1992, Ormerod
and Tyler 1991). However, when acidification removes fish from wetlands the abundance of
insect prey can increase (McNicol et al. 1990, Blancher and McNicol 1991, Blancher et al. 1992)
as can selection of the most productive wetlands by hens with young broods (Parker et al. 1992,
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Merendino etal. 1992, Merendino & Dennis 1993, Merendino and Ankney 1994), and the
number of waterfowl broods (McNicol and Wayland 1992). Nonetheless, insect densities do not
necessarily increase the reproductive success of aerial-foraging insectivorous species, e.g., Tree
Swallows, which are potential vulnerable to chemical bioaccumulation (St. Louis etal. 1990, St.
Louis and Barlow 1993, Blancher and McNicol 1991, Froese et al. 1998). Overall, calcium
deficiency appears to affect birds in acidified wetlands to a greater degree than metal toxicity
(Albers and Camardese 1993). Breeding pairs of 15 waterfowl species are more abundant in
Ontario wetlands with > 40 ppm total alkalinity than in less alkaline wetlands (Dennis et al.
1989, Merendino et al. 1992). In British Columbia as well, densities of several breeding duck
species were greater in ponds with higher levels of conductivity and calcium (Savard et al.
1994).
8.5 Salinization
Highly saline or alkali conditions are detrimental to some invertebrate and plant foods used by
many duck species, and high salinity is directly toxic to — or impairs the growth of- young
ducklings (Clark & Nudds 1991, Moorman etal. 1991). Those sensitive waterbirds may visit
saline wetlands, but often only when fresher wetlands are available nearby (Lokemoen and
Woodward 1992, Woodin 1994, Adair et al. 1996). Breeding densities of most duck and grebe
species in interior British Columbia were greater in ponds with higher conductivity, but marsh-
nesting species were unaffected (Savard et al. 1994). In a survey of part of the Canadian prairie,
2% of the wetlands were found to be potentially too saline to support waterfowl reproduction
(Leighton and Wobeser 1994).
Nonetheless, a few waterbird species occur regularly at very high densities in alkali wetlands
during the breeding season and/or migration, e.g., American Avocet, Black Stilt, Snowy Plover,
phalaropes, Killdeer, Horned Grebe, Tundra Swan; White-rumped, Semipalmated, and Baird's
sandpipers (Eldridge and Krapu 1993, Earnst 1994, Jehl 1994, Kingsford and Porter 1994,
Savard et al. 1994, Oring & Reed 1997, Rubega & Johnson 1997, Warnock 1997). These
relatively salt-tolerant species also occur in less saline wetlands, but their abundance often is
greatest in hypersaline wetlands, and is related to sharp seasonal peaks in the abundance of brine
shrimp and other salt-tolerant invertebrates. They characteristically travel hundreds of miles,
sometimes on a daily or weekly basis, in order to exploit such invertebrate foods during the short
times when the food peaks (Haig et al. 1998).
8.6 Sedimentation/ Turbidity
Sedimentation can affect birds by altering habitat structure, killing submersed vegetation, or
altering the abundance or availability of prey items, (see Sections 4 and 5). However, in ponds
in the interior of British Columbia, densities of breeding dabbling ducks were correlated
positively with wetland turbidity (Savard et al. 1994).
8.7 Thermal Alteration
Thermal alteration can affect birds the greatest by preventing ice to form in some northern waters
and by altering the seasonal abundance of prey items. No new studies were found documenting
the direct effects of thermal alteration of wetlands on birds, but habitat changes that will occur as
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a result of global warming are a significant long-term concern (Poiani & Johnson 1992, Poiani et
al. 1996).
8.8 Dehydration/Inundation
Dehydration/Drought
Drainage and some types of hydrological manipulation of wetlands have been well-documented
as contributing causes in the decline of many wetland bird species (Batt et al. 1989, Bortner et al.
1991, David 1994, DeAngelis et al. 1997). In Manitoba, for example, wetland drainage has
decreased the availability of breeding and brood-rearing areas for waterfowl (Rotella and Ratti
1992). As wetlands are drained or converted to other land cover types, local densities of
wetlands decline and mean distances between individual wetlands increase. A cypress dome
rehydration project in central Florida precipitated the return of 16 wetland bird species to the
wetland (Weller 1995). In contrast, in the Florida Everglades, fewer bird species were found in
hydrologically impacted sites than in reference sites (Gawlik and Rocque 1998).
Drought conditions also expose duck nests to greater predation as a result of diminished
vegetation vigor and density, and creation of access points to islands that formerly were
inaccessible (Hallock & Hallock 1993, Jobin & Pieman 1997). By reducing the number and
perhaps the variety of wetlands and the vegetation communities they contain, sustained regional
drought or widespread draw down of water tables diminishes avian richness, bird density, and
breeding success in many individual wetlands and wetland complexes (Higgins et al. 1992,
Bethke & Nudds 1993, Bancroft et al. 1994, Greenwood et al. 1995, Dobkin et al. 1998). As a
result of the hydrological impacts in the Everglades, wood storks begin breeding later in the
season, and in most years a late start is followed by nest failures (Ogden 1994). Anthropogenic
and natural reversals in wetland drying also have resulted in failed nesting attempts in wood
stork colonies in South Florida (Bancroft et al. 1994)
Complexes of wetlands — where permanently, semipermanently, seasonally, and temporarily
inundated sites are close together — are important for many waterbird species, especially in the
prairie region (e.g., Rotella & Ratti 1992a,b; Cowardin et al. 1995) and in other regions where
precipitation varies greatly from year to year (Meyers & Odum 1991, Fleming et al. 1994).
Permanent water need not directly adjoin nesting habitat in order for larger duck species to breed
successfully, and in fact sites farther from water can be more productive for Mallards, Northern
Pintail, and Northern Shovelers, so long as vegetative cover is adequate (Kantrud 1993).
Proximity of nest cover to permanent water may be more important to Green-winged Teal and
American Wigeon. In Maine, richness of aquatic bird communities in individual wetlands was
correlated with local wetland density although not with distance to the nearest wetland (Gibbs et
al. 1991).
Wetlands that are inundated for only brief periods each year, and that typically lack surface
connections to other water bodies (i.e., many "isolated" wetlands), are relied on almost
exclusively by several bird species, especially many shorebird species. Many such areas are
former wetlands that have been at least partly converted to rice fields, sod farms, soybean fields,
or other agricultural activities that leave substantial areas of bare, fine-particled soil. During wet
years in some regions, these wetlands also can be the only ones shallow enough to provide
acceptable foraging habitat for many waterbirds, and they provide essential habitat for shorebirds
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most years (Nesbitt and Williams 1990, Eldridge 1992, Bishop 1992, Shuford et al. 1998, Twedt
& Nelms 1999). In addition, isolated wetlands that are inundated only seasonally provide (a)
vital feeding and stopover habitat for migratory waterfowl during early spring in northern
regions, when many deeper wetlands remain frozen, and (b) roosting areas during high tide,
when located near tidal waters (Lovvorn & Baldwin 1996, Rottenborn 1996, Burger et al. 1997).
Waterbirds that use seasonal wetlands are highly mobile and adapted to exploiting "on short
notice" the brief seasonal peaks of foods in these areas, provided that hydrologically diverse
complexes of wetlands are maintained at a regional scale (Skagen & Knopf 1993, 1994, Farmer
& Parent 1997, Oring & Reed 1997, Robinson & Warnock 1997, Warnock 1997, Haig et al.
1998, Laubhan & Gammonley 1999). Approximately 68% of the breeding bird species that
utilize wetlands in the Colorado Plateau use wetlands that are inundated only briefly each year
(Adamus 1993).
Inundation
In the south-central United States, wintering waterbird richness, abundance, and projected
carrying capacity were found to be greatest on playa wetlands that were repeatedly drawn down
or irrigated to a shallow depth during early April, late June, and early August, and then flooded
for the winter in November. However, by initiating flooding of these playas in September, bird
use occurred over a longer seasonal period (Haukos & Smith 1993, Anderson & Smith 1999).
Pre-irrigation of former wetlands and fields, followed by winter flooding, was found to be
effective for supporting greater numbers of wintering Northern Pintail in California's Central
Valley as well (Barnum & Euliss 1991). Deeper drainwater evaporation ponds are also
important to waterfowl wintering in the Central Valley (Euliss et al. 1991). Pre-irrigation tends
to flush salts from underlying soils. Pre-irrigated fields, when flooded in winter, provide
expansive areas where ducks are relatively safe from predation. However, in the southern
Mississippi Alluvial Valley, larger concentrations of wintering shorebirds were found on
permanently-inundated wetlands whose water levels were drawn down, than on fields that were
flooded only during winter (Twedt et al. 1998).
Bottomland hardwood wetlands are a vital resource for many wintering birds (Zeller & Collazo
1995). Bird richness and abundance in a bottomland hardwood floodplain, both in winter and
summer, were not significantly associated with relative wetness or flooding duration (Wakeley &
Roberts 1994). Flooding of bottomland wetlands displaced wintering American Woodcock and
might have contributed to increased mortality (Krementz et al. 1994). The breeding success of
large wading bird species nesting near Lake Okeechobee, Florida, varied between years
depending on annual drought or flood conditions. The particular response was species-specific
(Smith and Collopy 1995). Prolonged hydroperiods may be necessary for the development of
populations of large fish (Fleming et al. 1994) that are selected for by Wood Storks, Great
Egrets, and Great Blue Herons. Avian richness in restored prairie wetlands was mostly greater
during a wet than a dry year (Hemesath & Dinsmore 1993).
Wetlands created by beaver contain significantly more waterbird species than inactive beaver
sites or potential beaver sites of the same size (Medin and Clary 1990, Grover and Baldassarre
1995). Beaver affect the species richness by impounding water and creating a diverse mosaic of
emergent vegetation, flooded dead trees, and open water. Beaver ponds comprising only 25% of
the wetland area in Ontario were used disproportionately by dabbling ducks (Merendino et al.
1995). In an examination of 70 beaver ponds, Grover and Baldassare (1995) determined that
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active beaver ponds had more species of birds than inactive sites. Active sites had more open
water, dead standing trees, surface water, and flooded emergents than inactive sites. Female
mallards in the St. Lawrence River Valley spent most of their time breeding in forested-live
wetlands (40%) and postbreeding time in forested-dead wetlands (35%) (Losito et al. 1995).
The diking and filling above the average high water level of coastal wetlands on the Great Lakes
has rendered much former habitat almost useless for nesting waterbirds (Prince et al. 1992).
Construction of reservoirs also impacts birds by eliminating many wetlands, while creating
others with less stable water levels (Nilsson & Dynesius 1994). Associated changes in river
morphology influence species composition of wintering waterfowl (Johnson et al. 1996). In the
Florida Everglades, reproductive success of an endangered subspecies — Cape Sable Seaside
Sparrow — was less during a period when water levels were kept at unusually high levels,
potentially flooding nests, increasing nest predation, and changing vegetation composition and
structure (Nott et al. 1998).
8.9 Effects of Vegetation Removal, Habitat Fragmentation
Much ornithological research during the past decade has focused on impacts of loss of natural
vegetative cover on songbird reproductive success. Most of these studies have focused on
upland forests, but an increasing number have examined grasslands, riparian systems, and
floodplains. In any event, most findings from upland landscapes are probably transferable to
wetlands with similar vegetation structure. Many studies of upland forest tracts of various sizes
(areas) situated in agricultural or urban landscapes continue to document a striking decrease (in
the smaller, partly fragmented tracts) in the occurrence, abundance, and/or reproductive success
of several neotropical migrant songbird species (e.g., Andren 1994, Askins 1995, Donovan et al.
1995, 1997, Friesen et al. 1995, Mclntyre 1995, Fauth et al. 2000). This is widely attributed to
disproportionate vulnerability of these species to predation and parasitism in smaller tracts
(Robinson et al. 1995), although some evidence from the floodplain forests of the Upper
Mississippi River suggests that predation on nests may actually be less in smaller isolated
patches (Knutson et al. 2000). Wooded patches smaller than about 100 hectares, and especially
those smaller than 16 hectares, generally do not support the full set of songbird species present in
larger wooded patches (Blake & Karr 1984, Robbins et al. 1989). Some breeding species return
annually to their natal sites even after their former habitat has been severely altered (Villard et al.
1995). Use of bird species assemblages as indicators of vegetation disturbance in non-wetland
environments is discussed by Hutto (1998).
From several studies it is now evident that what has been documented in uplands — that larger
connected patches of natural habitat, relatively unfragmented by roads, support more native bird
species than smaller, fragmented patches — is also true in wetland and riparian areas. These
wetland studies include: (a) a survey of 30 Ontario wetlands ranging in size from 13 to 1500
hectares (Findlay & Houlahan 1997), (b) an analysis of data from 18 forested wetlands in
Maryland (Schroeder 1996), (c) a survey of wetland and riparian habitat on 158 New England
lakes (Allen & O'Connor), (d) a survey of 40 Minnesota wetlands (Whited et al. 2000), and (e) a
survey of urban riparian areas in central California (Rottenborn 1999). Riparian areas and
wooded wetlands — the habitats that often are often most-used as breeding habitat by neotropical
migrant birds (Gates and Giffen 1991, Flather & Sauer 1996) - also happen to be characterized
by an unusually large variety and density of nest predators and parasites. Neotropical migrant
birds include dozens of species of warblers, vireos, tanagers, and other long-distance migrants.
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As a group these comprise a large portion of the breeding avifauna in most regions, and are
declining throughout developing regions in North America.
In addition to being sensitive to variation in land cover at a landscape scale, bird species
composition and density are very sensitive to structure and age of vegetation within a particular
site (e.g., Hanowski & Niemi 1990, Craig & Barclay 1992, Edwards & Otis 1999, Hanowski et
al. 1999). Bird richness and community structure have been compared among wetlands having
different vegetation cover types (e.g., Gibbs et al. 1991, Craig & Barclay 1992, Adamus 1992),
but comparisons can be confounded by differences among cover types with regard to bird species
detectability (Morrison et al. 1998). In Florida, riparian forests along blackwater rivers had a
somewhat different species composition than riparian forests along spring-fed rivers. Spring-fed
sites had more species, possibly because of their more complex vertical structure and greater
canopy closure (Leonard 1994). Neotropical migrants are especially sensitivity to the structural
complexity of habitat in bottomland hardwood forests (Pashley & Barrow 1993). Bird species
also differed among Colorado riparian areas with different vegetative structure (Finch 1991) and
in New Mexico riparian stands of different ages (Farley et al. 1994). Vegetation structure clearly
influenced species composition in central Iowa, where 48 breeding species were observed along
grassed waterways, but only 14 in surrounding crop fields. Breeding bird densities in the 44
grassed waterways were more than 3 times greater than in crop fields. Bird use peaked in July,
suggesting that grassed waterways should not be mowed until the end of August or early
September (Bryan and Best 1991).
Vegetation composition of islands located in lakes and wetlands also determines bird species that
nest there. Low shrubs, tall grasses, and dense herbaceous plants provide nesting cover for
Canada Goose, Mallard, Gadwall, and other ducks. Mallard, Blue-winged Teals, and Gadwall in
a North Dakota marsh had the highest nesting density and success on islands that were 335-1,085
m offshore, surrounded by 150-200 m of open water, and contained tall, dense, brush or forb
cover (Williams and Crawford 1989). Other species that prefer dense vegetation in wetlands
include Northern Harrier, Short-eared Owl, and Ring-necked Pheasant (Homan et al. 1993). One
North Dakota study that used herbicides to reduce vegetation cover found a reduction in densities
of Marsh Wren, Red-winged Blackbird, and Common Yellowthroat, up to two years after
application (Linz et al. 1993, Blixt et al. 1993). A Minnesota study found no positive correlation
between cover ratio (the ratio of open water to emergent vegetation) and numbers of Yellow-
headed Blackbird, Song Sparrow, or Sora (Olson 1992). However, avian richness in prairie
wetlands cannot always be predicted by vegetation structural diversity (Olson 1992).
The conversion of forested wetlands to emergent and open water wetlands can alter species
composition and richness of breeding birds both on-site and locally. For example, 53% of the
bird species that formerly used forested wetlands no longer occur regularly where such forests
have been mined and converted to emergent wetlands (Doherty 2000). A sustainable supply of
standing dead trees (snags) also is important in forested wetlands (Sedgwick & Knopf 1990).
Vegetated wetlands in Ontario were used to a greater degree than open water areas by Mallard,
American Black Duck, Wood Duck, Blue-winged Teal, and many other dabbling duck species
(Merendino et al. 1993, Merendino and Ankney 1994). Similarly, Connecticut tidal marshes
with a larger ratio of vegetation to open water contained more breeding bird species, as expected
(Craig & Beal 1992). However, most wading bird species avoid dense stands of vegetation
(Bancroft et al. 1994, Hoffman et al. 1994, Smith et al. 1995). Across a size range of 0.01 to 1.3
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hectares, shallow ponds created in tidal salt marshes for mosquito control were more attractive to
migrating waterfowl when they were smaller than 0.02 hectares or larger than 0.08 hectares
(Erwin et al. 1991). Migrating shorebirds were most attracted to ponds larger than 0.10 hectares
In most instances, achievement of successful wetland restoration is clearly tied to the recovery of
vegetation structure (Brown & Smith 1998, Zedler 1993, 1998) and the maintenance of
complexes of suitable wetlands in an appropriate landscape context (Morrison et al. 1994).
Thus, birds have sometimes been used to monitor progress in the condition of wetlands
following restoration (e.g., Sewell and Higgins 1991, Weller 1995). Breeding bird communities
of natural prairie potholes have been found to be more diverse than those of recently restored
wetlands (Delphey & Dinsmore 1993, VanRees-Siewert & Dinsmore 1996), although numbers
specifically of breeding ducks may not differ (Delphey & Dinsmore 1993). The lack of well-
developed vegetation zones that typify natural potholes likely led to the lower numbers or less
frequent encounters of American Goldfinch, Virginia Rail, Sora, Least Bittern, American Bittern,
Common Yellowthroat, Swamp Sparrow, and Red-winged Blackbird in recently restored
wetlands (Delphey and Dinsmore 1993). The development of a diverse community of
submersed aquatic plants may contribute to recovery of avian communities in some restored and
constructed wetlands (Weller et al. 1991, Leschisin et al. 1992, Mulyani & DuBowy 1993). In
peat-mined Canadian bogs that remained less vegetated than unmined bogs even 20 years after
abandonment, at least 10 species were less common in the bogs that had been mined, especially
using vacuum techniques (Desrochers et al. 1998).
Even when in pristine condition, wetlands are not necessarily covered completely with
vegetation at all times, and very dense stands of vegetation are unsuitable for several species
(Olson 1992, McMurl et al. 1993, Hemesath & Dinsmore 1993, Blixt et al. 1993). For example,
migrating shorebirds, as well as American Robin and Grasshopper Sparrow, were observed in
newly restored wetlands more frequently than in natural wetlands, probably due to their aversion
of dense vegetation (Delphey & Dinsmore 1993). As dense stands of vegetation are thinned, the
diversity of bird species using a wetland typically increases or remains stable (Blixt et al. 1993),
especially if open water begins to occupy spaces cleared in the vegetation. Small floating mats
of dead herbaceous vegetation, interspersed with open water, are important to some waterbirds
(Linz et al. 1997). In the Dakotas and Montana, islands in wetlands also had greater nest
densities and greater nest success than surrounding uplands (Lokemoen and Woodward 1992).
Logically, the greatest differences between bird composition in riparian/wetland habitats vs.
adjoining upland habitats have been found where the adjoining habitats are most dissimilar in
terms of vegetation structure (Strong & Bock 1990, Gates and Griffen 1991, Hooper 1991,
McGarigal & McComb 1992, Murray & Stauffer 1995, Karriker 1996). This is true in
agricultural as well as wooded landscapes. For example, in Saskatchewan wetlands, more birds
were found in natural wetlands and wetlands surrounded by organic farms, than in wetlands
surrounded by conventional farms or in wetlands surrounded by farms that minimized tillage but
still used chemicals (Shutler et al. 2000).
Even when vegetative structure of wooded sites does not change between years, species
composition within the sites may change if land cover in the surrounding landscape is altered
(Triquet et al. 1990, Richter & Azous 2001). Forested South Carolina wetlands surrounded by
pine woods supported several area-sensitive breeding species at the expense of species that
prefer field edges (Kilgo et al. 1997). Forested riparian areas surrounded by intact forests had
somewhat different bird species than those that were not, in Kentucky (Triquet et al. 1990),
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Pennsylvania (Croonquist & Brooks 1993), Vermont (Meiklejohn & Hughes 1999), and Idaho
(Saab 1999). Riparian strips and wooded wetlands have a more unique avifauna in regions with
homogeneous land cover, whether the homogeneity is represented by forests nearly unbroken by
logging and roads (LaRue et al. 1995, Anthony et al. 1996) or by agricultural land unbroken by
hedgerows and other patches of natural vegetation. Nonetheless, riparian strips usually do not
provide the same quality habitat or support the same richness and abundance of species as large,
wide tracts of lowland forest (Johnson & Brown 1990, Whitaker & Montevecchi 1999).
Several studies have attempted to identify (a) types and configuarations of edges between
habitats that minimize predation of bird nests, (b) minimal and optimal widths of riparian cover
necessary to maintain songbird richness at multiple scales, and (c) indicator species or bird
community metrics (response variables) that statistically are most sensitive to particular types of
human influences, at particular scales. In bottomland hardwood forested wetlands, predation on
bird nests is greater at forest-cropland edges than at forest-river edges or at the edge between
forest and natural levees (Saracco & Collazo 1999). Studies of buffer width basically have
concluded "the wider the better" if the aim is to protect the largest possible component of a
region's avifauna. However, in agricultural landscapes, the number of individual breeding birds
per unit area in wider corridors is not necessarily greater — only the number of species (Rich et
al. 1994, Darveau et al. 1995, Thurmond et al. 1995, Kilgo et al. 1998, Meiklejohn & Hughes
1999, Whitaker & Montevecchi 1999). The following table summarizes recent North American
research on bird response to riparian and wetland buffer widths.
Buffer or corridor width
(includes both stream
banks)
>60m
20m
411ft
164ft
14ft
40m
>500m
> 100m
>100m
175ft
Function
to sustain forest-dwelling
birds
to support several
ubiquitous species but not
most forest-dwelling species
to support full complement
of species
to support many sensitive
species
to maintain portions of the
bird community in disturbed
areas
to maintain canopy-
sensitive species in forested
headwater areas
to maintain the complete
avian community
characteristic of South
Carolina bottomland forests
to maintain a probability of
occurrence of at least 50%
of most breeding species
to maintain breeding
neotropical migrant species
did not support forest-
interior neotropical migrant
Habitat
coniferous
riparian forest
(balsam fir)
coniferous
riparian forest
(balsam fir)
riparian &
wetland
riparian &
wetland
riparian &
wetland
coniferous &
mixed
hardwoods
hardwoods
hardwoods
hardwoods
Location
Quebec
Quebec
Pennsylvania
Pennsylvania
Pennsylvania
Oregon
Maryland
Georgia
(Altamaha
River)
Researcher (s)
Darveau et al.
(1995)
Darveau et al.
(1995)
Croonquist and
Brooks 1993
Croonquist &
Brooks (1993)
Croonquist &
Brooks (1993)
Hagar(1999)
Kilgo etal. (1998)
Keller etal. (1993)
Hodges &
Krementz (1996)
Thurmond et al.
(1995)
-------
30-50 m
75-175 m
>60m
species
to maintain bird richness at
a landscape scale
to include 90% of species
breeding in a region's
riparian habitats
to maintain wintering
populations of Hooded
Warbler and Acadian
Flycatcher
hardwoods
mixed
urban riparian
eastern Texas
Vermont
Florida
Dickson et al.
(1995)
Spackman &
Hughes (1995)
Leonard (1994)
On a related note, a few studies have examined the apparent reluctance of forest birds to cross
areas of non-forest habitat, i.e., "gaps" that lack corridors (Machtans et al. 1996). During the
post-fledgling period, several songbird species in Quebec were only half as likely to cross a 50-m
wide field as cross a 50-m wide wooded area (Desrochers & Hannon 1997). Wooded routes
were preferred even when 3 times longer than shortcuts through openlands. Gaps narrower than
30 m did not function as barriers to movements of most songbirds. During the breeding period,
species that characteristically breed in forest interiors were least likely to cross gaps wider than
25 m (Rail et al. 1997). Similar results were found in the same region among wintering
songbirds (St. Clair et al. 1998). In Georgia, at least 5 breeding species avoided crossing the
100-m wide Altamaha River (Hodges & Krementz 1996). In North Dakota, post-fledgling birds
were found to move between wooded shelterbelts more often when the shelterbelts were
connected by wooded corridors than when structurally similar shelterbelts with the same
separation distance and area were not connected with corridors (Haas 1995). After clearcutting
of an Alberta forest, juvenile birds moved along the remaining riparian corridors (Machtans et al.
1996). Connectivity of natural cover at a landscape scale was a statistically important predictor
of bird species composition amid agricultural regions of Minnesota (Whited et al. 2000).
In contrast, in urban and desert landscapes of Arizona, where many bird species breed almost
exclusively in riparian areas (Germaine et al. 1998), the importance to migrant birds of wooded
connectivity (i.e., lack of gaps) was found to be less than the influence of total wooded acreage,
regardless of whether the wooded acreage is situated along a stream or as an oasis (Skagen et al.
1998). Although migrating songbirds used riparian habitats extensively, they did not appear to
be using riparian corridors as travel lanes during migration. Also, in more temperate forested
landscapes of the western United States, no clear evidence has been found linking forest
fragmentation to reduced avian abundance or productivity (Schieck et al. 1995, Tewksbury et al.
1998).
Effects of habitat fragmentation on songbirds are not limited to wooded habitats. In grassland
landscapes of Nebraska, herbaceous wetlands that are narrow (e.g., perimeter-area ratio <0.01)
support fewer breeding species, as do herbaceous wetlands smaller than about 30 hectares
(Helzer and Jelinski 1999). Narrow wetlands tended to have fewer individuals of Red-winged
Blackbird, Dickcissal, Upland Sandpiper, Western Meadowlark, Grasshopper Sparrow, and
Bobolink, and the latter 4 species also were less likely to use small wetlands. Areas more than 1
km from roads in Massachusetts were avoided by grassland birds (Forman et al. 2000). Wetland
area also was a significant predictor of breeding bird richness in herbaceous tidal marshes of
Connecticut (Craig & Beal 1992) and inland wetlands of Maine (Gibbs et al. 1991). Nest
success of some Missouri species also was found to be less in smaller grassland patches (Winter
& Faaborg 1999). In prairie pothole wetlands of eastern South Dakota, species that are most
vulnerable to loss of small wetlands are very mobile species that exploit resources over a broad
-------
region (Naugle et al. 2000). The excessive removal of vegetation by livestock — regardless of
the size, connectivity, and shape of wetland in which it occurs — can reduce onsite avian
diversity and reproductive success of some species (Anderson 1993, Ohmart 1994, Gilbert et al.
1996, Weller 1996).
Predation of bird nests in grassland and other open habitats has been shown to be greater (a)
close to edges with woods or plowed fields (Johnson & Temple 1990, Burger et al. 1994,
Pasitschniak-Arts & Messier 1995, 1996, Bellinger and Peak 1995), (b) in smaller patches of
natural grassland than in large patches (Kantrud 1993, Burger et al. 1994), (c) where human trails
go through sparse cover (Olson & Rohwer 1998, Miller et al. 1998), (d) in grazed areas (Gilbert
et al. 1996), and (e) where dikes or trails built on fill within a wetland make it easier for
predators to gain access (Peterson & Cooper 1991). Predation also may be less in large wetlands
because they are more likely to contain stretches of water too wide and deep for some
mammalian predators to cross (Pieman et al. 1993, Pieman & Schriml 1994, Esler & Grand
1993). Documented high rates of nest parasitism and predation are believed to contribute
significantly to a long-term decline of Song Sparrow in wetlands of coastal British Columbia
and possibly elsewhere (Smith & Arcese 1996, Rogers et al. 1997, Larison et al. 1998), as well
as to a decline of Willow Flycatcher in some western riparian areas (Harris 1991, Brown 1994).
Suburban ditches and hayfields near Boulder, Colorado, were repeatedly a sink rather than a
source of nesting Red-winged Blackbirds, due mainly to nest predation rather than cowbird
parasitism (Vierling 2000). In contrast, natural habitats (wetlands and tallgrass prairie) with
fewer nearby buildings were a source. Parasitism of Red-winged Blackbird nests by Brown-
headed Cowbird in Iowa was less at restored wetlands probably because of the lack of mature
trees from which to perch and search for nests (Delphey & Dinsmore 1993).
In southeastern regions where bottomland hardwood forests were once common, their removal
and fragmentation at a large scale has been associated with local or regional reduction in
abundance of the following breeding species (Burdick et al. 1989, Smith & Schaefer 1992,
Thurmond et al. 1995, Kilgo et al. 1998, Dickson et al. 1995):
Mississippi and Swallow-tailed Kites; Red-shouldered Hawk, Barred Owl, Ruby-throated
Hummingbird; Pileated, Red-bellied, and Downy Woodpecker; White-breasted Nuthatch,
Great-crested Flycatcher, Yellow-throated Vireo, Prothonotary Warbler, Northern Parula,
Swainson's Warbler, American Redstart, Black-and-White Warbler, Summer Tanager
Species that prefer thickets, such as White-eyed Vireo, Hooded Warbler, and Northern Cardinal,
may increase after logging. Species that are more prevalent in young stands (20-30 years old) -
including Acadian Flycatcher, Red-eyed Vireo, and Blue-gray Gnatcatcher (Wigley and Roberts
1994, Mitchell 1989, Kilgo et al. 1998) - also seem to tolerate narrower riparian strips. Wood
Thrush and Louisiana Waterthrush often occur in narrow wooded riparian areas in South
Carolina (Kilgo et al. 1998). However, in eastern Texas landscapes and perhaps elsewhere, some
of these species have been found exclusively or mostly in the widest riparian zones (Dickson et
al. 1995).
In northern regions, breeding species that appear to be less common in narrow than in wide
riparian strips or woodlots include the following (from Johnson & Brown 1990, Darveau et al.
1995, Spackman & Hughes 1995, Meiklejohn & Hughes 1999, Whitaker & Montevecchi 1999):
-------
Pileated Woodpecker, Yellow-bellied Flycatcher, Red-breasted Nuthatch, Veery;
Swainson's and Hermit Thrushes, Golden-crowned Kinglet, Ovenbird; and Blackpoll,
Black-throated Green, Bay-breasted, Blackburnian, and Cape May Warbler
Severe grazing, mowing, fire, or herbicide application at inappropriate times is detrimental to
many waterbird species and recovery of bird diversity in grazed areas may take several years
(Schultz & Leininger 1991, Johnson et al. 1991, Higgins et al. 1992, Gilbert et al. 1996, Dobkin
et al. 1998, Warkentin & Reed 1999). However, some studies (Clary & Medin 1993, Medin &
Clary 1990, 1991) found no statistically significant decline in overall avian richness and/or
abundance associated with grazing, or found reduced nest density but not reduced nest success
(Bowen & Kruse 1993). Effects depend on the grazing regime, the wetland plant community that is
being grazed, and other factors (Barker et al. 1990, Sanders & Edge 1998). In western rangelands,
wetland and riparian species that appear to be most sensitive to the immediate effects of grazing
include: Wilson's Phalarope, Willow Flycatcher, Yellow-breasted Chat, Yellow Warbler, Song
Sparrow, Savannah Sparrow, Spotted Towhee, and Red-winged Blackbird (Bock et al. 1993,
Dobkin et al. 1998, Sanders & Edge 1998). Even the reduction of vegetative cover by native
deer can adversely affect nesting songbirds; deer densities of 8-15 deer per km2 (about 20-40 per
mi2 ) were found to be the least at which widespread effects on songbird richness and abundance
were noted in Pennsylvania (DeCalesta 1994).
8.10 Disturbance from Human Visitation and Other Influences
Frequent visitation of wetlands by boaters and other recreationists can adversely some
waterbirds, especially near nesting colonies (Dahlgren & Korschgen 1992, Erwin et al. 1993,
Klein 1993, Knight & Gutzwiller 1995, Klein et al. 1995, Rogers & Smith 1997). Human
intrusion can disrupt bird feeding patterns (Skagen et al. 1991), reduce bird song which is vital to
reproductive success of most species (Gutzwiller et al. 1994), and cause at least temporary shifts
in bird community richness and abundance (Riffell et al. 1996). Powerlines and transmission
towers are also a source of mortality (Bevanger 1998).
Sport fishing has been suggested as a possible cause of a shift in the fish population from larger
predatory fish to smaller fish in a Canadian lake, and the resultant shift in avian species
composition (Gerrard et al. 1993). There were associated large increases in the abundance of
Common Loon, American White Pelican, Osprey, Great Blue Heron, and Herring, Ring-billed,
and Bonaparte's Gulls. No changes were seen in the abundance of mergansers, Bald Eagle, or
Common Tern.
Introduction of non-native fish, invertebrates, and plants also can affect wetland birds. For
example, large herbivorous fish such as carp compete directly with birds for submersed aquatic
plants, and also reduce water clarity (Bouffard & Hanson 1997). The introduced Zebra Mussel
provides abundant food for some diving duck species, but indirect effects on food chains are less
well known (Custer & Custer 1996, Hamilton & Ankney 1994). The introduced waterweed,
Hydrilla, also has apparently benefitted some species but its effects on others are unknown
(Esler 1990). Although purple loosestrife (Lythrum salicaria) has commonly been considered to
have few wildlife values, American Goldfinches nest in this habitat in New York wetlands
(Kiviat 1996).
-------
Global climate change has enormous potential to influence wetland birds, by influencing the
stressors described in the rest of this section. Potential effects are discussed by Larson (1993).
8.11 Wetland Monitoring
Spatial and Temporal Variation
Choice of appropriate sample sizes depends on measured variation in the target taxa and metrics.
Such coefficients of variation were calculated from previous avian studies in prairie pothole
wetlands (Adamus 1996), and are summarized from various studies elsewhere at:
http://www.im.nbs.gov/powcase/powvariation.html
In a central Maine bog, breeding birds appeared to select slightly different microhabitats during
consecutive years (Wilson et al. 1998). The same was noted for some wintering waterfowl
species in southern forested wetlands (Kaminski et al. 1993) and for avian richness in some
western riparian habitats (Sanders & Edge 1998). Variation in use of 87 Maine wetlands during
an 8-year period by 15 waterbird species was quantified by Gibbs et al. (1991). Interannual
variation in avian species richness in North American wetlands of various types surveyed by the
Breeding Bird Census program was compiled by Adamus & Brandt (1990). Common pitfalls in
the statistical analysis of multi-year data on animal use of habitats are described by Schooley
(1994). Also, in 6 Connecticut wetlands, 12 species changed their habitat associations among
seasons (Craig & Barclay 1992).
Changes in the avifauna of a series of prairie pothole wetlands on the scale of decades were
determined by Igl and Johnson (1997). Detecting a 5-10% decline in populations of some
shorebird species would require more than 10 years of survey effort, due to the difficulty of
surveying the highly mobile populations (Warnock et al. 1998). The importance of surveying
migratory shorebirds at very large spatial scales (e.g., regions) is highlighted by Haig et al.
(1998). Perhaps the most spatially-extensive wetland bird survey was conducted by Naugle et al.
(2000), who characterized 834 prairie wetlands and recommended region-scale wetland
conservation priorities based on habitat relationship models derived from statistical analysis of these
data.
Techniques
Some information applicable to surveying wetland birds is presented in Bibby et al. (1992),
Ralph et al. (1993, 1995), Gibbs & Melvin (1993), Ribic et al. (1999), Weller (1999), and
Adamus (2001). Whenever possible, surveys conducted for comparing wetlands or wetland
types using waterbirds should be done at night as well as during daytime, because habitat
selection often differs dramatically (Beyer & Haufler 1994, McNeil et al. 1992, Anderson &
Smith 1999). Technological advances such as radiotelemetry, remote time-lapse tape recording
and photography, and molecular markers are increasingly being applied not only to count birds,
but also to determine their movements and quantify their persistence and breeding success in
diverse wetlands, e.g., whether particular wetlands are "sources" or "sinks" for local birds.
Where birds that nest in cavities are present, bird boxes provide a convenient means of
monitoring reproductive success, with minimal disturbance and without the labor of having to
find nests. They have been used successfully to monitor impacts from heavy metals (Kraus
1989, Peterson and McEwan 1990) and acid precipitation (St. Louis and Barlow 1993). Another
-------
approach — time-budget analysis — involves documenting the hours a wetland is visited by
various species and usually requires purchase and installation of video equipment that
automatically photographs portions of the wetland at specified intervals. From viewing the
tapes, the duration of each activity (e.g. feeding) of visible birds in each photographed zone can
be determined. It is costly to implement for studies intended to survey more than a few
wetlands. However, time-budget analysis has demonstrated that estimates of bird density are not
necessarily sufficient to indicate a degraded wetland condition, i.e., a wetland with diminished
invertebrate densities (Eldridge & Krapu 1993). Even more labor-intensive are studies that
employ banding (Meyers & Pardieck 1993). Over long periods of time, they can provide
information on population dynamics of particular species, enabling judgments of whether a
wetland is a "source" or "sink" for breeders.
Metrics for Assessing Impacts to Wetland Bird Communities
Information on this topic is reviewed on EPA's wetland biomonitoring web page (Adamus
2001).
8.12 Literature Cited
-------
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