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Water: Monitoring & Assessment

8.0 Invertebrate Communities

Impacts on Quality of Inland Wetlands of the United States:
A Survey of Indicators, Techniques, and Applications of Community Level Biomonitoring Data
Excerpts from Report #EPA/600/3-90/073
(now out of print)


Discussions under the heading "Invertebrates" here include aquatic insects, freshwater crustaceans (e.g., amphipods, crayfish), aquatic annelids (e.g., worms), zooplankton, and terrestrial insects (e.g., the butterfly, bog elfin, and others listed by Niering 1985) that are found predominantly in wetlands.

Enrichment/eutrophication. Wetland invertebrates respond strongly to trophic condition. Abundance generally increases with increased nutrient concentrations (e.g., Cyr and Downing 1988, Tucker 1958) and species richness may decrease (Wiederholm and Eriksson 1979) or increase (Tucker 1958). Particular species assemblages of invertebrates have commonly been reported to be useful indicators of lake trophic state (Table 9) and may find similar usage in wetlands. These include:

  • aquatic worms (Oligochaeta) (Gatter 1986, Milbrink 1978, Lafont 1984, Lauritsen et al. 1985);
  • midges (Chironomidae) (Rae 1989, Wiederholm and Eriksson 1979, Winnell and White 1985);
  • snails (Gastropoda)(Clarke 1979a); and
  • clams (Sphaeriidae)(Clarke 1979b, Klimowicz 1959).

In particular, the ratios of (a) tubificid worms to aquatic insects, (b) the chironomid subfamilies Tanypodinae and/or Chironomini to the subfamily Orthocladiinae), and/or (c) cladocerans to rotifers, have been reported to increase with increasingly eutrophic conditions (Ferrington and Crisp 1989, Gatter 1986, Radwan and Popiolek 1989, Rosenberg et al. 1984). As species shifts occur with increasing eutrophication, chironomid species richness may decline; however, chironomid biomass and/or abundance increase (Ferrington and Crisp 1989, Johnson and McNeil 1988). Indeed, chironomid emergence was recommended as an efficient indicator of secondary production in lakes by Welch et al. (1988).

Organic loading/reduced DO. Excessive organic loading of surface waters, including wetlands, is known to alter community composition (CH2M Hill 1989), usually reduces invertebrate diversity and evenness (e.g., Sedana 1987), and sometimes reduces density and biomass (e.g., Hartland-Rowe and Wright 1975, Pezeshik 1987, Schwartz and Gruendling 1985, USEPA 1983). However, density and biomass of benthic invertebrates in a southern Quebec wastewater wetland was significantly greater than in unexposed wetlands (Belanger and Couture 1988). Density (Sedana 1987) and richness of invertebrates also increased in an Alabama pond after a single episodic addition of manure, but after four weeks richness declined to less than in a control pond (Deutsch 1988). Florida regulations for treated wastewater discharges to wetlands specify that "the Shannon-Weaver diversity index of benthic macroinvertebrates cannot be reduced below 50 percent of background levels as measured using standard techniques."

Under moderate loading, attached algae may increase and consequently, herbivorous mayflies and midges may dominate the community (Jones and Clark 1987). However, if turbidity and hydroperiod conditions allow submersed or floating-leaved aquatic plants (e.g., Lemna) to out-compete algae, other aquatic invertebrates may become dominant.

Table 9. Examples of Aquatic Invertebrates That May Indicate Eutrophic Conditions in Wetlands.

Compiled from the ERAPT database (Dawson and Hellenthal 1986) and the following references: Harvey and McArdle 1986, Rosenberg et al. (1984), Saether (1975), Strange 1976, and Walker et al. 1985, Wiederholm and Eriksson (1979). Note that these species may occur as well in wetlands that are NOT eutrophic, although usually in smaller proportion relative to other species.

Chironomus attenuatus Chironomus carus
Chironomus crassicaudatus Chironomus plumosus
Chironomus riparius Chironomus staegeri
Chironomus stigmaterus Chironomus tentans
Cryptochironomus blarina Cryptochironomus fulvus
Cryptotendipes casuarius Cryptotendipes darbyi
Cryptotendipes emorsus Dicrotendipes californicus
Dicrotendipes incurvus Dicrotendipes leucoscelis
Dicrotendipes modestus Dicrotendipes nervosus
Einfeldia natchitocheae Endochironomus nigricans
Glyptotendipes barbipes Glyptotendipes lobiferus
Glyptotendipes meridionalis Glyptotendipes paripes
Goeldichironomus holoprasinus Harnischia amachaerus
Harnischia boydi Harnischia edwardsi
Harnischia galeator Harnischia viridulus
Kiefferulus dux Lauterborniella varipennis
Leptochironomus nigrovittatus Omisus pica
Pagastiella orophila Parachironomus carinatus
Parachironomus directus Parachironomus hirtalatus
Parachironomus monochromus Parachironomus pectinatellae
Parachironomus scheideri Parachironomus tenuicaudatus
Paralauterborniella elachista Paralauterborniella
Paralauterborniella subcincta Paratendipes subaequalis
Pedionomus beckae Phaenopsectra profusa
Polypedilum digitifer Polypedilum halterale
Polypedilum illinoense Polypedilum simulans
Polypedilum trigonum Stenochironomus hilaris
Tribelos quadripunctatus Pseudochironomus fulviventris
Pseudochironomus richardsoni Calopsectra confusa
Calopsectra dendyi Calopsectra neoflavella
Calopsectra xantha Cladotanytarsus viridiventris
Micropsectra dubia Micropsectra nigripila
Tanytarsus buckleyi Tanytarsus quadratus
Tanytarsus recens Parametriocnemus lundbeckii
Cricotopus bicinctus Cricotopus remus
Cricotopus sylvestris Nanocladius alternanthera
Psectrocladius dyari Clinotanypus pinguis
Coelotanypus concinnus Coelotanypus scapularis
Coelotanypus tricolor Psectrotanypus vernalis
Ablabesmyia aequifasciata Ablabesmyia americana
Ablabesmyia annulata Ablabesmyia aspera
Ablabesmyia basalis Ablabesmyia cinctipes
Ablabesmyia hauberi Ablabesmyia illinoensis
Ablabesmyia mallochi Ablabesmyia monilis
Ablabesmyia ornata Ablabesmyia parajanta
Ablabesmyia peleensis Ablabesmyia philosphagnos
Ablabesmyia rhamphe Ablabesmyia tarella
Guttipelopia currani Labrundinia floridana
Labrundinia johannseni Labrundinia neopilosella
Labrundinia pilosella Labrundinia virescens
Monopelopia boliekae Procladius adumbratus
Procladius bellus Procladius culiciformis
Procladius denticulatus Procladius riparius
Tanypus carinatus Tanypus grodhausi
Tanypus punctipennis Tanypus stellatus

Table 10. Examples of Aquatic Invertebrates That Tolerate Low-Oxygen Conditions in Wetlands.

Compiled from the ERAPT database (Dawson and Hellenthal 1986) and its supporting documents (Beck 1977b, Harris et al. 1978). Note that these species may occur as well in wetlands that are NOT anoxic, although usually in smaller proportion relative to other species.


Callibaetis floridanus
Hexagenia limbata
Caenis diminuta


Chironomus attenuatus Chironomus chelonia
Chironomus crassicaudatus Chironomus plumosus
Chironomus riparius Chironomus stigmaterus
Chironomus tentans Cryptochironomus blarina
Cryptochironomus curtilamellatus Cryptochironomus fulvus
Cryptotendipes darbyi Dicrotendipes californicus
Dicrotendipes fumidus Dicrotendipes modestus
Dicrotendipes neomodestus Dicrotendipes nervosus
Endochironomus nigricans Glyptotendipes barbipes
Glyptotendipes paripes Glyptotendipes lobiferus
Psectrocladius dyari
Glyptotendipes meridionalis
Goeldichironomus holoprasinus Harnischia galeator
Harnischia viridulus Kiefferulus dux
Parachironomus monochromus Parachironomus tenuicaudatus
Paralauterborniella elachista Paralauterborniella subcincta
Paratendipes albimanus Phaenopsectra profusa
Polypedilum aviceps Polypedilum digitifer
Polypedilum halterale Polypedilum illinoense
Polypedilum nigritum Polypedilum obtusum
Polypedilum ontario Polypedilum scalaenum
Polypedilum tritum Pseudochironomus aix
Pseudochironomus chen Pseudochironomus richardsoni
Cladotanytarsus viridiventris Micropsectra nigripila
Tanytarsus buckleyi Brillia flavifrons
Cricotopus belkini Cricotopus bicinctus
Cricotopus politus Cricotopus remus
Cricotopus sylvestris Psectrocladius dyari
Coelotanypus concinnus Coelotanypus tricolor
Ablabesmyia aequifasciata Ablabesmyia aspera
Ablabesmyia mallochi Ablabesmyia monilis
Ablabesmyia rhamphe Larsia decolorata
Procladius bellus Procladius culiciformis
Tanypus carinatus Tanypus grodhausi
Tanypus neopunctipennis Tanypus punctipennis
Tanypus stellatus Chironomus attenuatus
Chironomus carus Chironomus plumosus
Chironomus riparius Chironomus stigmaterus
Chironomus tentans Cryptochironomus fulvus
Endochironomus nigricans Glyptotendipes barbipes
Glyptotendipes lobiferus Glyptotendipes meridionalis
Goeldichironomus holoprasinus Micropsectra nigripila
Cricotopus remus Procladius culiciformis
Tanypus carinatus Tanypus punctipennis
Tanypus stellatus

Table 11. Examples of Invertebrates That May Tolerate or Prefer Acidic Conditions in Wetlands.

Compiled from the ERAPT database (Dawson and Hellenthal 1986) and the following references: Beck 1977b, Kimerle and Enns 1968, Smock et al. 1981, Walker et al 1985a. Note that these species may occur as well in wetlands that are NOT acidic, although usually in smaller proportion relative to other species.

Ablabesmyia americana, A. aspera, A. basilis, A. hauberi, and A. parajanta, A. peleensis, A. philosphagnos

Chironomus (some species)
Cladopelma (some species)
Cladotanytarsus (some species)
Corynoneura taris
Cryptotendipes casuarius
Dicrotendipes incurvus, D. leucoscelis
Guttipelopia currani
Harnischia amachaerus, H. boydi
Krenosmittia (some species)
Labrundinia floridana, L. johannseni, L. neopilosella, L. virescens
Lauterborniella varipennis
Metriocnemus abdomino-flavatus, M. hamatus, M. knabi
Monopelopia tillandsia
Monopsectrocladius (some species)
Nilotanypus americanus
Nimbocera (some species)
Omisus pica
Orthocladius annectens
Pagastiella orophila
Parachironomus alatus, P. scheideri
Paramerina anomala
Polypedilum braseniae, P. nymphaeorum, P. obtusum
Procladius bellus
Tanypus neopunctipennis
Tanytarsus (some species)
Thienemannimyia senata
Tribelos quadripunctatus
Trissocladius (some species)
Dugesia tigrina
Nais (some species)
Limnodrilus hoffmeisteri
Aulodrilus piqueti
Crangonyx (some species)
Callibaetis diminuta
Caenis diminuta
Oxyethira (some species)
Palpomyia (some species)

In an isolated Florida cypress swamp dosed with treated wastewater, the following taxa were dominant:

Nais obtusa
Psychoda albicans, alternata
Chironomus riparius
Polypedilum convictus, flavus

Invertebrates that were absent (but present in untreated swamps nearby) included the following (Brightman 1976):
Lioplax subcarinata
Glyptotendipes lobiferous
Goeldichironomus sp.
Tanypus stellatus
Anomalagrion hastatum
Orthemis ferraginea

In another Florida wetland, McMahan and Davis (1978) detected no impact on terrestrial invertebrate diversity from wastewater additions, despite eutrophicated conditions that resulted. After addition of manure, some Alabama ponds previously dominated by Cladotanytarsus, Clinotanypus, and Procladius became dominated by Dero, Stylaria, and Physa (Deutsch 1988).

In a Vermont wastewater-impacted wetland, caddisflies, clams, snails, water spiders, crustacea, and all aquatic insects except midges were significantly impacted (Schwartz and Gruendling 1985). The impact was due largely to the shading out of submersed plant substrates by algal blooms. Consequently, herbivorous mayflies and midges can begin to dominate such communities (e.g., Jones and Clark 1987). However, if turbidity and hydroperiod conditions allow submersed or floating-leaved aquatic plants (e.g., Lemna) to out-compete algae, other aquatic invertebrates may become dominant.

Even in the absence of human-related wastewater influences, invertebrate communities in wetlands that naturally have low dissolved oxygen are sometimes depauperate compared to those naturally having greater oxygen (e.g., White 1985). Wetland invertebrates that appear to tolerate low oxygen levels (which typify even some undisturbed wetlands) are listed in Table 10. Ratios of tolerant to intolerant species have often been used to indicate ecological status of surface waters, and could be similarly tested for use in wetlands.

Contaminant Toxicity. The availability of vegetation may be particularly important to invertebrates in wetlands having contaminated, persistently anoxic, or highly saline sediments. In such situations, vegetation provides an colonization surface isolated from sediments, where contaminants often are concentrated; richness and abundance of epiphytic and nektonic invertebrate groups may thus remain high in well-vegetated wetlands (McLachlan 1975).

Under more severe exposure to contaminants (e.g., large ambient concentrations of dissolved metals), aquatic invertebrate species richness and density both decline, at least in shallower wetlands (Ferrington et al. 1988, Krueger et al. 1988, Winner et al. 1975). Richness and density can decline even with levels (of phenols and oil-water ratios) not known to be toxic in laboratory studies (Cushman and Goyert 1984).

Shifts in community composition occur as well. Specifically, shifts in structure away from aquatic insects and toward a community dominated by certain oligochaetes (aquatic worms) have been noted in sediments severely contaminated by heavy metals (e.g., Wentsel et al. 1978, Howmiller and Scott 1977, Winner et al. 1980). Areas that are at least moderately contaminated often are dominated by chironomid midges (Winner et al. 1980, Cushman and Goyert 1984, Rosas et al. 1985, Waterhouse and Farrell 1985) and other aquatic invertebrate species whose adults have wings and short life cycles, e.g., water bugs and water beetles (Borthwick 1988, Courtemanch and Gibbs 1979, Gibbs et al. 1981). However, responses to low levels of copper seem to be family- or genus- specific, rather than occuring at the "order" level of taxonomic

Table 10. Examples of Aquatic Invertebrates That Tolerate Low-Oxygen Conditions in Wetlands.

classification (Leland et al. 1989). Additions of heavy metals to aquatic ecosystems may increase the ratio of predators to herbivores and detritivores, at least initially (Leland et al. 1989). Nematodes may be particularly sensitive indicators of contaminant toxicity in wetlands that lack surface water; those of the subclass Adenophorea tend to be more sensitive than those of the subclass Secernentea (Bongers 1990, Platt et al. 1984, Zullini and Peretti 1986).

The commonly used herbicide, Atrazine, has been shown to cause shifts in community composition and emergence times of aquatic insects at a concentration of 2 mg/L (Dewey 1986). Other herbicides used in wetlands have been shown to increase the dominance of invertebrates tolerant of low dissolved oxygen, a result related to the large oxygen deficit commonly caused by decay of massive amounts of plants (Scorgie 1980). Also, oil and associated phenols reduced richness, diversity, and total abundance of aquatic insects in one set of wetland experiments (Cushman and Goyert 1984). The midge Cricotopus bicinctus and the aquatic worm Limnodrilus hoffmeisteri were more prevalent downstream of than upstream from an oil spill (Penrose 1989).

However, some midges (e.g., Nilotanypus fimbriatus) are reportedly very sensitive to oil (Rosenberg and Wiens 1976) and pesticides (Hanson 1952). Mayflies (except burrowing species) are particularly sensitive to metals (Leland et al. 1989, Wagerman et al. 1978), oil (Giddings et al. 1984, Cushman and Goyert 1984), and pesticides (Hurlbert et al. 1972, Ali and Stanley 1982, Van Dyk et al. 1975). Amphipods, at least the genera Gammarus and Hyallela, and the clam shrimp (Lynceus brachyurus) appear to be very sensitive to certain pesticides. As indicators of contamination, these freshwater shrimp have the added benefit of being relatively stationary (i.e., because they do not emerge and fly away like aquatic insects, their presence may be more indicative of the longer-term conditions of a wetland). Dosed populations have taken up to a year to recover. They occur in most wetlands with standing water, and their response to pesticides has been documented in prairie pothole wetlands (Borthwick 1988) and Maine bog ponds (Gibbs et al. 1981, Courtemanch and Gibbs 1979). They also have been reported as absent from stormwater treatment wetlands while present in nearby unexposed wetlands (Horner 1988).

It is conceivable that other crustacea, such as crayfish, respond similarly. However, few community-level data are available. Crayfish are damaged by copper levels of greater than 0.5 mg/L (Hobbs and Hall 1974), cadmium levels greater than 10 mg/L (Fennikoh et al. 1978), and mercury levels greater than about 2 mg/L (Doyle et al. 1978).

In wetlands that lack permanent standing water (e.g., bogs, floodplains), data on heavy metal toxicity from terrestrial invertebrate studies may be pertinent. A summary of such studies by Bengtsson and Tranvik (1989) reports the following:

  • Species richness and, less often, total abundance of terrestrial invertebrates declines with increasing metal concentration;
  • Rare species appear more sensitive than common, widespread species;
  • Least sensitive groups include soft-bodied invertebrates such as earthworms, terrestrial herbivores such as ants and weevils, and invertebrates that inhabit the upper soil layers;
  • Oribatid mites, the nematode suborder Dorylaimina, and many ground beetles (Carabidae) are highly sensitive, whereas springtails (Collembola) as a whole are less so.

The authors suggest maximum allowable concentrations for lead of less than 100-200 mg/kg; less than 100 mg/kg for copper; less than 500 mg/kg for zinc, and less than 10-50 mg/kg for cadmium.

Other thresholds of invertebrate toxicity for metals and/or synthetic organics are given by Johnson and Finley (1980), USEPA (1986), EPA's "AQUIRE" database and the US Fish and Wildlife Service's "Contaminant Hazard Reviews" series that summarizes data on arsenic, cadmium, chromium, lead, mercury, selenium, mirex, carbofuran, taxaphene, PCBs, and chlorpyrifos.

Although not directly manifested in changes in community structure, physical deformities of individuals often accompany severe pollution. Midges with deformed mouth parts were noted in areas of synthetic-coal-derived oil pollution (Cushman and Goyert 1984).

Acidification. Knowledge of acidification effects on wetland invertebrate communities comes mainly from studies in acidified lakes and streams exposed to mine drainage. As compared to circumneutral or slightly alkaline waters, acidic waters (natural or recently induced acidity) generally have less invertebrate biomass and/or species richness, lower ratio of consumers to producers, and fewer clearly dominant taxa (e.g., Friday 1987, Hall and Likens 1980, Harvey and McArdle 1986, Letterman and Mitsch 1978, Parsons 1968, Smock et al. 1981, 1985, Thorp et al. 1985, Walker et al. 1985a, Warner 1971). However, several studies, e.g., those of some acidic "blackwater" streams, have detected no significant differences in lake or stream invertebrate numbers or richness attributable to pH differences (e.g., Bradt et al. 1986, Bradt and Bert 1987, Collins et al. 1981, Crisman et al. 1980, Kelso et al. 1982, Winterbourn and Collier 1987). The effects of acidification may interact with and possibly be overshadowed by trophic conditions of wetlands (Brett 1989, Kerekes et al. 1984, Schell and Kerekes 1989).

Shifts in community composition are probably the most frequently measured effect of acidification. Particularly acid-sensitive are species of gastropods (snails), pelecypods (clams and mussels), daphnids, ephemeropterans (mayflies), amphipods (freshwater shrimp), and some midges (particularly the subfamilies Chironominae and Orthocladinae) (Allard and Moreau 1987, Bell 1971, Friday 1987, Hall et al. 1980, Harvey and McArdle 1986). Some of the first species to be affected by acidification are crustacea-- the predaceous copepod, Epischura lacustris (Sprules 1975), and the freshwater shrimp, Hyalella azteca (Zischke et al. 1983) and Gammarus lacustris. Taxa reported to be more prevalent under acidic conditions include oligochaetes, acarids (water mites), the phantom midge, Chaoborus, and midges of the subfamilies Tanypodinae and possibly Chironomini (Allard and Moreau 1987, Bradt and Bert 1987). A few caddisflies, freshwater sponges, dragonflies, water bugs (Corixidae), water beetles (Dytiscidae), and Tanytarsini midges tolerate weakly acid conditions (Fowler et al. 1985, Walker et al. 1985a). Species of midges and caddisflies known to occur under acidic conditions are listed in Table 11, based on data compiled by Beck (1977b) and others.

Salinization. Naturally saline, nontidal wetlands typically have low diversity of aquatic invertebrates (Kantrud 1989) and are dominated by brine shrimp (Artemia), brine flies (Ephydra), and a few species of midges and aquatic worms. Severe increases in salinity of freshwater habitats also can diminish invertebrate community biomass and species richness. However, rather few data have been collected specifically from inland brackish wetlands, so relative tolerances of species to increased salinity are poorly known.

Other taxa known to be relatively tolerant include certain species of midges, mosquitoes, aquatic worms, dragonflies, water bugs, and water beetles (Kreis and Johnson 1968). Crayfish generally require salinities less than 15 ppt (Loyacano 1967). Former salt marshes that were converted to freshwater wetlands were found to have fewer midges of the subfamily Orthocladiinae than expected (Walker et al. 1985a).

If more information were available on tolerances, such data might be used (e.g., as a ratio of salt-tolerant to salt-intolerant species) in conjunction with background chemical data to indicate stress to wetlands from irrigation runoff water, cultivation of saline soils, coastal saltwater intrusion, or other salinity sources.

Table 11. Examples of Invertebrates That May Tolerate or Prefer Acidic Conditions in Wetlands.

Compiled from the ERAPT database (Dawson and Hellenthal 1986) and the following references: Beck 1977b, Kimerle and Enns 1968, Smock et al. 1981, Walker et al 1985a. Note that these species may occur as well in wetlands that are NOT acidic, although usually in smaller proportion relative to other species.

Ablabesmyia americana, A. aspera, A. basilis, A. hauberi, and A. parajanta, A. peleensis, A. philosphagnos

Chironomus (some species)
Cladopelma (some species)
Cladotanytarsus (some species)
Corynoneura taris
Cryptotendipes casuarius
Dicrotendipes incurvus, D. leucoscelis
Guttipelopia currani
Harnischia amachaerus, H. boydi
Krenosmittia (some species)
Labrundinia floridana, L. johannseni, L. neopilosella, L. virescens
Lauterborniella varipennis
Metriocnemus abdomino-flavatus, M. hamatus, M. knabi
Monopelopia tillandsia
Monopsectrocladius (some species)
Nilotanypus americanus
Nimbocera (some species)
Omisus pica
Orthocladius annectens
Pagastiella orophila
Parachironomus alatus, P. scheideri
Paramerina anomala
Polypedilum braseniae, P. nymphaeorum, P. obtusum
Procladius bellus
Tanypus neopunctipennis
Tanytarsus (some species)
Thienemannimyia senata
Tribelos quadripunctatus
Trissocladius (some species)
Dugesia tigrina
Nais (some species)
Limnodrilus hoffmeisteri
Aulodrilus piqueti
Crangonyx (some species)
Callibaetis diminuta
Caenis diminuta
Oxyethira (some species)
Palpomyia (some species)

Burial/sedimentation. High rates of sedimentation (7 cm/yr) resulted in lower diversity, richness, and total community biomass in a southern river system (Cooper 1987). Fine-particle sediments, particularly if anoxic, support reduced diversity and richness of invertebrates (Wilbur 1974). Species of mayflies and chironomids Species of mayflies and chironomids that feed mainly on algae are particularly affected, while burrowing invertebrates might be expected to be least-affected. In Lake Erie, the abundance of tubificid worms was correlated with the sediment accumulation rate and organic carbon flux (but not to organic carbon) (Robbins et al. 1989). Excessive sedimentation may be indicated by absence of the freshwater bryozoans, e.g., Pectinella magnifica, and the fingernail clam Sphaerium rhomboideum (Cooper 1987, Cooper and Burris 1984).

Turbidity/Shade; Vegetation Removal. Removal of aquatic bed vegetation can increase algae in wetlands, thus increasing the ratio of herbivorous species (e.g., certain mayflies) to detritivorous species (e.g., certain midges and worms). Submersed plants and logs have among the highest densities and species richness of any aquatic substrate, e.g.:

Armstrong and Nudds 1985, Boerger et al. 1982, Chubb and Liston 1986, Crowder and Cooper 1982, Durocher et al. 1984, Dvorak and Best 1982, Floyd et al. 1984, Gilinsky 1984, Hall and Werner 1977, Kallemeyn and Novotny 1977; Kimble and Wesche 196, Krecker 1939, Krull 1970, Menzie 1980, Minkley 1963, Miller et al. 1989, Mittelbach 1981, Poe et al. 1986; Scheffer et al. 1984, Schramm et al. 1987, Teels et al. 1978, Voigts 1976, Ware and Gasaway 1978, Wetzel 1975.

Indeed, equations for predicting the density of aquatic invertebrates in submersed vegetation (lacustrine aquatic bed) have been developed by Cyr and Downing (1988), using data on biomass of individual macrophyte species and season. Thus, removal or loss of aquatic vegetation due to shading/turbidity can be expected to profoundly affect the invertebrate resource (e.g., Bettoli 1987, Vander Zouwen 1983). On the other hand, selective removal of dense macrophyte stands can increase density, biomass, and/or richness of remaining invertebrate communities (e.g., Beck et al. 1987, Broschart and Linder 1986, Kaminski and Prince 1981, Kenow and Rusch 1989, Murkin and Kadlec 1986). Removal of the canopy of one forested floodplain wetland had little effect on aquatic invertebrate richness and density (Boschung and O'Neil 1981). The degree to which vegetation removal has a neutral or beneficial effect on macroinvertebrates may depend partly on the type of removal procedure (e.g., mechanical thinning, ditching, burning, herbicides, crayfish introduction) and the spatial patterns created (Nelson and Kadlec 1984).

Non-aquatic invertebrates may also respond to removal of woody and emergent vegetation. For example, a decline of wetland spider richness accompanied peat harvesting in a bog (Koponen 1979).

Thermal Alteration. Heated effluents generally reduce the richness of invertebrate communities in wetlands and may either increase or decrease their density and productivity (Gibbons and Sharitz 1974, McKnaught and Fenlon 1972, Nichols 1981, Poff and Matthews 1986, Whitehouse 1971, Wiederholm 1971). Increases in secondary productivity are the result of higher primary productivity associated with warmer temperatures and longer growing seasons. Crayfish generally cannot tolerate temperatures greater than about 30 C (Becker et al. 1975). Backswimmers (Corixidae) and midges appear to tolerate moderately warmed surface waters (Gibbons and Sharitz 1974). Temperatures of over 40 C apparently do not significantly affect the life cycle of the midges Chironomus sp., Tanypus neopunctipennis, or Tanytarsus sp.; where deep, soft substrates are available as refugia for burrowing species, damage from thermal increases may be lessened (Coler and Kondratieff 1989). The ratio of burrowing oligochaetes, nematodes, gastropods, chironomid midges, and nektonic invertebrates to other aquatic invertebrates might thus be tested as one indicator of thermal disturbance.

Dehydration, Inundation. Water levels profoundly affect the abundance and community composition of invertebrates (Reid 1985, Wiggins et al. 1980). Addition of permanent open water to a non-permanently flooded wetland increases the opportunity for invasion by many submersed and floating-leaved species that provide complex substrates for aquatic invertebrates. This consequently can result in an increase in on-site species richness, and perhaps increased density, of wetland invertebrates. For example, inundation of emergent wetlands was noted to increase the density, biomass, and richness of invertebrates (Huener 1984), and cause a shift in community composition toward herbivores and detritivores (Murkin and Kadlec 1986). For Mississippi River borrow pit wetlands, "days flooded" was the most significant factor explaining invertebrate density in a multivariate regression; flooding in the sampled wetlands ranged from 24 to 115 days annually, with a mean of 81 (Cobb et al. 1984).

However, if inundation in some wetlands is prolonged (throughout the growing season) and deep, the resulting oxygen and light deficits may result in diminished richness and density of aquatic plants (Ebert and Balko 1987). Prolonged growing-season flooding, when it occurs in wetlands that have no prior history of such flooding, results in diminished invertebrate density and richness (Driver 1977, Hynes and Yadev 1985, Neckles et al. 1990). In forested floodplain wetlands, invertebrate species richness and abundance decrease with increasing soil moisture and flood frequency (e.g., Uetz et al. 1979) and with disruption of normal sequencing of flooding (Sklar and Conner 1979).

When wetlands that normally contain standing water are almost totally dehydrated for short periods (i.e., "drawdown"), the result is usually a major increase in nutrients, algae, and invertebrate density (Benson and Hudson 1975, Reid 1985, Wegener et al. 1974). This effect may be less pronounced if a dense canopy prevents sufficient light for algal growth, and exchange rates of wetland water with adjacent waters are minimal. Also, less mobile taxa, such as freshwater clams, may be particularly sensitive to drawdown. They can become stranded and perish during rapid drawdown unless underlying sediments remain saturated and soft so individuals can burrow down into the saturated zone (Jiffry 1984).

Although invertebrate density may increase following reflooding of dehydrated wetlands, invertebrate richness may not, particularly if sediments have become heavily oxidized and hardened during exposure (Hunt and Jones 1972). If wetlands are dehydrated irregularly and rapidly (e.g., by frequent passage of large ships) or for long periods (e.g., reservoir fluctuations), both abundance and richness of invertebrates can decline (Hale and Baynes 1983, Smith et al. 1987).

Invertebrate taxa can be classified into groups (response guilds) related to their life cycles and preference for particular wetland hydroperiods. Conceivably, ratios of these groups (e.g., density-weighted ratio of short-lived/mobile species to longer-lived/immobile species) could be tested as an indicator of wetland hydrologic status, as has been done with midges (Driver 1977) and water beetles (Hanson and Swanson 1989). In prairie pothole wetlands, chironomid diversity was also found to increase with permanency of the hydroperiod (Driver 1977), although contrary evidence is presented by Neckles et al. (1990). Individual taxa might be assigned to the following response groups (Delucchi 1987, Jeffries 1989, McLachlan 1970, 1975, 1985, Wiggins et al. 1980):

  • Overwintering Residents: disperse passively; include many snails, mollusks, amphipods, worms, leeches, crayfish.
  • Overwintering Spring Recruits: reproduction depends on water availability; include most midges, some beetles.
  • Overwintering Summer Recruits: reproduce independent of surface water availability, requiring only saturated sediment; include dragonflies, mosquitoes, phantom midges.
  • Non-wintering Spring Migrants: mostly require surface water for overwintering, adults leave temporary water before it disappears in spring or summer; includes most water bugs, some water beetles.

Thus, changes in density-weighted ratios of response groups, monitored from a large regional set of wetlands, might be used to indicate changing hydrologic conditions over time. However, additional research may be needed because some recent evidence suggests that certain taxa (species of Dytiscidae, Corixidae, Ceratopogonidae, Ephydridae, and even Chironomidae) may be unaffected by water regime in some situations (Neckles et al. 1990).

Fragmentation of Habitat. We found no explicit information on wetland invertebrate community response to fragmentation of regional wetland resources. A study of prairie potholes indicated increased diversity with increased wetland size, and the author suggested that might be due to the increased distance of smaller areas from larger and more stable wetlands (Driver 1977). Increased richness and interspersion of plant forms within a wetland can result in increased macroinvertebrate richness and numbers (Voigts 1976).

One can surmise that as the distance between wetlands with colonizers becomes greater, species with narrow environmental tolerances and which do not disperse easily might be most affected. Indeed, in a study of essentially identical wetlands, Jeffries (1989) found that statistical clusters of invertebrate taxa were defined by the distance and surface water connection of their associated wetland from a much larger regional water body. However, even apparently "immobile" species such as amphipods and clams have some capability for dispersal (Swanson 1984).

Landscapes where wetlands are interspersed with uplands can have almost 70 percent more invertebrate species than those containing only uplands (Coulson and Butterfield 1985). In lakes, the species richness of mollusks (Aho 1978, Lassen 1975), midges (Driver 1977), and crustaceans (Fryer 1985) increases with increasing lake area.


Natural factors that could be important to measure and (if possible) standardize among wetlands when monitoring anthropogenic effects on community structure of invertebrates include:

age of wetland (successional status), water or saturation depth, conductivity and baseline chemistry of waters and sediments (especially pH, alkalinity or calcium, and organic carbon), sediment type, current velocity, presence of fish, stream order or ratio of discharge to watershed size (in riverine wetlands), density, type, and form of vegetation and woody debris (particularly, total surface area), ratio of open water to vegetated wetland, and the duration, frequency, and seasonal timing of regular inundation, as well as time elapsed since the last severe inundation or drought.

Sampling methods for wetland or lake littoral invertebrates are described in Downing and Rigler 1984, Edmondson and Winberg 1971, Fredrickson and Reid 1988b, Isom 1986, Murkin and Murkin 1989, Witter and Croson 1976, and others. Although addressing streams, the book by Elliott (1971) is an important reference for sampling program design and data analysis.

Larval aquatic invertebrates can be found in wetlands throughout the year. If wetlands can be sampled only once, then the late wet season or beginning of the dry season, if they coincide with the growing season, are usually the recommended time, as density and richness tend to be greatest then (Marchant 1982). Alternatively, if conditions among a series of years are to be compared and the primary desire is to minimize variability, then dry-season measurements made just before the onset of flooding may be best (McElravy et al. 1989). However, the chronology of density peaks can vary even among wetlands in close proximity, possibly due in some cases to differences in predation (Campbell 1983).

In either case, and particularly in disturbed and intermittently flooded wetlands, caution is needed to schedule sampling to coincide with phenologies of particular taxa (Sklar 1985). For example, one might want to avoid sampling immediately after a synchronous emergence of the usually dominant species. Maximum information is often obtained when most invertebrates are within a size range (later instars) retained by nets used to sample them, and can be identified with greatest confidence. For biomass estimates, Hanson et al. (1989) reported that samples collected at 4- and 6- week intervals were very similar to those based on 9 biweekly collections. For a bog stream monitored over 23 months, Boerger et al. (1982) reported a 17-fold variation in midge densities, and even greater variation was reported by Gatter (1986).

The choice of equipment depends largely on the wetland microhabitat to be sampled. Different assemblages of wetland invertebrates inhabit sediments (benthos), rooted plants or algae (phytomacrofauna), open water (nekton), and the surface film (neuston). Subsequent data analysis can use groupings based on ecological niches associated with each taxon (e.g., Cummins and Wilzbach 1985).

A significant problem in analyzing wetland invertebrate data arises from difficulties in determining the spatial dimensions of the area from which a sample was drawn. Accurate estimates of density (individuals per unit area) are difficult to achieve due to difficulties in accurately measuring the complex wetland substrate (submerged plants, tree trunks, emergent plant stems, logs, etc.). To address this, some investigators have removed the substrate along with the collected sample, weighted both, and reported density as weight or number of organisms per unit weight of substrate. In some cases regression coefficients have been calculated to convert plant weights to plant area, which may be further converted to invertebrate density (Downing 1986). Another approach has been to base comparisons among similar wetland habitats on similarity indices and richness (per number of individuals), rather than on density and biomass.

If the objective is to sample invertebrate communities attached to wetland plants (e.g., snails, many mayflies) and the water column, sweep nets (dip nets) are commonly used. These are the familiar long-handled insect nets. They may be used in water or air, so long as vegetation is not dense. Usually, they are either swept through a standard length of vegetation, or placed on the bottom and hauled vertically through the water column in a rapid stroke. They are convenient to use, and are particularly suited for capturing large (e.g., crayfish) or quick-moving species not collected by other methods, such as adult dragonflies and water striders. Disadvantages include user variability and the fact that their samples are not strictly quantitative, since the unit of area swept is difficult to accurately determine (Adamus 1984, Plafkin et al. 1989).

Trials by Furse et al. (1981) and Friday (1987) indicate that at least 80 percent of the species found to be present in a particular aquatic plant bed using 5 to 10 sweeps can be captured in half that number. In trial comparisons against a modified Gerking sampler (see below), Kaminski and Murkin (1981) found sweep nets to be just as effective in sampling water-column taxa, although Gillespie and Brown (1966) had come to the opposite conclusion. In wetland studies, sweep nets have been used by Borthwick (1988), Courtemanch and Gibbs (1979), Smith et al. 1987, Voigts 1976, White 1985, and others.

Another option for sampling plant-dwelling invertebrates in wetlands involves directly clipping the vegetation and returning it in an enclosed box to the lab. This can be used for both submersed and emergent plants, and provides more precise quantification than does use of sweep nets. Vacuum suction can also be used to remove small invertebrates from foliage in the field (Southwood 1981). Downing and Cyr (1985) found the most cost-effective quadrat size for clipping to be 500 cm2. Plants were enclosed in a 6-liter plastic box. Clipping aquatic macrophtyes in quadrats of varying sizes yielded five times higher populations than did sampling with Gerking, Macan, Minto, or KUG samplers. Gates et al. (1987) described a sampler useful for taking simultaneous samples of sediment invertebrates and plant-dwelling invertebrates. They found this to give results for plant invertebrates at least as precise and sometimes more accurate than obtained by clipping macrophytes.

A third option for sampling invertebrates of wetland plants involves use of artificial substrates. Plants are not sampled directly, but rather, plastic plants or other sterile surfaces (e.g., Hester-Dendy plate samplers) are totally submersed in the wetland water column and allowed to be colonized over a period of at least a month (Macan and Kitching 1972). Because they standardize surface area and texture, collections from substrate samplers are highly comparable to each other, making them attractive for use in monitoring of water column water quality. They also are lightweight, can be used in areas difficult to sample by other means (e.g., deep rivers), and sample processing is relatively clean. However, disadvantages include the fact that a return trip to the wetland is required, vandalism may be a problem, their use is limited to wetlands with surface water, they sample only epiphytic species, and representativeness can be questioned (Adamus 1984).

In an aquatic bed wetland, Gerrish and Bristow (1979) used plastic mimics of the pondweed, Potamogeton richardsonii, interspersed among live experimental plants. Although this yielded no significantly different numbers of invertebrates or species per unit of surface area than were found on real plants, aquatic worms were significantly more common on the artificial substrates, and the substrates did not accurately reflect the densities of invertebrates on the nearby Myriophyllum or Vallisneria plants.

Natural substrates initially devoid of organisms can also be used as colonization substrates. For example, plant litter was placed in boxes made of hardware cloth by Batema et al. (1985) and White (1985), for sampling macroinvertebrates in eastern floodplain forests. Artificial substrates are often ineffective for collecting large crustaceans (e.g., crayfish) and mollusks.

If the objective is to sample invertebrate communities inhabiting wetland sediments, then dredges-- also called grab samplers (Ekman, Ponar, etc.)--are often used. They essentially consist of a box with jaws that is lowered onto the sediment. The jaws enclose a specified area of bottom, and retrieve sediments and associated organisms to a sediment depth of about 5 cm. Dredges are used only where surface waters of at least 0.5 m in depth are present, and are not effective where there are rocks, aquatic plants, or logs to jam the jaws. They have been used in wetlands by Bradt and Bert (1987), Driver (1977), and Krull (1970). Estimates of density are only crudely quantitative because jaws seldom close tightly, allowing organisms to escape. Large organisms (e.g., crayfish), water column organisms, and fast-moving species are poorly sampled.

Another option for sampling sediments is to use core samplers. Unlike grabs, corers do not have jaws, and instead rely on compactive force or suction to retrieve sediments. They suffer the same disadvantages as dredges. Samples may be more precisely quantitative, but the mean size of organisms effectively captured may be smaller, due to the narrowness of corers. Core samplers may be the only option for quantitatively sampling sediment organisms in wetlands that lack surface water, and a variety of designs are available (e.g., Bay and Caton 1969, Coler and Haynes 1966). Core samplers are widely used in paleoecological studies of wetlands. Florida regulations for monitoring treated wastewater discharges to forested wetlands specify that, if a core sampler is used, devices with minimum sampling area of 45 cm2 be used, and that the number of samples at a given station within the wetland be that number needed to be 90% certain of being within 15% of the mean diversity of the population. Where aquatic plants interfere, some investigators have suggested a saw blade might be welded to the leading edge of the corer, for clipping heavy roots and stems (e.g., Murkin and Kadlec 1986). Where sediments are frozen, metal ice spades have been used to collect samples (e.g., Jacobi 1978).

Where sediments or soils are not covered by water (e.g., in peat bogs), pitfall traps and soil extraction techniques can also be used to augment vacuum sampling and sweep-net sampling, and may produce the highest densities and species richness (Coulson and Butterfield 1985).

If the objective is to sample invertebrates that inhabit the water column, tubular samplers (e.g., "Gerking samplers", "stovepipes", "Hess samplers", "box samplers") can be used. These are wide cylinders that enclose a standard area of bottom and usually are not designed for effectively penetrating the sediment. In some, the bottom can be sealed off with a sliding door, plug, or similar feature once the sampler is in place. Some have been fitted with a reinforced cutting edge on the bottom. Designs are described by Freeman et al. (1984), Gerking (1957), Korinkova (1971), Hiley et al. 1981, Legner et al. 1975, Mackay and Quadri (1971), Martin and Shireman (1976), Minto (1977), and Swanson 1978. They are not effective for catching quick-moving organisms, burrowing species, very large taxa, many epiphytic species, or for use in flowing water.

Emergence traps and funnel traps consist of nets or funnels anchored at and just above the water surface. They passively collect aquatic insects as they pass into their winged adult stage and emerge from the water column. Funnel traps are used to collect swimming, air-breathing insects as well as emerging species (e.g., Greenstone 1979, Henrickson and Oscarson 1978, Kaminski and Murkin 1981). Traps--either submerged, at the water surface, or above it-- can be fitted with lights to increase their attraction to some adult insects, for example:

Aiken 1979, Apperson and Yows 1976, Carlson 1971, Carlson 1972, Espinosa and Clark 1972, Hungerford et al. 1955, Husbands 1967).

A variety of designs for emergence and funnel traps have been tried, for example:

Corbet 1965, Daniel et al. 1985, Deonier 1972, Lammers 1977, Lemke and Mattson 1969, McCauley 1976a,b, Pritchard and Scholefield 1980, Rosenberg et al. 1980, Voigts 1973, and Washino and Hokama 1968, and some were evaluated by Kimerle and Anderson (1967).

Use of funnel and emergence traps is limited to wetlands containing open patches of surface water during the growing season, when most insects emerge. They can be used in both still-water and slow-flowing wetlands, particularly those difficult to sample by other means, and samples are relatively debris-free and easy to process. Because they are left in place (sometimes for many weeks), they avoid the problem encountered by other samplers of missing key species due to inappropriate time of visit.

Because emerging insects come from a variety of microenvironments, emergence traps can integrate well the extreme spatial heterogeneity within many wetlands. On the other hand, this makes it impossible to standardize or determine the unit of area measured. Thus, they would not be suitable for tracing the leading edge of an effluent plume within a small wetland. Samplers designed for passively collecting terrestrial insects (e.g., light traps, pitfall traps, malaise traps) encounter the same problem. Also, many of the wetland invertebrates most sensitive to pollution do not emerge (e.g., amphipods, aquatic worms, snails), so are not collected by emergence traps. Initial purchase of traps can be costly, and vandalism may be problematic.

In prairie potholes, conical emergence traps were situated at 3 m intervals perpendicular to shore (Driver 1977). To detect immediate effects of pesticide application, Gibbs et al. (1981) emptied emergence traps every two hours, from 6 AM to 8 PM. Normally, traps are left in place for many days or weeks. Welch et al. (1988) used submerged funnel traps to catch emerging midges in a lake. They found no difference in total catch between 0.142-m2 and 0.283-m2 trap sizes. Traps with inverted funnels inserted in the jar necks caught more pupae than traps without funnels, and total catch in the traps without jars was 58 percent of the catch in traps with funnels. Rosenberg et al. (1984) submerged their funnel traps, situating them at depths of 1, 2, 3.5, and 4.5 m.


In general, quantitative data on wetland macroinvertebrates has not been uniformly collected from a series of statistically representative wetlands in any region of the country. Thus, it is currently impossible to state what are "normal" levels for parameters such as seasonal invertebrate density, species richness, or biomass, and their temporal and spatial variability, in any type of wetland.

Perhaps the closest approximation of such a data set is that of Giese et al. (1987). These invertebrate data were collected in part from streams flowing through relatively pristine floodplain wetlands, and thus help serve as a regional baseline for bottomland hardwood wetlands. Collecting a single, timed (30-minute), series of dip-net samples from each site, the investigators found an average of 50-60 invertebrate taxa, containing an average total of about 800 individuals. The Shannon diversity index averaged 4.17 to 4.67 in these relatively pristine lowland streams.

Also, the U.S. Geological Survey is presently initiating a program (NAWQA) to monitor stream invertebrate communities inhabiting a carefully selected sample of watersheds throughout the United States. Although wetlands will not be a specific target of the monitoring, the spatial and regional variability of invertebrates may become better known from this probability sampling approach. Another regionally extensive project was undertaken by Corkum (1989) in the Pacific Northwest/Alaska, and resulted in an ecologically-based classification of stream types for that region. Other data on wetland macroinvertebrates is selectively summarized in Table 12.

Coefficients of variability for invertebrates in streams range from about 0.2 to 0.8 (Eberhardt 1978). Few such values apparently have been published for wetlands. Variation in invertebrate density among habitats within wetlands has been documented in some cases, for example:

Beck 1977a, Erman and Erman 1975, Gatter 1986, Kansas Biological Survey 1987, Neuswanger et al. 1982, Thorp et al. 1985.

However, only a few studies in the U.S. have quantified invertebrate community differences among a series of wetlands in a region, and have mostly focused on lacustrine or riverine wetlands. These include:

Bradt et al. 1986, Campbell 1983, Cobb et al. 1984, Erman and Erman 1975, Ferren and Pritchett 1988, Garono and MacLean 1988, Haack et al. 1989, Hepp 1987, Kallemeyn and Novotny 1977, Krull 1970, Lowery et al. 1987, Mathis et al. 1981, McAuley and Longcore 1988, Stoddard 1987, Teels et al. 1978.

Although data exist that quantify year-to-year variation in invertebrate community structure in other surface waters (e.g., McElravy et al. 1989), such studies have apparently not been published for wetlands. Conceivably such unpublished data may be available from sites of the U.S. Department of Energy's National Environmental Research Park system, as well as the following sites of the National Science Foundation's Long Term Ecological Research (LTER) program (that contain studied wetlands): Illinois Pool 19 site, Illinois-Mississippi Rivers sites, New Hampshire Hubbard Brook riparian forest, Oregon Andrews Experimental Forest riparian forest, and Michigan Kellogg Biological Station site.

Quantitative published data on composition of aquatic invertebrate communities appears to be most available for submersed vegetation (aquatic bed wetlands), particularly in the Southeast and Prairie pothole region. Apparently such data are most limited for wetlands that are saturated but mostly lack standing water (e.g., bogs), as well as for playas and non-Southeastern riparian wetlands. Information is most available on impacts of hydrologic alteration, acidification, and nutrients, and least on impacts of salinization, sedimentation, thermal alteration, and habitat fragmentation.

Even qualitative lists of "expected" aquatic invertebrates in wetlands of various types do not appear to have been compiled, either nationally or by individual states. The USFWS has begun to compile such lists (pers. comm., Buck Reed, USFWS, St. Petersburg, FL) and some publications in the "community profile" series of the USFWS (Appendix C) mention particular taxa known to occur in wetlands.

Table 12. Examples of Invertebrate Density and Biomass Estimates from Wetlands.

BIOMASS (g/m2)

State type* N min.value max.value citation

Lowery et al. 1987
AR** Pfo

Cobb et al. 1984

Tebo 1955
LA Pab 48 0.5 15.7
Sklar 1985
LA Pab ? 4.2 6.8
Sklar & Conner 1983
MS Pfo ?
Baker et al. 1988
WA Pfo 18 2.5 5.7
Meehan-Martin & Swanson 1988
WI Pem ? 0.6 1706
Schmal & Sanders 1978
SD Pem 220 1.3 8.5
Broschart & Linder 1986
CA P(fen)
8.5 Erman & Erman 1975

DENSITY (number/m2)


type N min.value max.value citation

>10,000 Lowry et al. 1987
AR*** Pfo

mean=2967 Cobb et al. 1984
CA Pem 230 6952 23,857 Erman and Erman 1975
FL Pfo

1,102 Brightman 1984
FL Pfo

2.5 Brightman 1984
IA Pem

>20,000 Voights 1976

4,108 Tebo 1955
KS Pem ? 508 18,676 Kansas Biol. Surv. 1987
LA Pab ?
76,990 Sklar and Conner 1979
LA Pfo

16,198 Sklar 1985
LA Pfo

12.5 Sklar 1985
LA Pfo ? - 16,000 Sklar & Conner 1979
LA Pfo 70
mean=95/grab Beck 1977a
Pem 13

mean=2900/grab Beck 1977a
MI Lab

>10,000 Lowery et al. 1987
MO Pfo

>9,000 Batema et al.
MS Pfo ? 1675 9248 Baker et al. 1988
GA Pab ?
mean=12,093/m2 Smock & Stoneburner 1980
MO Pfo 4
5045 Batema et al. 1985
IL-MO L 33 247 4321 Jones et al. 1985
KY Pem 84 739 5143 Bosserman & Hill 1985
NJ Pem ? 196 335,547 Gatter 1986
WI Pem ?
35,730 Schmal et al. 1978
MI Lab ? 7665 13,243 Hiltunen & Manny 1982
SD Pem 175 584 5929 Benson & Hudson 1975
SD Pem 220 3533 15,193 Broschart and Linder 1986
OR R 64 33 15,700 Kreis & Johnson 1968
OR Pem ? 11 1745 Fishman 1989

* wetland codes (Cowardin et al. 1979): Pab=palustrine aquatic bed, Pem=palustrine emergent, Pfo=palustrine forested, L=lacustrine, R=riverine

** includes TN and MS

***includes LA, MS, and MO

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