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Water: Monitoring & Assessment

2. Algal and Microbial Communities as Indicators of Prairie Wetland Integrity

Bioindicators for Assessing Ecological Integrity of Prairie Wetlands
Report # EPA/ 600/ R-96/ 082
September 1995

Contents:

2.1 Ecological Significance and Suitability as an Indicator
2.2 Potential Indicator Metrics
2.3 Previous and Ongoing Monitoring in the Region
2.4 Response to Stressors

2.4.1 Algae and Microbial Communities as Indicators of Hydrologic Stressors
2.4.2 Algae and Microbial Communities as Indicators of Vegetative Cover Condition
2.4.3 Algae and Microbial Communities as Indicators of Wetland Salinity
2.4.4 Algae and Microbial Communities as Indicators of Sedimentation and Turbidity
2.4.5 Algae and Microbial Communities as Indicators of Excessive Nutrient Loads and Anoxia
2.4.6 Algae and Microbial Communities as Indicators of Pesticide and Heavy Metal Contamination

2.5 Monitoring Techniques

2.5.1 Direct Sampling
2.5.2 Indirect Sampling Through Measurement of Processes
2.5.3 Analysis of Historical Conditions
2.5.4 Bioassay Methods

2.6 Variability and Reference Points

2.6.1 Spatial Variability
2.6.2 Temporal Variability

2.7 Collection of Ancillary Data
2.8 Sampling Design and Required Level of Sampling Effort
2.9 Summary


2.1 Ecological Significance and Suitability as an Indicator

In prairie wetlands, four algal assemblages are commonly defined: phytoplankton (algae suspended in the water column), metaphyton (unattached and floating or loosely associated with substrata), and two assemblages of attached algae, namely epiphytic algae (attached to plants) and benthic or epipelic algae (attached to sediments) (Crumpton 1989). The regional pool of species is probably largest for the benthic and phytoplanktonic forms, although individual samples of communities are usually taxonomically simple. Microbes prevalent in prairie wetlands include bacteria, viruses, yeasts, and microscopic fungi.

Because they form the basis of the food chains, algal and microbial communities are of critical ecological importance in prairie wetlands. The prairie wetland animal community depends on solar energy that has been converted by photosynthesis to biomass by algae or vascular plants (Neill and Cornwell 1992). The algae and vascular plants provide a substrate and food source for microbial communities, and the microbes in turn are consumed extensively by a wide variety of detrivorous invertebrates. Algae, and to a lesser degree vascular plants (Campeau et al. 1994), are also consumed directly by invertebrates. In some prairie wetlands, the levels of algal and microbial productivity approach those of vascular plants.

The relative importance of microbes vs. algae as supporters of prairie wetland invertebrates depends on interactions between microbes and algae, season, and water regime and chemistry of a particular wetland. Based on a statistical analysis of invertebrate biomass in submersed plant beds in 11 eastern Canadian lakes, Lalonde and Downing (1993) concluded that aquatic invertebrate biomass was highly correlated with the biomass of littoral phytoplankton, epiphyton, and vascular plants, a situation also noted in prairie wetlands (Murkin et al. 1991). However, Murkin et al. (1992) suggest that algae's role in determining the horizontal distribution of invertebrates within a prairie wetland might not be as great as it would seem. They based this conclusion on their failure to find spatial overlap between maximum density of macroinvertebrates and maximum epiphytic biomass (measured as chlorophyll) in Delta Marsh. Similarly, in most of the Cottonwood Lake wetlands that LaBaugh and Swanson (1993) studied, they found that seasonal and annual changes in microinvertebrate abundance were not statistically related to the abundance of algae. They speculated that microinvertebrates might be depending more on microbial biomass. Ducks which feed on wetland invertebrates also do not congregate in wetlands that have the most algae, although ducks that use algae-rich wetlands were found in a Saskatchewan study (Gloutney 1993) to spend more time feeding. Using isotope ratios to investigate food webs in Delta Marsh, Neill et al. (1992) found that emergent vascular plants and the algae and microbes attached to them, rather than submersed macrophytes or metaphyton, were the most important sources of organic matter to the invertebrate consumers that were most abundant during June.

In addition to their value as food for wetland animals, algae influence invertebrates and vascular plants in at least five other ways: (1) phytoplankton and metaphyton reduce light penetration of the water column, (2) epiphytic and benthic algae mediate the levels and flux of nutrients, contaminants, and oxygen across ecotones (e.g., sediment-water column) and plant surfaces, (3) all algae (when they respire, die, and decompose) can diminish dissolved oxygen in the water column and sediments, (4) some blue-green algae fix (add) gaseous nitrogen to the water column, thus enriching wetlands, (5) some blue-green algae are toxic to other organisms. Any of these phenomena can alter the basic habitat structure of prairie wetlands, and thus, the populations of invertebrates, amphibians, fish, and waterbirds. Moreover, microbial populations can alter the structure of habitat available to invertebrates because they are the major assemblage responsible for decomposition, and decomposition is usually the primary factor responsible for offsetting excessive accumulation of plant litter in prairie wetlands. Excessive plant litter accumulation, with associated increase in microbial densities, can create oxygen deficits in sediments and surface waters (Baird et al. 1987, Barica 1984, Barica and Mathias 1979), which can impair seed germination, damage invertebrate communities, and further retard decomposition.

Certain microbial communities also can detoxify some chemicals. For example, microbes associated with wetland plants can detoxify some synthetic organic compounds (Hodson 1980) such as pentachlorophenol (Pignatello et al. 1985), the herbicides glyphosate (Goldsborough and Beck 1989) and atrazine (W. Crumpton, Iowa State Univ., personal communication), as well as detergents (Federle and Schwab 1989).

Particular microbial communities also can reduce, via the process of denitrification, the overenrichment of wetlands. This is important at landscape and regional scales not only because wetlands are among the most effective ecosystems for removing nitrogen (Groffman and Tiedje 1989), but also because they intercept much of the runoff and groundwater before it reaches larger, more permanent waterbodies. Protection and enhancement of nitrogen removal functions of wetland microbial communities is important to maintaining and restoring important public uses of downslope lakes, rivers, and aquifers. In Iowa semipermanent wetlands, Davis and van der Valk (1978b) reported removal of 86% of the inputted nitrate and 78% of the ammonia. Two open wetland complexes in North Dakota removed 13 and 58% of the tributary nitrate, as compared to a drained wetland complex in which there was a > 10-fold increase in nitrate (Malcolm 1979). An unvegetated wet basin in South Dakota that was loaded with municipal wastewater removed 1765 kg N/ha (White and Dornbush 1988). On a regional basis, Jones et al. (1976) in northwestern Iowa found that among 34 watersheds, those with a large percentage of land as wetlands had less nitrate in streamflow than those with a small percentage of wetlands.

Denitrification rates may be equal or greater at the beginning and end of the growing season than during mid-summer (Christensen 1985, Myrold 1988, Zak and Grigal 1991). Thus, denitrification functions in wetlands may be of greatest value in removing nitrate during years when runoff inputs occur early or late in the growing season. However, if runoff resulting from the spring melting of snow surrounding wetlands occurs prior to ice-out in wetlands, the runoff flows under the wetland ice, purging the basins of anoxic, ammonia-rich water which can subsequently be released into receiving waters (if seasonal connections exist) without being substantially denitrified, thus causing water quality problems (NDDHCL 1990). This adverse impact might be more likely to occur in landscapes dominated by semipermanent and permanent basins, because they tend to remain frozen longer than temporarily flooded wetlands.

As indicators of wetland integrity, microbial communities have several characteristics that are usually considered to be advantages (Adamus and Brandt 1990):

  • tight linkage to fundamental processes (e.g., decomposition, denitrification, respiration)
  • easily collected and transported samples
  • USEPA protocols available (but may need adaptation to wetlands)
  • immediate response to contamination
  • measurable in wetlands year-round, and in wetlands that lack surface water
  • sensitive to presence of some contaminants, and assay protocols are available (e.g., Ames test, Microtox test)
  • sufficient published information to identify a few "indicator taxa" that clearly are associated with particular stresses

Characteristics that are usually considered to be disadvantages when using microbial communities to determine wetland condition include:

  • lack of a response that is identifiably stressor-specific
  • laborious and slow identification (plate culture); process measurements can be difficult to interpret with regard to ecological significance
  • rapid turnover of individuals of most species requires frequent sampling; microbes do not integrate conditions over time very well
  • naturally great micro-spatial variation
  • impractical for detecting bioaccumulation

Algae have several characteristics that are usually considered to favor their use for monitoring ecosystem integrity (Adamus and Brandt 1990):

  • Algae are pivotal in food webs, and maintain tight linkage to fundamental processes (e.g., photosynthesis, respiration).
  • USEPA protocols for sampling both structure and function of algal communities are available (may need modification to wetlands).
  • Standardized collection procedures are available and impact the monitored wetland only minimally.
  • Tolerances and indicator value are well-known, particularly with regard to nutrients.
  • Rapid reproduction rates and short life cycles make algae and microbes sensitive to short-term impacts.
  • Algae are generally immobile and thus reflect conditions at a particular site, making them useful for in situ exposure assessments and whole-effluent bioassays.
  • Decay-resistant remains (diatom frustules and pigments) provide a means for establishing historical reference conditions in a wetland.

Characteristics that are usually considered to be disadvantages when using algae to determine wetland condition include:

  • laborious identification of many taxa
  • rapid turnover requires frequent sampling
  • complicated interpretation of spatial patterns because of drifting cells of unattached species
  • relatively insensitive to heavy metals and insecticides (Hellawell 1984)
  • difficult to link responses of many algal taxa to a specific stressor
  • poorly documented linkages to other components of the food chain in prairie wetlands

2.2 Potential Indicator Metrics

Various measurements and metrics can be applied to algal and/or microbial samples, for use in characterizing conditions in reference wetlands, identifying the relative degree of past disturbance of a prairie wetland, or assessing the current inhibition of key processes:

  • Richness of species and functional s (per unit volume of sample, or per thousand randomly-chosen individuals).
  • Number and biomass of cells per unit sample, or chlorophyll-a concentration per unit volume.
  • Proportional density and richness of species reputedly tolerant to a named stressor.
  • Degree of temporal variability in richness, density, and/or biomass (expressed as coefficient of monthly or weekly variation).
  • Month during which maxima of richness, density, and/or biomass occur (generally, or for particular s, e.g., blue-green algae, benthic algae).
  • Litter decomposition rate (or less precisely, the depth of fibric litter and proportional weight of dead vascular plant vegetation), used as an indicator of microbial activity.
  • Denitrification Enzyme Activity (DEA).

The specific ways some of these metrics have been or could be interpreted as an indication of stressed conditions are described in Section 2.4.

In situ methods of measuring algal production (e.g., uptake rates of carbon radioisotope, oxygen production) are not considered here. These methods are described generally by Stevenson and Lowe (1986), and protocols applicable to prairie wetlands are given in Hooper and Robinson (1976), Gurney and Robinson (1989b), and Britton and Greeson (1988). Estimates of algal production are strongly influenced by the choice of method used to estimate production, and temporal variability of measurements can be enormous.

2.3 Previous and Ongoing Monitoring in the Region

Field studies of algae have been published in at least 19 papers covering at least 124 prairie wetlands. A survey of five wetlands by LaBaugh and Swanson (1988) sampled only the water column, whereas surveys of about 50 wetlands by Kling (1975), and two wetlands by Shamess et al. (1985) included benthic and epiphytic algae as well. Information on species composition of metaphytic, epiphytic, and benthic algal communities also is very limited. Apparently no published studies have described the species composition of microbial communities in prairie wetlands.

Most algal and microbial studies in the region have considered only the physiological processes (e.g., respiration, production) associated with algae or microbial communities in general. Almost without exception, these studies have been completed during a single year's growing season, and in nearly all cases in permanent prairie wetlands or lakes. In a few cases (Barica 1975, Barica et al. 1980, Hickman and Jenkerson 1978), these process-based studies have incidentally mentioned the dominant taxa that were found. Identified assemblages of microbes in the region apparently have not been cultured in the laboratory to determine their functional characteristics, e.g., potential as denitrifiers or their role in methanogenesis. Decomposition processes have been measured in a few instances, generally without identifying the microbial taxa responsible (e.g., Davis and van der Valk 1978a,b, Neely and Davis 1985b, Wrubleski et al. 1993). Two studies (LaVeglia and Dahm 1974, Johnson 1986) measured process rates in the microbial communities of a few prairie wetlands using various physiological and chemical indicators: respiration, monophosphatase activity, electronic transfer system (dehydrogenase) potential, glucose mineralization, ammonium production, nitrification, and sulfur oxidation. Taxa responsible for the measured processes were not identified.

Montana's water quality monitoring agency currently is using epiphytic and benthic algae (specifically, diatoms) on a trial basis as an indicator of the condition of about five prairie wetlands that represent varying degrees of potential impairment. The information collected on species composition could yield valuable insights into sampling variability and habitat relationships. Ongoing research on algae in prairie wetlands by the National Hydrologic Research Institute in Saskatoon, Saskatchewan, is focusing on (a) transfer of algal energy to zooplankton and (b) effects of herbicides on algae (Appendix K).

2.4 Response to Stressors

2.4.1 Algal and Microbial Communities as Indicators of Hydrologic Stressors

Declining wetland water levels often raise the water temperature and concentrate dissolved nutrients that exist in the water column, and mobilize some of the nutrients from shoreline sediments and from plant litter that has become exposed. Such increases in nutrient concentration sometimes cause algal blooms in the remaining surface water (Schoenberg and Oliver 1988). Inundation can have the opposite effect, sometimes diluting and chemically binding nutrients to bottom sediments, cooling the water column, increasing algal competition with vascular plants, and thus reducing biomass of some algal taxa. However, inundation typically increases the leaf surface area available for colonization by some algae, and provides increased opportunities for dispersal of some algal taxa into and out of a wetland.

Species Composition

Changes in the density of phytoplankton -- as compared with metaphytic, epiphytic, and benthic algae -- might suggest that water levels in a wetland have changed within recent days or perhaps weeks. Specifically, recent (within a year) inundation often decreases the ratio of phytoplankton biomass (per unit area) to biomass of the other algal community components (Hosseini and van der Valk 1989a,b). This occurs as higher water levels reduce canopy coverage of vascular plants, increase light penetration and the area of substrate available for colonization, and dilute the levels of nutrients that otherwise would support the proliferation of rapidly-growing phytoplankton (Hooper and Robinson 1976, Gurney and Robinson 1988). Increased water levels can also differentially reduce phytoplankton density and productivity by creating habitat space for zooplankton, which graze selectively on the phytoplanktonic forms of algae. During the first year after flooding of one wetland, the biomass and productivity of both metaphyton and attached algae increased, whereas only the metaphyton continued this increase into the second year (Hosseini and van der Valk 1989a,b). Long-term changes in wetland water regimes might be inferred by collecting diatom remains using sediment cores (see Section 2.6.3) and determining what proportion of the found species (or pigments) are ones that are characteristically associated with drought or wet conditions, as inferred from salinity tolerances given for 143 taxa by Fritz et al. (1993) and for 62 taxa by Blinn (1993).

Species Richness

Data are insufficient to characterize the response of algal richness to changes in water levels of prairie wetlands. Sampling of five wetlands in the Cottonwood Lakes area identified 245 taxa in the two semipermanent wetlands, 159 in two seasonal wetlands, and 98 in a saline wetland (LaBaugh and Swanson 1988). The collective list from all five wetlands totalled 306 taxa, and 80% of these were present in the two semipermanent wetlands, 52% in the two seasonal wetlands, and 32% in the saline wetland.

Density, Biomass, and Productivity

Because flooding creates additional habitat space for both planktonic and epiphytic algae, the total biomass and production of algae in a wetland can increase, even when production of phytoplankton per unit area drops (Hosseini and van der Valk 1989a,b). Thus, total algal biomass or productivity of a prairie wetland, even if it could be measured accurately, would be a confusing indicator of water level change.

Decomposition

One of the few studies conducted on this topic in a prairie wetland (Wrubleski et al. 1993) found that leaching and decomposition of aboveground litter was more rapid in six of eight plant species in a flooded wetland than in a dry wetland; only the litter of cat-tail (Typha) and common reed (Phragmites) showed no difference between wet and dry treatments. The depth of flooding was inconsequential, but there were important differences in decay rate among taxa, with Chenopodium decomposing the slowest and Scolochloa and Scirpus lacustris the fastest. The nutrient content of litter from Phragmites actually increased over time, probably indicating the flood-related development of a rapidly growing microbial and epiphytic algae community, a phenomenon noted in other prairie wetlands as well (e.g., Neely and Davis 1985b). The relative extent of plant litter, as estimated coarsely from low-altitude photographs, was found to be a poor indicator of past hydrologic conditions in one prairie wetland (van der Valk and Squires 1992).

Other Microbial Processes

Hydrologic conditions affect denitrification, a microbial process that is of considerable importance because it improves water quality by removing excessive amounts of dissolved nitrogen. 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, e.g.:

soil water content was the most dominant factor controlling denitrification activity, followed by the concentration of ammonium, total soil respiration, and nitrate. (van Kessel et al. 1993)

Measurements of denitrification in a South Dakota wetland soil indicated that conditions of less than 22% volumetric soil moisture completely inhibit denitrification (Lemme 1988). A wetland does not have to be exposed to runoff for very long to reach these moisture levels and remove nitrate (i.e., convert and export nitrogen as a gas). Microbial communities that support denitrification develop rapidly in newly created wetlands (Duncan and Groffman 1994).

 

It remains unclear under which water regime denitrification is greatest. For example, Kantrud et al. (1989) state, "It would seem that temporary and seasonally flooded wetlands would be especially efficient in removal of excess nitrogen." There are at least two reasons why this might be so. First, fluctuating water levels that typify temporary and seasonal wetlands might be expected to enhance denitrification so long as (a) anaerobic conditions still occur, (b) moisture levels in the upper soil layers are not too severely depleted (i.e., pore space is 30-60% water-filled; Linn and Doran 1984, Lemme 1988), (c) carbon supplies also are not limiting (Fraser et al. 1988), and (d) salinity conditions are not extreme. Second, soil temperature might be expected to be warmer in temporary and seasonal wetlands during much of the year, due to their shallow depths.

However, other logic suggests that semipermanent and permanent wetlands might be more effective than temporary and seasonal wetlands for removing nitrate. Because semipermanent and permanent wetlands are usually groundwater discharge or flow-through systems, they are less susceptible to drought, and by definition, remain saturated and thus favorable to denitrification for longer periods. Prolonged drought in temporary wetlands not only results in moisture deficits inhospitible to denitrifying microbes, also can result in loss (via mineralization) of organic matter essential for sustaining denitrifiers. Organic matter content of soils in semipermanent and permanent wetlands generally seems to be greater than in temporary and seasonal wetlands (however, Loken 1991 reported less organic matter in soils of semipermanent groundwater discharge wetlands; he attributed this to high salinity of these basins inhibiting their productivity).

2.4.2 Algal and Microbial Communities as Indicators of Changes in Vegetative Cover

Species Composition

Algae and microbes respond quickly and persistently to changes in vegetative cover. As grazing, mowing, fire, and other factors decrease the amount of plant litter in prairie wetlands, the composition of algal and microbial communities can shift from characteristically epiphytic species to benthic or phytoplanktonic species. Long-term changes in plant cover of a wetland might be inferred by collecting diatom remains using sediment cores (see Section 2.6.3), and determining what proportion of the found species are ones that are characteristically shade-tolerant.

Species Richness

Species richness of algal communities sometimes declines with removal of vegetative cover (Seelbach and McDiffett 1983).

Density and Biomass

Algal and microbial biomass and density can either decrease (Rabe and Gibson 1984) or increase (Seelbach and McDiffett 1983) as vascular plant cover becomes sparser.

Decomposition, Other Microbial Processes

Decomposition rates can be retarded somewhat by wetland plant litter that has accumulated excessively (Godshalk and Wetzel 1978). However, microbial density and denitrification are generally greater in unplowed prairie soils than in plowed soils where plant litter is mostly removed (Linn and Doran 1984). Some rooted plants are capable of enhancing microbial populations and processes by (a) transferring nitrates from the sediment into aboveground tissues and eventually into the water column; (b) providing a carbon substrate (e.g., plant litter); (c) speeding the diffusion of oxygen (via roots) into otherwise anaerobic subsurface zones, especially during mid-growing season; and (d) increasing diurnal and seasonal fluctuations in the water table, and consequently the oxidation status, as a result of evapotranspiration. Densities of denitrifying microbes might be greatest where soil organic matter reaches a maximum just below the soil surface, but above the depth limit of the root zone (Parkin and Meisinger 1989). In this zone, impeded lateral flow increases the time available for nitrate loads to interact with prolific microbial populations present in the surrounding root masses.

2.4.3 Algal and Microbial Communities as Indicators of Wetland Salinity

Species Composition

In the prairie region, a salinity threshold of about 1000 mg/L separates algal species that are relatively salt-tolerant from ones that are not (Prepas and Trew 1983). Long-term changes in salinity of a wetland might be inferred by collecting diatom remains using sediment cores (see Section 2.6.3), and determining what proportion of the found species are ones that are characteristically salt-tolerant. Salinity limits and optima for 142 diatom taxa found in inland lakes of North America are presented by Fritz et al. (1993) and Blinn (1993), and might be used for reconstructing past salinity conditions in a wetland.

Species Richness

In saline lakes of western North America, the richness of diatom taxa is negatively correlated with specific conductance, with greatest richness corresponding to specific conductance of less than 45 mS (Blinn 1993). Diatom richess is greatest in waters where specific conductance is primarily the result of NaCl, or where concentrations of MgSO4 are intermediate, rather than where carbonate ions are dominant (Blinn 1993). Because local groundwater regimes play a major role in determining the ion chemistry of prairie wetlands, slight changes in groundwater flow might noticeably alter diatom species composition and richness.

Density, Biomass, and Production

Algal productivity increases with conductivity up to about 3000 µS/cm, and decreases at higher salt concentrations (Reynolds 1979). Moreover, chlorophyll-a, an indicator of algal biomass, occurs at lower concentrations in highly saline prairie lakes than in fresher ones, i.e., ones with < 1000 mg/L total dissolved solids (Barica 1978). An empirical model is available for predicting the nutrient status of saline prairie lakes, given information on their conductivity and chlorophyll-a content, but the model is not accurate where the ratio of total nitrogen to total phosphorus is < 12 (Bierhuizen and Prepas 1985, Campbell and Prepas 1986).

Decomposition

Apparently no studies of salinity effects on decomposition have been conducted in prairie wetlands. One study elsewhere found that decomposition was slower in inland wetlands having greater salinity (Reice and Herbst 1982).

2.4.4 Algal and Microbial Communities as Indicators of Sedimentation and Turbidity

A relatively low biomass and density of algae and microbes can indicate wetlands that receive chronically elevated inputs of sediment. Even more indicative might be a shift in species composition.

2.4.5 Algal and Microbial Communities as Indicators of Excessive Nutrient Loads and Anoxia

Species Composition

Algae and microbes respond more quickly to nutrient additions than do submersed vascular plants (Crumpton 1989). Among algal assemblages in prairie wetlands, phytoplankton and epiphyton respond immediately to small, repeated nutrient additions, whereas metaphyton demonstrate a delayed but large and enduring response, and benthic algae respond hardly at all (Murkin et al. 1994a). When nutrients are added in only a single dose, phytoplankton show a stronger response than epiphyton (Gabor et al. 1994).

Species composition of algal communities (especially diatoms) has a long history of use as an indicator of the relative state of enrichment of a water body. Moreover, algal species composition reflects not only the total level of nutrients, but also the ratio of two nutrients, phosphorus and nitrogen. One study of a prairie wetland (Barica et al. 1980) showed that a large ratio of biomass of green algae (Chlorophyta) to blue-green algae (Cyanophyta) can indicate that an oversupply of nitrogen, relative to phosphorus, has occurred within a few months. This pattern has been supported by wetland studies in other regions (e.g., Michigan bogs, Hooper 1982) that found that Euglenophytes (one-celled, mobile green algae) in particular respond to increases in ammonium and Kjeldahl nitrogen (rather than to nitrate alone), as well as to other substances associated with decomposing organic matter. The indicator status of a large variety of algal taxa with regard to enrichment is cataloged in several publications (e.g., Prescott 1968, Lowe 1974, Leclercq and Maquet 1987, Descy and Coste 1990, Richardson and Schwegler 1986). From such listings, the long-term changes in nutrient status of a wetland might be inferred once diatom remains or pigments from sediment cores (see Section 2.6.3) are collected and analyzed to determine the proportion of the found species (or pigments) that are characteristically associated with eutrophication.

Among microbial assemblages, photosynthetic microbes appear to respond more immediately to nutrient additions than do most other microbes (Pratt and Cairns 1985a). In some wetlands, enrichment increases the number of facultative-anaerobic bacteria (e.g., Streptococci, Enterobacteriaceae and aerobic spore forms, e.g., Bacillus spp., Pseudomonas alcaligenes, and Aeromonas spp.). Mesotrophic ponds can have elevated numbers of fluorescent pseudomonads, whereas oligotrophic waters can have more denitrifiers (Pseudomonas fluorescens and Vibrio spp.) (Schmider and Ottow 1985).

Species or Form Richness

Algal or microbial species richness is not a precise indicator of enriched conditions because it can either increase (e.g., Morgan 1988, Pratt et al. 1985) or decrease (e.g., Hooper 1982, Schindler and Turner 1982) in response to nutrient addition.

Biomass or Density

Algal (Murkin et al. 1991) and microbial (Tate and Terry 1980, Schmider and Ottow 1985) biomass or density are strong indicators of a wetland's degree of enrichment. Increasing duration and frequency of algal blooms can be a sign of increasing enrichment of a wetland.

Decomposition

Microbial populations, and consequently decomposition, are at least temporarily accelerated by enrichment in some wetland types (e.g., Dierberg and Ewel 1984). However, it is not apparent that relatively high rates of decomposition are a sign of atypical enrichment in prairie wetlands, and over the long term, enrichment could reduce decomposition rates in a wetland if it results in anaerobic conditions becoming widespread.

Other Algal Indicators of Enrichment

In prairie wetlands, Hooper-Reid and Robinson (1978a) found statistical relationships between nutrient enrichment (or impoverishment) and various physiological indicators: alkaline phosphatase activity, nitrogenase activity, ratio of protein to carbohydrate and lipid, and silica uptake rate. The strength of these relationships varied within the growing season, and in contrast, Murkin et al. (1994a) found no such statistical relationships. Formation of polyphosphate bodies within algal cells has also been used as an indicator of phosphate oversupply (Stevenson and Lowe 1986). However, many anatomical and physiological approaches are relatively labor-intensive and are often more appropriate for use in research than in routine monitoring.

Other Microbial Processes

The activity of denitrifying microbes is probably greater in wetlands of greater fertility (e.g., moderately alkaline clays with adequate organic matter). For example, microbial biomass in North Dakota soils was found to be greater in areas underlain by siltstone than in areas underlain by less fertile sandstone or shale parent material (Schimel et al. 1985). Denitrification rates might be greater in wetlands that have been exposed to nutrient runoff than in relatively pristine wetlands (personal communication, J. Kadlec, Utah State Univ., Logan). Tillage and fertilization of soils over time also might increase the suitability of remaining soil carbon as an energy source for denitrifying microbes (Groffman et al. 1992).

2.4.6 Algal and Microbial Communities as Indicators of Pesticide and Heavy Metal Contamination

Species Composition

Algal blooms commonly occur in wetlands following the application of herbicides to kill vascular plants. Benthic algae sometimes are the first to increase, as they benefit from the opening of the canopy. By stabilizing bottom sediments somewhat and thus reducing turbidity, their establishment can pave the way for metaphyton such as Chara, which can reduce turbidity even further (Crawford 1981). A shift in community composition from large filamentous chlorophytes (green algae) to smaller diatom species and blue-green algal species -- particularly those of the order Chaemaesiphonales -- is another possible sign of herbicide effects on a wetland (Goldsborough and Robinson 1983, Gurney and Robinson 1989a, Hamilton et al. 1987, Herman et al. 1986). However, whether this occurs can depend on the particular herbicide that is applied. Limited data from laboratory assays (Peterson et al., in press) suggest that (a) glyphosate might differentially inhibit diatoms, a key food of snails and midge larvae; (b) diquat might cause a shift from diatoms and blue-green algae to unicellular green algae; and (c) atrazine, hexazinone, simazine, and tebuthiuon might allow nuisance filamentous blue-green algae to become more dominant than other algal assemblages.

Effects of heavy metals and selenium on algae and microbes have been studied elsewhere (e.g., Crane et al. 1992), but have received little study in prairie wetlands. Algal taxa that might be potential indicators of heavy metal contamination are identified in several studies from other regions (Lange-Bertalot 1979, Maeda et al. 1983, Deniseger et al. 1990), and microbial taxa that are potential indicators of contaminants are documented by Baath (1989) and Dean-Ross and Mills (1989).

Species Richness

Algal and microbial species richness is probably a weak indicator of wetland contamination with toxic substances, but remains untested in prairie wetlands. Microbial diversity sometimes declines with exposure to hydrocarbon pollutants (Atlas et al. 1991) but not necessarily in response to heavy metals (Dean-Ross and Mills 1989).

Biomass and Density

In response to contaminants, the total biomass or density of algae and microbes can either decrease (e.g., Whitton 1971) or increase. Decreases are due generally to inhibition of reproduction and growth, whereas increases typically occur when contaminants are differentially toxic to animals that otherwise would graze on algae (e.g., Hurlbert et al. 1972), or to plants whose shading otherwise limits algal growth. Algae that inhabit sediments (benthic algae) appear to remain inhibited by some pesticides for a longer period than are algae that are attached to substrates above the sediment surface (Gurney and Robinson 1989a). This suggests that the ratio of benthic species to non-benthic species might be a useful indicator of persistent, sediment-adsorbed contaminants. However, the total biomass or density of algae and microbes is a poor indicator of contamination. This is especially true if the numbers of cells, rather than their volume, is the monitored indicator (Gurney and Robinson 1989a).

It cannot be assumed that contaminants that are harmless or harmful to vascular plants will usually have the same effect on algae. Many alga taxa are more sensitive than vascular plants to particular contaminants, especially those that inhibit photosynthesis (Fletcher 1990). Effects of contaminants on algal and microbial communities of prairie wetlands specifically have only recently been studied, beginning with a mesocosm study by Johnson (1986). A widely used herbicide, atrazine, was found to reduce algal productivity and growth by more than 40% when present at concentrations > 1 mg/L (Johnson 1986). Some evidence suggests that atrazine concentrations as low as 0.001 mg/L might be capable of altering algal species composition (deNoyelles et al. 1982) and biomass (Herman et al. 1986); effects may depend on duration of exposure (Jurgensen and Hoagland 1990). Some attached algae can develop resistance to atrazine after exposure to 0.050 mg/L (Detenbeck et al. 1993). Atrazine concentrations of up to 0.008 mg/L were found in a survey of 42 prairie wetlands in nine South Dakota counties (R. Ruelle, personal communication, USFWS, Pierre, SD), and concentrations of 0.001-.005 mg/L occur most of the time in agricultural streams entering the Great Lakes (Frank et al. 1979). A concentration of 0.413 mg/L would be expected to occur immediately after a 0.5-ha prairie wetland is sprayed at recommended dosages (Sheehan et al. 1987). Reviewing other toxicity data, Sheehan et al. (1987) concluded that the expected in-wetland concentrations of 7 of 21 herbicides used in the prairie region could be toxic to algae if wetlands were sprayed directly.

Recent laboratory testing of 23 pesticides (20 herbicides, 2 insecticides, 1 fungicide) at realistic, environmentally expected concentrations resulted in impacts to a wide range of alga taxa from nine of the pesticides, five of which were triazine herbicides (Peterson et al., in press). Least damaging to algae was the fungicide propiconazole, and the herbicides picloram, bromoxynil, and triclopyr. Field assays indicate triclopyr might be relatively nontoxic to wetland vascular plants as well (Gabor et al. 1993). Johnson (1986) also found that two other herbicides (triallate and treflan) actually stimulated photosynthetic productivity by 20-30% two weeks after application. However, triallate can be highly persistent under some conditions (Sheehan et al. 1987), and long-term effects were not determined. Carbofuran also mildly stimulated algal growth when present at concentrations of 10 and 100 mg/L. Phorate showed no effects, and fonofos inhibited algal growth only after 30 days, suggesting that a degradation product was responsible for toxicity. After applying another popular herbicide, glyphosate (Roundup), to prairie wetland mesocosms at a typical rate (2.5 L/ha), Shaw (1992) also reported a mild stimulatory effect on phytoplankton productivity at concentrations < 0.1 mg/L. However, greater concentrations depressed algal productivity (as measured by 14C uptake) in a group of four wetlands, and the author noted that lower concentrations of glyphosate could be just as toxic to algae in waters of relatively low calcium and magnesium content. In a survey of 10 other Saskatchewan potholes, Shaw (1992) found a glyphosate concentration > 0.1 mg/L in only one.

Decomposition

Even when applied at concentrations 50 and 100 times normal field rates, one soil insecticide (AC 92,100) had no apparent effect on decomposition rates in a prairie hydric soil (LaVeglia and Dahm 1974). No data are available for other pesticides.

Other Microbial Processes

A dosing study in a prairie wetland mesocosm of six herbicides (atrazine, fonofos, carbofuran, phorate, treflan, triallate) found that none had a significant impact on indicators of microbial functions (glucose mineralization, oxygen consumption, alkaline phosphatase activity, respiratory electron transfer system/dehydrogenase activity) (Johnson 1986). Similarly, an insecticide dosing study of an Iowa hydric soil found no impacts on some other indicators of microbial function (LaVeglia and Dahm 1974). Applying 50-100 potentially toxic contaminants to microbial communities and a variety of other organisms in laboratory bioassays, Blum and Speece (1991) found that chemicals that were highly toxic to a popular test organism -- fathead minnow -- were almost always toxic to a major denitrifier, Nitrosomonas. Two microbial assemblages responsible for decomposition (aerobic heterotrophs and methanogens) were less sensitive to the same contaminants.

Bioaccumulation

Apparently no studies have examined the role of algal and microbial communities as sinks for heavy metals or pesticides in prairie wetlands.

2.5 Monitoring Techniques

2.5.1 Direct Sampling

Methods for monitoring algal or microbial communities are described by Stevenson and Lowe (1986), Aloi (1990), and Britton and Greeson (1988). Microbial communities, especially assemblages of bacteria, are notoriously difficult to characterize because of the selectivity of culture techniques (Atlas 1984). Nonetheless, various bacterial strains can be placed in assemblages that likely have ecological significance (Mills and Wassel 1980).

Algae can be sampled at any season, but algal biomass is often greatest during the later part of the growing season (e.g., Crumpton 1989, Hooper 1978, Hooper-Reid and Robinson 1978a). In semipermanent wetlands, it may be advisable to sample metaphyton during sunny weather, because sunlight makes the metaphyton mats more buoyant and thus easier to see and sample.

Chlorophyll-a is sometimes sampled from the water column as an indicator of algal biomass. Some studies in prairie wetlands (e.g., Hickman and Jenkerson 1978) show it being only weakly correlated with measures of algal biomass (dry weight) and productivity, while others (Hosseini and van der Valk 1989a,b) report strong correlation. Cell volume is also sometimes used as an indicator of production (e.g., Shamess et al. 1985), but cell surface area seems to be a more accurate surrogate (Hooper-Reid and Robinson 1978b).

Algal communities in wetlands are generally collected from sediment samples, water column samples, artificial substrates, or natural organic substrates. Methods are:

Sediment sampling. Algae and microbes can be sampled from sediment surfaces in all types of prairie wetlands. Piston corers, plastic syringes, or other suction devices can be used. For example, Shamess et al. (1985) used a plexiglass corer to remove the top 2 cm of sediment when sampling benthic algae in a Manitoba wetland. In Florida cypress swamps, Dierberg and Brezonik (1982) sampled the nitrifying bacteria of surface sediments using a sterile piston corer and a plastic syringe with an attached tube.

Water column sampling. Whenever standing water is present for more than a few days, algae and microbes can be counted from samples of the water column of prairie wetlands (Robarts et al. 1992). Volumetric tube containers (Gurney and Robinson 1988) or fine-mesh nets have been used to collect samples. Vertically-integrating, automated samplers also can be used (Schoenberg and Oliver 1988). Surface microlayers (top 250-440 micrometers) can be sampled using fine nets or screens mounted on a frame (Estep and Remsen 1985).

Artificial substrates. Artificial substrates (initially sterile materials placed in a wetland and subjected to natural colonization) sometimes integrate the algal and microbial assemblages from a large variety of microhabitats (Henebry and Cairns 1984, Goldsborough and Robinson 1986). Microbes or algae can be monitored by installing plexiglass plates or similar inert, sterile surfaces in prairie wetlands at any time when surface water is likely to be present for several days. The substrates are colonized by attached algae and microbes during this period, and then retrieved for analysis. In prairie wetlands, cellulose acetate substrates roughened with sandpaper were used by Hooper and Robinson 1976, Hooper-Reid and Robinson 1978a,b, and Shamess et al. 1985; whereas acrylic rods were used by Hosseini and van der Valk 1989a,b, and Murkin et al. 1992.

Natural substrates. Epiphytic and benthic algae can be sampled using a quadrat approach, in which a frame is placed over a standard-sized area of bottom and substrates are scraped (Hooper and Robinson 1976). Frame sizes of 10 x 10 cm (Atchue et al. 1982) and 1-2 m2 (Schoenberg and Oliver 1988) have been used in other regions.

To accurately estimate algal and microbial density, the surface area of substrate must be quantified. This can be a daunting task in the case of epiphytic algae, where plant surface areas need to be measured. Some investigators have approached this by measuring surface areas of a random sample of plants, sometimes with the use of a digital scanner, then measuring their volumes (by displacement) or dry weights and developing area-volume or area-weight calibration curves. The curves can be used to estimate plant surface area from future, simpler measurements of the volume or weight of other plants of the same species.

Bacterial and fungal abundance are usually estimated as colony forming units (CFU) using plate count techniques. However, concerns have been raised about the validity of this technique for monitoring fungi; use of low-nutrient culture media (rather than the typical enriched media) are also recommended (Baath 1989).

Use of more than one sampling method is recommended, because different taxa occupy different habitats. For example, the data of Shamess et al. (1985) indicate that species richness of one prairie wetland ranged from 18 to 35 species, and that of another ranged from 26 to 41 species, depending on which of five components of the algal community were sampled, and how they were sampled(1)

. When results of sampling all five components were pooled, species richness of the first wetland was 78 and that of the second was 80; thus, no single component of the algal community contained more than 45-51% of the species. The number and proportion of species that were unique to one component of the algal community was smallest for the smooth substrate-colonizing component (1 endemic species, comprising 4% of all species on smooth substrate) and greatest for the phytoplankton component (17 endemic species, comprising 41% of all phytoplanktonic species).

2.5.2 Indirect Sampling Through Measurement of Processes

Decomposition

Procedures for measuring rates of decomposition in prairie wetlands are detailed by Murkin et al. (1989) and Davis and van der Valk (1978a,b). In the latter instance, the authors collected fresh standing litter immediately after first frost. They clipped six, 1 x 1 m quadrats of each species, at 15-m intervals along a transect parallel to shore. They placed plant litter in nylon mesh bags on the sediment surface or suspended in the water column. They then removed a few bags periodically for about one year (more often at first). Silt and invertebrates were removed and samples were dried to a constant weight. The investigators noted that, by excluding litter-processing invertebrates, the bags might not precisely represent the natural rate of decomposition.

Denitrification Enzyme Activity (DEA)

A method for determining the relative activity level of important denitrifying bacteria in soils was applied to wetlands by Groffman and Tiedje (1989). Requirements include a laboratory with a gas chromatograph, a gas manifold (to make samples anaerobic), and facilities to do chloroform-incubation methods of carbon analysis, chloramphenicol microbial inhibition, and nitrogen gas measurement, from samples brought in from the field. The initial laboratory investment is approximately $20K, and exclusive of the gas analysis tasks, one person can run 50-100 samples per day, with a cost of $150/month for expendable supplies (P. Groffman, personal communication, Institute of Ecosystem Studies, Millbrook, NY).

Other Measures That May Reflect Microbial Processes

Respiration and other functional activities of microbial communities can be estimated by a variety of indirect methods. Methods for measuring respiration of entire ponds or wetlands are available (Madenjian et al. 1990). Probably the best-known microbial bioassay technique is the Microtox Standard Assay Procedure, which has been used to measure microbial stress in prairie wetlands potentially exposed to pesticides (Ruelle and Henry 1993). Measurements of the relative rates of lipid biosynthesis (Fairchild et al. 1984) are another expression of microbial function. Stressed microbial communities also sometimes have altered adenylate (ATP, ADP, AMP) energy charge ratios. Microbial biomass can be indirectly monitored by comparing levels of adenosine triphosphate (ATP) to ash-free dry weight (Meyer and Johnson 1983). The rates at which microbial communities colonize sterile substrates introduced to a wetland, and the characteristics of the colonizing community, can also be used to indicate impacts from contaminant toxicity (Cairns et al. 1992).

2.5.3 Analysis of Historical Conditions

The nutrient status of a wetland during pre-settlement periods can sometimes be inferred by examining photosynthetic pigments or structural remains found in wetland sediments. In particular, levels of chlorophyll, chlorophyll derivatives, carotinoids, and myxoxanthin (a pigment associated with eutrophic blue-green algae) can be used to infer nutrient status, at least in permanent water bodies where past hydrological effects on species composition can be presumed to be insignificant. Samples are collected with corers; pigments can then be extracted with solvents and partitioned using chromatographic methods. Methods are described by Leclercq and Maquet (1987) and Agbeti and Dickman (1989). A baseline was established using such an approach in one prairie wetland (Begres 1971), and a project involving analyses of cores from 50 prairie wetlands is ongoing (S. Fritz, personal communication, Limnological Research Center, University of Minnesota, Minneapolis).

2.5.4 Bioassay Methods

A review of laboratory, outdoor mesocosm, or in situ bioassay methods involving algae is beyond the scope of this document. Use of bioassays to explore contaminant toxicity to algae in prairie wetlands has been relatively limited. Examples include studies by Gurney and Robinson (1989a), Wayland and Boag (1990), Johnson (1986), and Ruelle and Henry (1993). Impacts of phorate, an organophosphate insecticide, on microbial populations were not detected using a culture test, the Microtox test (Dieter et al. 1994).

2.6 Variability and Reference Points

2.6.1 Spatial Variability

Species Richness

One of the few studies to survey algal richness in prairie wetlands (LaBaugh and Swanson 1988) sampled pothole wetlands representing five different hydrochemical environments in the Cottonwood Lakes area. When lists from all five wetlands and all six sampling dates (months) were pooled, the species total was 306 -- clearly more species than are usually found in any wetland's non-algal flora or fauna. Seventy-six algal species were found over the course of one season on three species of plants in a single shallow lake in Manitoba (Pip and Robinson 1982).

In other regions, studies that have compared protozoan communities among wetlands include Henebry et al. (1981) and Pratt et al. (1985). The former study, covering 13 Michigan wetlands over a 5-year period, found a range of 93 to 365 protozoan species. The latter study, covering 28 Florida ponds, found a range of 112 to 410 species, with a mean of 338 species in non-artificial ponds. Functional structure of the resident protozoan fauna changed slightly from year to year, but wetlands in the same geographic region and experiencing the similar climatic patterns had similar proportions of species in each functional (Pratt et al. 1985).

Density, Biomass, and Production

Phytoplankton standing crop (biomass) is often expressed as chlorophyll-a, and can peak at 0.481 mg/L in some prairie lakes (Barica 1975, Barica et al. 1980). Phytoplankton averaged only 0.029 mg/L during the growing season in one Saskatchewan lake (Hickman and Jenkerson 1978), 0.002-0.006 mg/L in six prairie wetlands (Gloutney 1993), 0.001-0.380 mg/L in the Cottonwood Lake semipermanent wetlands (LaBaugh and Swanson 1993), and 0.010 mg/L in a saline prairie lake (Campbell and Prepas 1986). Metaphyton standing crop in a prairie wetland that had been flooded for two years averaged 66 g/m2 over a growing season (Hosseini and van der Valk 1989b) and peaked at 151 g/m2 dry weight. Metaphyton in another prairie marsh was measured as 200 g/m2 (van der Valk 1986). Among six Saskatchewan prairie wetlands, the biomass of epiphytic algae peaked at only 0.025 to 0.105 g/m2 (Gloutney 1993). In Delta Marsh, epiphytic biomass was estimated to range from 2.3 to 32.3 g/m2 (Hooper and Robinson 1976). Chlorophyll-a from Delta Marsh's epiphyton varied from < 0.01 to about 0.05 g/m2, and was greatest within a cat-tail stand,m from its edge with open water (Murkin et al. 1992). Chlorophyll-a from epiphyton in some saline prairie lakes in Alberta averaged less than 0.07 g/m2 (Campbell and Prepas 1986).

In shallow prairie lakes, phytoplankton densities can exceed 300,000 cells per ml (Hickman and Jenkerson 1978); bacterial densities can exceed 10,000,000/mL (Campbell and Prepas 1986). Phytoplankton primary productivity averaged 196.77 mg C/hr/m3 and 196.77 mg C/hr/m2 in a shallow prairie lake (Hickman and Jenkerson 1978), and in prairie wetlands the production of all algae combined generally ranges up to a few hundred gC/m2/yr (Crumpton 1989, Murkin and Batt 1987). The production of epiphytic algae is perhaps greater on emergent than submersed vascular plants (Hooper and Robinson 1976), although it is difficult to standardize estimates of available surface area of plants. Within emergent plant communities, the level of epiphytic algal biomass varies largely with spatial and temporal variation in nutrient availability, e.g., from 2.3 to 32.3 g C/m2 (Hooper and Robinson 1976, Hooper-Reid and Robinson 1978a).

Decomposition Rate

Few estimates are available to describe the variability of decomposition rates among and within wetlands. The half-life of fallen emergent plant litter in two Iowa prairie lakes ranged from 128 to 1011 days, depending on plant species, season, and other factors (Davis and van der Valk 1978a,b, Neely and Davis 1985b, Neely and Baker 1989).

Denitrification Enzyme Activity (DEA)

Measurements of DEA within a wetland vary somewhat spatially. Variability (coefficient of variation) of DEA measurements within wetlands ranges from 33-89%, and the coefficient of variation among true replicates is about 10% (P. Groffman, personal communication, Institute for Ecosystem Studies, Millbrook, NY). Measurements of denitrification using alternative methods are highly variable and difficult to compare (Seitzinger et al. 1994), but DEA measurements are believed to be generally less variable because they integrate conditions over time.

2.6.2 Temporal Variability

Species Richness

In the Cottonwood Lakes area, data indicate considerable monthly variability in algal richness (LaBaugh and Swanson 1988). Species turnover rates have not been quantified for prairie wetlands.

Density

In a study of five wetlands in the Cottonwood Lakes area, the following relationships were noted (LaBaugh and Swanson 1988):

Basin Type Month(s) with the Most Peaking Taxa (peaks) Month with the Fewest Peaking Taxa
Seasonal
(n = 2 basins)
June (38 and 42 taxa) May and July
Semipermanent
(n = 2 basins)
October (32 and 55 taxa) and May (59 taxa) June and September

Decomposition Rate

Interannual differences in decomposition rates and patterns of two species (Scolochloa and Scirpus lacustris) in a flooded prairie wetland were negligable (Wrubleski et al. 1993).

Denitrification Enzyme Activity (DEA)

No published information was found on interannual variation in prairie wetlands of DEA or other microbial functions.

2.7 Collection of Ancillary Data

It is easier to separate the anthropogenic from the natural causes of impairment of community structure if data are estimated or inferred simultaneously on the following features of particular importance to algae and microbes:

age of wetland and its successional status, light penetration (water depth, turbidity, shade), temperature, sediment oxygen, general chemistry of waters (particularly pH and conductivity), leaf surface area and stand density of associated vascular plants, density of grazing aquatic invertebrates, and moisture regime (e.g., time elapsed since last runoff, inundation, or desiccation event).

All of these features vary to a large degree naturally, as well as in response to human activities such as soil tillage, compaction, and erosion; fertilizer and pesticide application; and water regime modification.

2.8 Sampling Design and Required Level of Sampling Effort

For most algal communities, sample processing and species proportional counts (assuming 500 individuals) take 2-3 hours per sample (Stevenson and Lowe 1986).

2.9 Summary

The enormous diversity of algae (probably over 500 taxa in prairie wetlands) and the position of algae and microbes at the base of the food chain suggests their considerable ecological importance. It also highlights a need for monitoring of key processes supported by algae and microbes, and continued research to associate various rates of these processes (e.g., decomposition) with the seasonal sequencing and occurrence of particular species compositions. Published estimates of interwetland and interannual variability of algal and microbial taxonomic composition in prairie wetlands are nearly nonexistent.

Most algal and microbial communities recover quickly from acute disturbances. Because of this, direct sampling of algal and microbial communities will fail to detect many acute disturbances. Alternatively, indirect examination of the pigment from just the portion of the diatom community that accumulates seasonally in sediment traps might provide some indication of current wetland conditions.

Algal species composition, and to a lesser degree species richness, demonstrates diagnostic responses to changes in vegetative cover, salinity, excessive nutrient loads, and sedimentation/turbidity (Table 1). Algae also respond sensitively to changing water regime and pesticide/heavy metal contamination, but existing information is too limited and confounding effects are too prevalent to currently allow widespread use of algae to diagnose impairment of prairie wetlands from these two stressors.

Table 1. Summary Evaluations of Possible Algal and Microbial Indicators of Stressors in Prairie Wetlands.

Evaluations are based on technical considerations, not cost or practicality. A rating of FAIR or POOR is assigned when too few data (FD) suggest potential as an indicator, or when confounding effects (CE) of other variables often overshadow those of the listed stressor, with regard to effects on the indicator.

Stressors Possible Indicators Evaluation
Hydrologic conditions Species Composition
Richness
Density, Biomass, Productivity
Decomposition
Denitrification
FAIR (CE)
UNKNOWN FAIR (CE)
POOR
FAIR
Changes in vegetative cover conditions Species Composition
Richness
Density, Biomass, Productivity
Decomposition
GOOD
FAIR (FD)
POOR (CE)
POOR (FD)
Salinity Species Composition
Richness
Density, Biomass, Productivity
Decomposition
GOOD
GOOD
FAIR
POOR (FD)
Sedimentation & turbidity Species Composition
Richness
Density, Biomass, Productivity
Decomposition
FAIR (FD)
FAIR (FD)
FAIR (CE, FD)
POOR (FD)
Excessive nutrients & anoxia Species Composition
Richness
Density, Biomass, Productivity
Decomposition
Denitrification
GOOD
POOR
GOOD
POOR (CE)
FAIR (CE)
Herbicides Species Composition
Richness
Density, Biomass, Productivity
Decomposition
FAIR (FD)
POOR (FD)
FAIR (CE)
POOR (FD)
Insecticides Species Composition
Richness
Density, Biomass, Productivity
Decomposition
POOR
POOR
POOR
POOR
Heavy Metals Species Composition
Richness
Density, Biomass, Productivity
Decomposition
GOOD (CE)
POOR
POOR
POOR

1. 11      The methods were: acetate colonization substrate (smooth, roughened), epiphyton sampling (scraped from Typha stems), epipelon sampling (from core samples), and phytoplankton sampling (using a tube sampler). Each wetland was sampled 17 times during the growing season by each method.


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