3. Vascular Plants as Indicators of Prairie Wetland Integrity

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

Contents:

3.1 Ecological Significance and Suitability as an Indicator
3.2 Potential Indicator Metrics
3.3 Previous and Ongoing Monitoring in the Region
3.4 Response to Stressors

3.4.1 Vascular Plants as Indicators of Hydrologic Stressors
3.4.2 Vascular Plants as Indicators of Vegetative Cover Condition
3.4.3 Vascular Plants as Indicators of Wetland Salinity
3.4.4 Vascular Plants as Indicators of Sedimentation and Turbidity
3.4.5 Vascular Plants as Indicators of Excessive Nutrient Loads and Anoxia
3.4.6 Vascular Plants as Indicators of Pesticide and Heavy Metal Contamination

3.5 Monitoring Techniques

3.5.1 Ground-based Sampling
3.5.2 Aerial Methods
3.5.3 Potential or Historical Vegetation
3.5.4 Bioassay Methods
3.5.5 Bioaccumulation

3.6 Variability and Reference Points

3.6.1 Spatial Variability
3.6.2 Temporal Variability

3.7 Collection of Ancillary Data

3.8 Sampling Design and Required Level of Sampling Effort

3.8.1 General Considerations
3.8.2 Asymptotic Richness
3.8.3 Power of Detection

3.9 Summary

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3.1 Ecological Significance and Suitability as an Indicator

In prairie wetlands, three growth forms of vascular plants usually are defined: emergent, floating-leaved, and submersed plants. To varying degrees, these form discrete zones within wetlands. Wetlands containing all three forms (and many subforms and species) are generally those with a core area that is flooded permanently, but whose water levels otherwise vary greatly from year to year. In such wetlands, nutrients are more available and support substantial invertebrate densities and waterbird use.

The foliage and/or seeds of several species of vascular plants are consumed regularly by waterfowl (see Appendix A). The food values of wetland vascular plants are attributable both to their being consumed directly (mainly by microinvertebrates and tadpoles) and to their serving as an attachment surface and secondary energy source for algae, amphibian larvae, and invertebrates.

Also, as noted in Section 3.0, vascular plants as well as algae influence the fertility of prairie wetlands by harnessing solar energy through photosynthesis. In contrast to algae which release the stored solar energy almost immediately after their death, the energy from vascular plants is made available by microbes slowly, over a period ranging from weeks to months. Thus, during seasons when algal populations are at a minimum, the energy originally trapped by vascular plants could well be a significant source of energy for the invertebrates which later in the growing season are consumed by waterbirds (Nelson and Kadlec 1984). Of particular importance is the value of vascular plants in prairie wetlands as a substrate for growth of attached algae and microbes (Campeau et al. 1994), and as habitat cover (shelter) that protects invertebrates, amphibians, and birds from predators and severe weather (Murkin et al. 1992). Evidence from other regions (Hanson 1990) suggests that the type of wetland plant can influence the size structure of invertebrate communities (the ratio of large to small individuals), and thus perhaps the value of the invertebrate community to waterbirds. 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.

Vascular plants are also important because they influence the amount, rate, and seasonal timing of nutrient and contaminant cycling across the water column - sediment ecotone. Some plants remove nutrients directly from the water column, thus seasonally tieing up some of the nutrients that otherwise might support nuisance algal blooms. In the sediment, plant roots can take up nutrients (and contaminants). In some cases, plant roots can transfer both nutrients and harmful substances to plant foliage, making these substances more available to food chains. Wetland plants also help maintain wetland water quality by stabilizing shorelines and reducing wind-driven resuspension of sediments that otherwise impair light penetration and reduce primary productivity.

Characteristics of vascular plants that are usually an advantage for indicating wetland integrity include:

• Immobility: plants reflect site conditions and are practical for use in in situ exposure assessments.
• Interpretability of gross patterns of spatial distribution as indicators of condition; patterns are interpretable from a distance without requiring permission for access to private property (e.g., through interpretation of aerial imagery).
• Sensitivity to a wide variety of stressors (especially hydrology, salinity, and changes in vegetative cover); sensitivities of individual species are relatively well known; submersed plants are especially sensitive (e.g., to turbidity, overenrichment).
• Taxonomy known better than for any non-animal assemblage, and identification to genera is mostly straightforward.
• Sampling techniques and community metrics are well-developed.

Vascular plant characteristics that usually are considered disadvantageous for indicating wetland integrity include:

• Lagged response to stressors (episodic stresses may not be reflected).
• Relative insensitivity to insecticides and heavy metals (but can bioaccumulate these).
• Difficulty sampling some assemblages (e.g., submersed species).
• Laborious identification of some assemblages.
• Difficulty characterizing communities during the dormant season.

3.2 Potential Indicator Metrics

The following measurements and metrics were considered, as applied to plant communities, 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, functional assemblages, and rare species (per unit area, or per thousand randomly-chosen individuals) in extant communities and in seed banks.
• Number and biomass of stems per unit area, or cumulative shoot length, or canopy cover per unit area (when measured at the time of its annual maximum, this is commonly used as a proxy for annual plant production).
• Relative dominance and richness of species, particularly of species reputedly tolerant or intolerant to a named stressor (measured both in extant communities and seed banks).
• Interannual variability in richness, density, and/or biomass.
• Germination rate of seeds in sediment samples.
• Bioaccumulation of contaminants.

The specific ways some of these metrics have been or could be interpreted as an indication of stressed conditions are described in Section 3.4.1. Various in situ methods of measuring macrophyte production (e.g., uptake rates of carbon radioisotopes, oxygen production, photosynthetic rate, respiration) are not considered in this document. In some instances, whole-system repiration rates, integrated over a 24-hour period, might usefully indicate wetland functional integrity.

3.3 Previous and Ongoing Monitoring in the Region

Virtually all studies of prairie wetlands mention plant species that are dominant in the studied wetland. Studies focusing primarily on vascular plants number at least 59 and cover in excess of 2200 prairie wetlands (Appendix J). The metrics most commonly measured in plant surveys are relative abundance and percent cover (canopy density). Although long-term monitoring of mature vegetation has been conducted in the Cottonwood Lakes wetlands and perhaps other areas, apparently only eight long-term studies (> 7 years of data) have been published.

No state agencies responsible for prairie wetlands currently monitor plants as indicators of wetland ecological integrity. At a regional level, USEPA's EMAP has documented plant communities in 30-40 wetlands that span gradients of water regime, probable disturbance, and geography. Variables that are being measured include species richness, areal cover, cover ratio, and amount of standing dead litter.

At more local levels, species composition and density of plants are being tested for possible use as indicators of the success of wetland restoration efforts in Iowa (Galatowitch 1993a,b), Minnesota (Madsen 1988, Sewell 1989), and perhaps elsewhere. Research on ecological relationships affecting plant communities continues to be conducted at NPSC and at the state universities. Personnel from The Nature Conservancy and/or state Natural Heritage programs conduct botanical surveys of prairie wetlands in relatively undisturbed localities, and periodically check a few of the wetlands known to contain plant species of state, regional, or national importance due to their rarity. In wetlands known to contain regionally-rare species, monitoring of these species can be used to indicate long-term integrity of the wetland. In prairie pothole wetlands of the United States, examples of regionally rare species are Napaea dioica, Carex formosa, Eleocharis wolfii, and Astragalus neglectus (interpreted from information provided to USEPA in 1992 by The Nature Conservancy).

3.4 Response to Stressors

3.4.1 Vascular Plants as Indicators of Hydrologic Stressors

Species Composition

The species composition of vascular plant communities, especially the submersed forms, is an excellent indicator of hydrologic conditions that have occurred in prairie wetlands as recently as within the past 2-4 years (Stewart and Kantrud 1972b, Millar 1973, Weller and Voigts 1983, Weller et al. 1991, Squires and van der Valk 1992). As little as 2 cm of standing water in a wetland can result in development of a plant community that differs by 75% from the one that develops when surface water is absent (van der Valk and Davis 1978).

In general, a shift over a 2-3 year period toward increased dominance by submersed species and a decline in emergent species (both adult plants and propagules) can indicate recent periods of wetter-than-usual conditions (Weller and Spatcher 1965, Millar 1973, van der Valk 1981, Weller and Fredrickson 1974, van der Valk and Squires 1992, van der Valk et al. 1994). Conversely, reduced dominance of submersed and other "obligate" wetland species (or species that typify semipermanent wetlands, see Appendix F), in association with increased dominance of emergent and "facultative" wetland species that are often annuals or biennials (or species that typify temporary wetlands, see Appendix F), can be used to determine if relative drought conditions have occurred recently (van der Valk and Davis 1978, Pederson 1981, Poiani and Johnson 1989, Poiani et al. 1995). For a few species indexed in Appendix F, quantitative data on water depth ranges and related hydrologic variables also are available.

More specifically, van der Valk (1981) noted that recent hydrologic conditions of a prairie wetland could often be surmised from current species composition, and less recent conditions can be inferrred from the seed bank. Certain conditions can be deduced retrospectively by knowing the proportion of extant species that characteristically (a) are annuals, perennials, or vegetative reproducers (see Appendix A), (b) are mudflat-germinators vs. standing-water germinators, and (c) have seeds that remain viable for long periods of time, vs. those with short-lived seeds which depend highly on dispersal. For example, if few mud flat annuals (species that have long-lived seeds and can only become established when there is no standing water) are found in seed banks underlying open water, it suggests that standing water has probably persisted for many years. Under ideal conditions, seed bank analysis can be used to infer conditions that existed up to 70 years ago (Wienhold and van der Valk 1989). Unfortunately, autecological knowledge is insufficient to assign many prairie wetland species to the predictive categories (especially b and c above) or to describe their typical seed longevity. This limits the utility of this deductive approach.

If an attempt were made to define hydrologic criteria for protecting various kinds of prairie wetland plant communities or assemblages of species, definitions of optimal heights and durations of water levels (and their variability) would be two important components. Using data on surrounding landscape factors (runoff potential) and anthropogenic uses, quantitative estimates of expected water heights and durations could be made for wetlands within particular landscapes, and from this, the types of plant communities that should be present could be predicted (e.g., Poiani and Johnson 1993a,b,c, Poiani et al. 1995). However, a major limitation of implementing this approach currently is that numeric estimates of adult plant and seed tolerances for water depth and duration have been made for only a few prairie wetland species⁽¹⁾

. Complicating this further is the fact that hydrologic tolerance thresholds vary somewhat within species, as a result of genetic variation and confounding influences of season (Meredino et al. 1990, 1991) and chemical conditions at the time that flooding or drawdown occurred.

In general, flooding of prairie wetlands tends to have a greater effect on community composition than occasional drought (van der Valk and Squires 1992). The specific effects of hydrologic alteration depend on flooding depth, season, frequency, duration, initial water levels, sediment type, dominant plant species, and other factors. Nonetheless, vascular plants are potent indicators of long-term hydrologic change in wetlands.

Community Zone Locations

The presence, position, and heterogeneity of vegetation zones, each representing a recognizable plant community, can be used as an indicator of wetland integrity. For example, if a particular zone is absent but was expected based on knowledge of a wetland's hydrology, then some perturbation can cautiously be assumed to have occurred (van der Valk and Welling 1988).

 

When wet conditions continue for at least two years in wetlands that formerly had a temporary or seasonal moisture regime (i.e., they become semipermanent or permanently inundated), the shallower zones of emergent vegetation often die off first, leaving a central deep zone of emergents surrounded by open water (Millar 1976). At least in situations where water level increases are large and abrupt, emergent communities disappear rather than shift to shallower areas (van der Valk 1994). During droughts, emergents can become re-established in the center of a wetland basin. Overall, hydrologic changes usually causes only slight shifts in the position of the submersed, emergent, and meadow zones (Harris and Marshall 1963, van der Valk and Davis 1976, Johnson et al. 1987). Vegetation zone boundaries in larger prairie wetlands tend to be less distinct than in small wetlands (Johnson et al. 1987), so their use as an indicator can be restricted in large wetlands.

Species Richness (of Mature Plants)

Plant species richness is a complex indicator of prior hydrologic conditions. Low richness can indicate prior periods of drought (Driver 1977, Galatowitsch 1993a,b), partly because prolonged dehydration facilitates the formation of homogeneous stands by a very few species. Perhaps less often, low richness can indicate prior periods of inundation when floods drown many emergent species. In some instances, prairie plant communities that are initially species-rich tend to suffer less loss of productivity as a result of drought than communities that initially are species-poor (Tilman and Downing 1994). Semipermanent wetlands generally have greater plant species richness than temporary or seasonal wetlands of the same size. This is partly because water persistence and depth is sufficient to allow submersed and floating-leaved plants, as well as terrestrial forms, to exist and successfully reproduce.

When increased species richness results from wetter conditions, it is often because increased water levels tend to fragment the monotypic stands of emergent species (especially cat-tail, bulrush, pickerelweed, and allow partial invasion by submersed and floating-leaved species that add to diversity (Botts and Cowell 1988, McIntyre et al. 1988). Wetlands that are more persistently wet also can have higher usage by birds and other mobile animals that introduce seeds, thus diversifying the wetland plant community (Pip 1987b).

In part of Delta Marsh, reflooding of various units to different depths after a drawdown that had lasted one or two years had several effects (van der Valk and Squires 1992, van der Valk et al. 1994). Reflooding to original water levels greatly increased the variety of vegetation forms (class richness, and number of classes containing multiple species), whereas reflooding to water levels that were 1 m above the original resulted in a halving of species richness. Reflooding to intermediate water levels gave unclear results.

After a drawdown and gradual reflooding of an Iowa semipermanent wetland, the peak in total species richness occurredyears after the drawdown, whereas the richness of just the submersed and floating-leaved species did not peak until 4-5 years after drought. In their Delta Marsh studies, van der Valk and Squires (1992) reported a lag time of at leastyears before small (< 1 m) changes in wetland water level could be detected using wetland plants.

Species Richness (of Seed Banks)

The species richness of seeds in wetland seed banks (seeds that lie dormant in soils and sediments) can be used to crudely approximate the time elapsed since hydrologic alteration occurred in wetlands that now are dry. This requires calibration to conditions currently present in similar but hydrologically unaltered wetlands. For example, Wienhold and van der Valk (1989) found that mean species richness in a series of unaltered wetlands was 12.3, but fell to 7.5, 5.4, 5.0, 3.2, and 2.1 species in potholes drained up to 5, 10, 20, 40, and 70 years ago. After wetlands had been drained for more than 20 years, about 60% of the species disappeared from the seed bank.

Biomass, Production, and Cover Ratio

For emergent plants, decreases in aerial cover can signify recent inundation events (van der Valk and Squires 1992), and increases in aerial cover can indicate recent drought events. If periods of dehydration are brief (a few hours or days), the aerial extent of emergent plants, especially those with rigid stems (e.g., cat-tail, common reed) is unlikely to change. The presence of "hemi-marsh" conditions -- relatively equal proportions of open water and emergent wetland -- indicates that a wetland has probably undergone a wet-dry cycle of about 2-7 years. During such a cycle, droughts lasting more than one year (Welling et al. 1988a) and flooded conditions lasting at least one winter (Millar 1973, McKee et al. 1989) can damage or impair the recruitment of many wetland emergent plants. For submersed plants, flood conditions lasting at least two years are required for establishment of some species (e.g., Potamogeton pusillus, Millar 1976). In Iowa, prairie wetlands restored within the last year had significantly more floating-leaved plants than those restored two years previously (Hemesath 1991). Emergent plant cover was less, and open water area was greater, on restored wetlands that had been drained > 30 years ago than on restored wetlands that had been drained more recently.

Models are available for predicting the duration of floods and droughts within a wet-dry cycle that likely produced a particular number and configuration of wetland types (Woo et al. 1993, Larson 1995) and associated species assemblages (e.g., Poiani and Johnson 1993a,b,c, Poiani et al. 1995). Similarly, models are available for predicting future plant species composition, given a particular hydrologic change (Poiani and Johnson 1993a,b,c, van der Valk et al. 1989, USFWS 1993, de Swart et al. 1994). A computer simulation by Poiani and Johnson (1993a,b,c) indicated that for a particular depth class, the ratio of open water to wetland vegetation (the cover ratio) was always greater when water levels were held constant for five years than when they fluctuated during that time. Thus, a relatively large cover ratio might be considered potentially indicative of a relatively sustained input of water to a wetland.

Aerial cover of submersed and floating-leaved plants changes less with increased inundation than does the aerial cover of emergent vegetation. However, cover of non-emergent species is usually reduced somewhat due to increased wave action and turbidity (van der Valk and Davis 1978). For example, although flooding to 1 m above normal in Delta Marsh, allowed bladderwort (Utricularia vulgaris) to invade stands of cat-tail, whitetop, and common reed, the flooding completely eliminated beds of sago pondweed (Potamogeton pectinatus) (Murkin et al. 1991). At the opposite extreme, complete drawdown lasting for much of a growing season has catastrophic effects on most submersed species. However, a few submersed and floating leaved species can survive up to a year of dessication e.g., Callitriche palustris, Potamogeton gramineus, Myriophyllum spicatum, Lemna minor, Spirodela, Naias, Potamogeton pectinatus, Marsilea mucronata, and Ceratophyllum demersum (Stewart and Kantrud 1972b, Cooke 1980, Davis and Brinson 1980, Nichols et al. 1989, van der Valk and Davis 1978).

Most studies of flooding effects in prairie wetlands have focused on semipermanent wetlands. A study of seasonal wetlands found that production (especially aboveground production) of one emergent species, Scholochloa festucacea, increased in response to 10 weeks of flooding to a depth of 25 cm in the spring (Neill 1994).

Seed Density

The density of seeds in wetland seed banks can be used to crudely approximate the time elapsed since hydrologic alteration occurred in wetlands that now are dry. This requires calibration to conditions currently present in similar but hydrologically unaltered wetlands. For example, Wienhold and van der Valk (1989) found that mean total seed density in a series of unaltered or recently (within five years) altered wetlands was 3600-7000/m², but declined to 1400, 1200, 600, 300, and 160 seeds/m² for the 10, 20, 30, 40, and 70 years after drainage.

Germination Rate

Seedlings of most emergent species in prairie wetlands fail to germinate from areas where sediments are not exposed until after early July. Germination is particularly hindered if the previous growing season had been especially wet, thus supporting exceptional biomass of submersed plants (as well as of emergent species) whose residual litter can smother emergent seedlings (van der Valk 1986). Consequently, the presence of viable emergent plant seeds but not seedlings sometimes can indicate that a drawdown, if it occurred, did not occur until late in the preceding growing season (Welling et al. 1988b, Meredino et al. 1990), that the wetland was overgrown with submersed plants and filamentous algae the preceding year; and/or that conditions prior to the growing season dramatically affected microbial and invertebrate communities that normally decompose most litter.

3.4.2 Vascular Plants as Indicators of Changes in vegetative cover Condition

Species Composition

Species composition can be used to loosely indicate the severity and type of process that has removed vegetation in a prairie wetland. A dominance of relatively tall and robust, adventive species or their hybrids, as opposed to shorter emergent species, is one possible sign of a lack of periodic disturbance from livestock, burning, mowing, or cultivation (Stewart and Kantrud 1972a). A relative scarcity of highly palatable (to cattle) plant species can also signify that intensified grazing has occurred during drier years. Plants that are annuals (see Appendix A) tend to be the most affected by early-season mowing (Stewart and Kantrud 1972a). Other emergent species tend to be affected differentially by herbicides and fire (Thompson and Shay 1985). Information on shifts in community composition that can result from grazing and other removal processes is summarized by Kantrud et al. (1986a) and Kirby et al. (1992). Indicator species that suggest the occurrence of previous grazing, tillage, or mowing in prairie wetlands are shown in Appendix A, which was expanded from Stewart and Kantrud (1972), Millar (1973), and Kantrud et al. (1989). However, both Walker and Coupland (1970) and Stewart and Kantrud (1972) concluded that land use factors (e.g., intensity of grazing, mowing, tillage), as compared with hydrologic and water chemistry influences, are less important determinants of species composition, except during drought years.

Community Zone Locations

Cover removal differentially affects the outermost zones of prairie wetlands because these are the most accessible to equipment and grazing livestock. Presence of a deep emergent zone surrounded by open water can be a sign of intense grazing (Stewart and Kantrud 1972a).

Species Richness

Decreases in wetland plant richness are sometimes a reflection of reduced intensity of disturbance from grazing, fire, or other removal processes in prairie wetlands. Conversely, long-term increases in species richness can indicate periodic occurrence of these disturbances. However, species richness is a poor indicator of vegetation removal, because in some instances, the thinning of stands of native vegetation allows new plant species to obtain a competitive foothold. Although this can increase the species richness initially, species richness can decline over the long term if new species, as is often the case, are ones that are highly aggressive and tend to form homogeneous stands. Even atyears after being restored, several Iowa wetlands had significantly lower species richness within each of their vegetative zones (except the deepwater zone) than did natural wetlands (Galatowitsch 1993).

Biomass, Aerial Cover, and Cover Ratio

Decreases in aerial cover and biomass of wetland vegetation, and increase in the cover ratio, are a strong indicator of recent fire, haying, excessive grazing, or contamination with herbicides have occurred recently or for a prolonged period. Conversely, long-term increases in aerial cover can indicate relative absence of these disturbances. A survey of restored prairie wetlands in Iowa (Galatowitsch 1993) found that shoot densities of emergent vegetation were less in restored wetlands than nearby natural wetlands.

3.4.3 Vascular Plants as Indicators of Wetland Salinity

Species Composition

The species composition of vascular plants is an excellent indicator of wetland salinity in prairie wetlands (Stewart and Kantrud 1972b, Looman 1981). Most freshwater macrophytes cannot tolerate more than 10 ppt dissolved salts (Reimold and Queen 1974, Ungar 1974). Of the 195 major plant species in prairie wetlands, the general categories of salinity that describe the occurrence of 157 (79%) species are known (Appendix A), and quantified tolerance (or preference) ranges are available for 120 species based on presence or absence in wetlands spanning a salinity gradient (Appendix E). For many species, exact thresholds of salinity tolerance vary by the type of salt (Mg, Na, etc.), life stage, genetic population, duration of exposure, temperature, and other factors (Lieffers and Shay 1983).

Species Richness

Diminished species richness can be a sign of hypersaline conditions (Reynolds and Reynolds 1975).

Biomass and Cover Ratio

Compared to most prairie wetlands, hypersaline wetlands probably have lower aerial cover and biomass of emergent vegetation, but only a few data are available (e.g., Wali 1976). An analysis of experimental data from the Delta Marsh suggested that, at least for some species, the influence of salinity on vascular plant production can overshadow the effects of other factors associated with water level change (Neill 1993). In Australia, saline lakes sometimes support greater densities of submerged macrophytes (and consequently waterfowl) because salts enhance the flocculation of suspended clay particles, thus increasing water clarity and light penetration (Kingsford and Porter 1994). In contrast, in England increases in lake salinity may cause shifts in plant community dominance from submerged plants to phytoplankton, at least at intermediate levels of nutrient loading (Moss 1994). In Ohio, fertilization of salt-tolerant inland wetland plants with nitrogen spurred their production (Loveland and Ungar 1983). In a study of four prairie wetlands, Fulton et al. (1979) and Fulton and Barker (1981) found that "vegetation density was a better indicator of soil properties than was the type of vegetation."

Germination Rate

Seed germination rates are inhibited by high (> 2 mS/cm) salinities, as occurs most often during drawdown events (Smith 1972, Galinato and van der Valk 1986).

Biomarkers

Preliminary experiments by Mendelsson and McKee (1992) suggest that the concentration of proline, an amino acid, in plant tissue can indicate the occurrence of salt stress that occurred during the previous five days.

3.4.4 Vascular Plants as Indicators of Sedimentation and Turbidity

Species Composition and Community Zone Locations

A shift in community composition from submersed species to emergent and floating-leaved species can signal recent increases in turbidity, which could be attributed to suspended sediment and/or phytoplankton (Niemeier and Hubert 1986, Hough and Forwall 1988). Submersed plants in northern latitudes often have only a brief period during which they must grow sufficiently to reach the water surface. If turbidity or other factors inhibit their growth during this period, or if a late winter shortens the growth period, submersed plants may not reach the water surface before their growth is stunted by algal blooms that begin in early summer (Engel and Nichols 1994). Submersed plants are generally less tolerant of increased turbidity than are benthic algae and phytoplankton (Dennison et al. 1993).

Some of the submersed plant species of greatest value to waterfowl and invertebrates do not persist when the Secchi disk depth is less than about 0.3 m (Chambers and Kalff 1985, Kantrud 1990). Secchi transparencies of 0.2-0.4 m are common for brief periods during algal blooms in prairie wetlands (Barica 1975). Tolerances of submersed plants to turbidity are defined more accurately by the light compensation point of each species, the point where its photosynthesis equals its respiration (Dennison et al. 1993, Kahl 1993). In Wisconsin, submersed species that are most tolerant of turbidity are characterized by rapid growth during the early spring, summer leaf canopies, and winter tubers or rhizomes (Engel and Nichols 1994). Data on relative depth maxima and ranges for many submersed species are compiled in Davis and Brinson (1980), and a more refined data base is currently being developed by USEPA's Wetlands Research Program (N. Detenbeck, personal communication, USEPA Environmental Research Laboratory, Duluth, MN). The more shade-tolerant submersed species can perhaps be identified from information on turbidity in Appendix A.

The relative extent of submersed species also declines with increasing sediment input when the sediment so fills a depression that water depths become too shallow and standing water fails to persist through the growing season (Edwards 1969). Similarly, in very shallow, temporary wetlands, sedimentation probably increases the dominance of non-wetland species.

The species composition of emergent species might also indicate relative degree of sedimentation, because mature plants and perhaps their seeds differ with respect to tolerance to burial (van der Valk et al. 1981). Although stiff-stemmed emergents (e.g., cat-tail, common reed) seem least affected by sedimentation, comparative data for other prairie species are mostly lacking.

Species Richness

Turbidity and sedimentation can reduce the species richness of wetland plants (e.g., Engel and Nichols 1994), both directly and because of secondary hydrologic impacts. If the effects of sediment on germinating seeds can be assumed to be similar to the effects of accumulating plant litter, then the results of van der Valk's (1986) experiments in a prairie wetland suggest that the species richness of adult emergent wetland plants would be reduced by increased sediment. The viability of the seeds of most of these plants, however, is relatively unaffected when covered by up to 35 cm of sediment (van der Valk and Davis 1979); removal of overlying sediment should allow them to germinate.

Biomass and Cover Ratio

As described above, turbidity definitely decreases the aerial cover of submersed macrophytes. Decreases in aerial cover or biomass of emergent species are probably not as strong an indicator of sedimentation, but data are lacking. If the effects on germinating seeds of sediment can be assumed similar to the effects of accumulating plant litter, then the results of van der Valk's (1986) experiments in a prairie wetland suggest that the shoot density of emergent wetland plants would be reduced. In deep permanent basins, sedimentation can increase the area of substrate within the euphotic zone that can be colonized by wetland plants thereby supporting increased aerial cover of emergents.

Germination Rate

Repeated burial by as little as 5 cm of sediment per year can be detrimental to seedlings of some emergent species (van der Valk et al. 1981). In Delta Marsh, a sediment layer of at least 1 cm significantly reduced seed germination of many emergent species, and a layer of 4-5 cm prevented germination of most of the species tested (Galinato and van der Valk 1986).

3.4.5 Vascular Plants as Indicators of Excessive Nutrient Loads and Anoxia

Species Composition and Community Zone Locations

Declines in submersed plants, and increases in emergent and especially floating-leaved plants (e.g., Nuphar, Lemna, Wolffia), are one sign of increased inputs of nutrients to wetlands, especially in wetlands that initially had been nutrient-poor. Although a few submersed species (Ceratophyllum demersum, Utricularia vulgaris) appear to tolerate moderate nutrient additions, most submersed species decline. This is because algae respond more quickly than vascular plants to nutrients; consequently they proliferate in open water areas in the form of light-obscuring blooms that limit submersed macrophyte growth and reproduction (Mulligan et al. 1976, Phillips et al. 1978). Most emergent species (except those capable of forming floating mats) increase less rapidly in response to waterborne nutrient inputs than do floating-leaved species (Ozimek 1978, Shimoda 1984, Graneli and Solander 1989), because the emergent plants obtain nutrients mainly from the sediment, whereas most floating-leaved plants obtain nutrients directly from the water column. Enrichment can also stress individual emergent plants, if it causes algal blooms that, upon collapse, deprive sediments of oxygen for prolonged periods as they decay (e.g., Hartog et al. 1989, McDonald 1955, Barica 1984, Barica and Mathias 1979).

Of the emergent species, various annual species, as well as cat-tail and common reed, often dominate enriched wetlands (Kadlec 1979, Hartland-Rowe and Wright 1975, Finlayson et al. 1986, Kadlec 1990, Kadlec and Bevis 1990). Species that are most efficient in transferring oxygen to their roots might flourish the most under highly enriched conditions, when sediment oxygen levels decline (Barko and Smart 1983). Cat-tail, for example, effectively transfers oxygen and also requires only trace amounts of dissolved oxygen for germination (Leck and Graveline 1979). Chemical conditions appear to have a greater influence on the presence of rarer, perennial species than on the occurrence of aggressive, common species (Pip 1979). Many reports, especially in the European literature, have categorized individual wetland plant species according to their nutrient-level preferences, and thus as to their potential as indicators of eutrophication (Moyle 1945, Pip 1979, Stewart and Kantrud 1972b, Swindale and Curtis 1957, Wiegleb 1981, Zoltai and Johnson 1988, Husak et al. 1989). Plant species composition is less effective in reflecting moderate enrichment than severe enrichment. This is because algal and microbial communities are often initially more effective than vascular plants in assimilating inputs of nutrients (Richardson and Marshall 1986).

Species Richness

Increasing species richness of herbaceous plants, particularly of emergent species, can signal moderate increases in the fertility of a wetland (Pip 1987a,b, Graneli and Solander 1988). However, severe enrichment can decrease species richness for reasons given above (Lachavanne 1985, Lind and Cottam 1969, Tilman 1987, Hough et al. 1989, Toivonen and Back 1989). Although plant richness is correlated positively with nutrient levels (especially phosphorus) in prairie wetlands, water chemistry variables explain less than half the total variability in plant richness (Pip 1987a).

Biomass and Cover Ratio

Increases in aerial cover and biomass of non-submersed wetland plants are commonly one sign of enrichment. Variables that describe plant characteristics having short turnover times, such as aboveground biomass and leaf area of annual plants, can be relatively sensitive indicators of enrichment. Cat-tail biomass and production respond to annual fluctuations in nitrate, making cat-tail a possible indicator of erratic inputs of nutrients (Davis 1989). In one fertilization experiment in a prairie wetland, the aboveground production of cat-tail (Typha glauca) increased 19% and that of burreed (Sparganium eurycarpum) increased 57% (Neely and Davis 1985a). However, enrichment of prairie wetlands does not always increase plant biomass and production in the long run. This is because the shading and smothering effect of litter that remains following one year's excessive production can inhibit the germination and growth of individuals in successive years (Nelson and Anderson 1983, Neill 1990).

Biomarkers

Experiments by Mendelsson and McKee (1992) and others suggest that the concentration of alcohol dehydrogenase in plant roots can indicate the occurrence of oxygen stress, perhaps from overenrichment, within the previous five days.

3.4.6 Vascular Plants as Indicators of Pesticide and Heavy Metal Contamination

Limited data on the effects of pesticides and metals on prairie wetland plants are contained in tables published by Sheehan et al. (1987), and in USEPA's PhytoTox data base. A host of factors associated with actual applications influence contaminant toxicity and plant mortality (Doust et al. 1994), and can include:

Environmental Factors: water temperature, organic content, pH, alkalinity, suspended solids.

Dose Factors: concentration, the specific compound or formulation (inert ingredients), frequency of application or exposure, duration of exposure.

Biotic Factors: the plant species, its life stage (season), degree of simultaneous stress from other factors.

Species Composition

Wetland plant species differ in their tolerances of various heavy metals, selenium, and herbicides. Consequently, changes in species composition can indicate past and ongoing incidents of exposure to these contaminants.

Although few studies have compared relative sensitivities of various wetland plants to a particular contaminant, data compiled from many single-species experiments involving heavy metals suggest that emergent plants might be generally more tolerant of heavy metals than submersed plants, which in turn might be more tolerant than algae (Outridge and Noller 1991). Duckweed (Lemna) appears to be particularly sensitive to the heavy metals cadmium and nickel, and chromium concentrations of 10 mg/L are inhibitory (Huffman and Allaway 1973). Cat-tail can tolerate lead, copper, and chromium accumulations of at least 10 m g/g (dry weight) of aboveground biomass; zinc accumulations in cat-tail can reach 25 m g/g (dry weight) without apparent ill effects (Mudroch and Capobianco 1979). Common reed can tolerate industrial wastewater with high levels of heavy metals (e.g., up to 250 micrograms/g sediment copper concentrations), as do bulrushes (Seidel 1966). In an Ontario river, submersed species (Elodea, Ceratophyllum, and Myriophyllum) appeared to be less tolerant of industrial wastes than floating-leaved and short, rooted aquatic plants (Potamogeton, Nuphar, and Nymphaea), which were in turn less tolerant than cat-tail and common reed (Dickman et al. 1980, 1983, Dickman 1988).

By knowing the characteristic sensitivities of various plant species, plant community composition can be used cautiously to infer past exposure to particular contaminants. For example, application of one herbicide resulted in reduced dominance of Phragmites australis but increased dominance of particular Lemna, Callitriche, and Potamogeton species.

Floating-leaved herbaceous plants are sensitive to the physical effects of oil compounds associated with herbicide applications, and growth of the duckweed Spirodela oligorhiza is affected by PCB concentrations of 5 mg/L (Mahanty 1975). Cat-tail can tolerate petroleum oil concentrations of 1 g/L (Merezhko 1973) and, along with common reed (Phragmites), appeared to be the most tolerant macrophyte downstream from an industrial effluent source in Ontario (Dickman 1988). Bulrushes are killed by phenol concentrations of 100 mg/L and abnormalities occur at large phenol concentrations, but new shoots form quickly (Seidel 1966).

Species Richness

Species richness of vascular plants in prairie wetlands would be expected to decline in response to atypically high loadings of certain heavy metals and especially to chronic exposure to herbicides. However, documenting data are lacking.

Biomass and Cover Ratio

Herbicides have an obviously direct effect in reducing the aerial coverage of selected vascular plants to which they are applied directly, especially early in the growing season. Indirect effects also can occur. Emergent cat-tails can be killed by herbicides applied for control of submersed plants (Newbold et al. 1974). Low concentrations of the popular herbicide, glyphosate sometimes stimulated growth of sago pondweed, Potamogeton pectinatus, a major submersed plant in prairie wetlands (Hartman and Martin 1985), but glyphosate is lethal to most emergent wetland plants at typical application levels (Sheehan et al. 1987).

Relatively few chemical assays have been conducted in field mesocosms in the prairie region. The commonly used herbicide, atrazine, has been examined the most. A review of published literature on atrazine effects found a wide range of concentrations (0.050-1.310 mg/L) associated with adverse effects on one prairie wetland species, Potamogeton perfoliatus (Hofmann and Winkler 1990, Swanson et al. 1991). The concentration that would be expected to occur immediately after a 0.5-ha prairie wetland is directly sprayed with atrazine would be about 0.413 mg/L (Sheehan et al. 1987), and chronic levels in prairie wetlands appear to be much lower than this (Ruelle and Henry 1993). A concentration of 1 mg/L atrazine caused a 50% decline in biomass of three macrophytes (Lemna, Ceratophyllum, and Elodea) over a 30 day period in a prairie wetland mesocosm (Johnson 1986). The pesticides carbofuran, fonofos, phorate, treflan, and triallate had no statistically significant effect on biomass of the plants mentioned above (Johnson 1986). Effects of herbicides on flowering of plants and storage of energy reserves (as seeds or tubers) has received little study in prairie wetlands, and could have important implications for waterfowl that feed extensively on these items and their associated invertebrates just prior to migration in late summer and early fall. Effects of heavy metals and selenium on prairie wetland plants are mostly unstudied. Selenium is toxic to some wetland plants at concentrations greater than 1.25 mg/L (Ornes et al. 1991).

Bioaccumulation

Wetland plants rapidly take up selenium (Ornes et al. 1991) and bioaccumulate many other heavy metals (Freemark et al. 1990). Cat-tails are among the few plant species that have been analyzed for bioaccumulation of contaminants in prairie wetlands (True and Dornbush 1984).

Biomarkers

Activity of the enzyme, peroxidase, has demonstrated usefulness in aquatic plants as a marker of previous exposure to several metals and organic contaminants (Byl 1994).

3.5 Monitoring Techniques

3.5.1 Ground-based Sampling

Many documents provide detailed guidance on sampling herbaceous vegetation (e.g., Higgins et al. 1994). Some address wetlands specifically. These include Dennis and Isom 1983, Downing and Anderson 1985, Fredrickson and Reid 1988a, Moore and Chapman 1986, Mueller-Dombois and Ellenberg 1974, van der Valk 1989, Phillips 1959, Schwoerbel 1970, and Woods 1975. Some methods for measuring production of prairie species are described by Neill (1992, 1993).

If wetlands can be sampled only once, mid-growing season is usually the recommended time. However, many plants are apparent and/or identifiable only for a few weeks of the growing season. Thus, if the aim is to estimate annual production or quantify community composition accurately, repetitive visits that account for the diverse phenologies of wetland species should be implemented (Dickerman et al. 1986, Smith and Kadlec 1985). Ideally, annual visits should be timed to coincide with year-specific weather conditions, rather than calendar date. Trampling of herbaceous vegetation and compaction of saturated soils during even a single site visit can induce community changes detectable in subsequent visits. Thus, field crews should be as small as possible and follow the same path in and out of a wetland. In deeper wetlands, use of underwater SCUBA transects is sometimes appropriate (Schmid 1965). If herbaceous wetland vegetation must be sampled destructively in order to obtain specimens for identification (e.g., in very turbid or deep waters), then equipment such as dredges, oyster tongs, plant grappling hooks, and steel garden rakes can be used (Britton and Greeson 1988). Equipment designed specifically for sampling submersed macrophytes is described by Dromgoole and Brown (1976), Macan (1949), Satake (1987), Woods (1975), and others. However, whenever possible, plants should be identified in the field rather than collected.

3.5.2 Aerial Methods

The EMAP monitoring effort is using low-altitude aerial metric photographs (scale about 1:2400) to measure vegetation cover annually on each of 48 wetlands. Low-altitude imagery has been used successfully in several previous studies in the prairies (Kreil and Crawford 1986, Welling et al. 1988b, van der Valk and Squires 1992, van der Valk et al. 1994) to differentiate emergent plant communities, stem density classes, and in some cases, species. Filters and image processing techniques can also be used to highlight various spectra, such as those sensitive to chlorophyll-a (Patience and Klemas 1993). Under ideal circumstances, such an approach might be used to indicate the presence, relative biomass, and condition of particular wetland species.

When communities or species can be distinguished reliably from aerial images, the relative spatial extent (percent cover) of the communities or species can be measured more accurately and cost-effectively than from ground transects. In the van der Valk and Squires (1992) study, false infrared imagery from an altitude of 610 m was used. Aerial videography is also being used more frequently for measuring wetland structure and determining wetland integrity (Cowardin and Sklebar 1993), and can delineate cover types covering as little as 1 m² in some instances (Olson 1992). As demonstrated by Welling et al. (1988b) and described by Caldwell and van der Valk (1989), aerial imagery can be used to make maps which, when overlaid with bathymetric maps derived from field transects, are useful for measuring the depth ranges of various plant species and provide essential information that cannot be derived from quadrat data. Although aircraft are commonly used for obtaining the photographs, tethered balloons with remotely triggered cameras also show some potential for use in monitoring prairie wetlands (Edwards and Brown 1960), and cost about $20 each, excluding the camera (Davis and Johnson 1991).

3.5.3 Potential or Historical Vegetation

Because the vegetation in prairie wetlands shows such tremendous interannual variability, seeds lying in wetland soils or sediments (the "seed bank") are often sampled in lieu of or in addition to mature plants. The assumption is that, because seeds decompose much more slowly than foliage (over decades rather than months), they indicate conditions not necessarily shown by live vegetation. Similarly, pollen from prairie wetlands remains intact in anaerobic sediments for long periods, and under some circumstances can be identified to species under a microscope, potentially yielding information on the nutrient status and water levels present historically in the wetland (Watts and Bright 1968, Vance and Matthewes 1994).

The simplest and most economical method of analyzing the seed bank is known as the seedling emergence (or seedling assay) method. Procedures are described by van der Valk and Davis (1978) and Galatowitsch and van der Valk (1994). Each sample consists of about 1000 cm² of substrate removed to a depth of 5 cm. Large organic matter (leaves, etc.) is removed, sometimes using a coarse sieve, and the sample is distributed in a shallow (<cm) layer in two pans. In one pan the sediment is kept covered with a centimeter of water, and in the other it is merely kept moist. The pans are kept in a greenhouse or sunlit shelter for the duration of the growing season, or longer in a heated facility if necessary. As seeds sprout, they are removed, enumerated, and identified by comparison with a reference collection of seedlings. An adjacent pan containing sterile soil is also tracked to ensure that seedlings in the other two pans are the result of seeds contained in the sediment, rather than blown in. Results of using this method with samples from prairie wetlands are reported by Welling et al. 1988a,b, Pederson 1981, Merendino and Smith 1991, Galatowitsch 1993, van der Valk and Davis 1979, and Poiani and Johnson 1988, 1989.

The seedling emergence method requires time to establish a reference collection by taking seeds from known (fully identified) mature plants and growing the seeds to the seedling stage. Because seedlings of many species (e.g., Carex, Cyperus, Bidens) can seldom be identified until they are fully mature, they must at first be tentatively ed by appearance and counted with representatives of each taxa grown to maturity for identification. The success of the seedling emergence method hinges on the veracity of the assumption that seedling density is proportional to seed density when in fact, there are frequently situations where mature plants are obvious in a wetland but use of the seedling emergence method produces none of their seedlings (van der Valk and Davis 1978, Welling et al. 1988a,b, Wienhold and van der Valk 1989). This situation arises because (a) appropriate incubation conditions are unknown for some species; (b) some wetland plants produce seeds only rarely (they normally reproduce vegetatively); and (c) there is often considerable inter- and intraspecific variation in seed set, seed viability, and seedling competition. For these reasons, seed bank analysis techniques are inadvisable where the objective is to inventory rare species. In particular, some seedbank assays tend to overestimate annual "weedy" species (those with readily geminable seeds) while underestimating certain species that are especially sensitive to competition and/or which inhabit the drier zones of wetlands (S. Galatowitsch, personal communication, Univ. Minnesota, St. Paul). Nonetheless, other evidence (Poiani and Johnson 1988) suggests that the influence of such variation on the overall ability to characterize a wetland's past or potential vegetation is, at least some of the time, probably minor.

3.5.4 Bioassay Methods

A review of laboratory, outdoor mesocosm, or in situ bioassay methods involving vascular plants is beyond the scope of this document. Use of bioassays to explore contaminant toxicity to plants in prairie wetlands has been relatively limited. Examples include studies by Wayland and Boag (1990), Johnson (1986), and Ruelle and Henry (1993). Both mature plants and seeds have been assayed. Test species and protocols for assaying the effects of pesticides or other contaminants on wetland plants are proposed by Freemark et al. 1990, Swanson et al. 1991, Smith 1991, and Doust 1994.

3.5.5 Bioaccumulation

Methods for collecting wetland plants and assessing bioaccumulation of contaminants in plant tissues are described in Moser and Rope (1993b).

3.6 Variability and Reference Points

3.6.1 Spatial Variability

Species Richness

As a point of reference, approximately 922 herbaceous vascular plant species characteristic of pothole or riverine wetlands have been recorded in the North Plains region, which includes the Dakotas and eastern parts of Montana, Wyoming, and Colorado, but not parts of the prairie region in Minnesota and Iowa (Reed 1988). In prairie counties of North Dakota, about 135 (15%) of the North Plains region's wetland species are considered "rare" by the state's Natural Heritage Inventory (Appendix I); this represents 55% of the 245 rare plant species occurring in any habitat in North Dakota. Most of the rare wetland plants are associated with riverine, bog, or forested wetlands rather than with typical prairie pothole basins.

About 191 (21%) of all the North Plains wetland species occur frequently as dominants in prairie pothole wetlands (Appendix A). Among the dominant species are 17 species (9%) officially , of which five are annuals. The total number of annuals that sometimes dominate prairie wetlands is 33 (17% of all dominant species in prairie wetlands) (Appendix A).

In a survey of 112 prairie potholes in southern Manitoba, Pip (1979) found more than 47 vascular plant species. In a subset of 39 of the sites, she found 32 species with a mean of 5.3 species per wetland, and noted that even wetlands that were adjacent seldom had similar floras. In a larger (n = 177) set of Manitoba wetlands, she found a mean of 4.9 species per wetland. In a survey of 261 vegetation stands in 82 prairie potholes in northeastern Montana, Lesica (1993) found a total of 173 species, of which 12 were exotics. Among seven wetland complexes ranging in size from 7 to 15 ha, he found between 48 and 74 species per complex, and between six and 11 plant communities. Species richness was not strongly correlated with community richness. From five natural wetlands in eastern North Dakota, Kreil and Crawford (1986) reported a total of 64 vascular plant species. From 140 quadrats (each 1.0 x 0.5 m) in temporary and seasonal wetlands of eastern North Dakota, Hubbard et al. (1988) collectively found 41, 38, 20, and 16 species (on Tetonka, Parnell, Worthing, and Southam soil types, respectively). In part of the Delta Marsh, 65 plant species were found during four growing seasons, although no more than 48 of these were present in any single year (Squires and van der Valk 1992).

In a survey of 20 semipermanent wetlands in prairie regions of Iowa, Galatowitsch (1993) found a total of 158 species, of which 106 were species that typify wetlands. Collectively, there were 133 species in ten natural wetlands (range 41-63 per wetland, mean = 45.8) and 83 species in ten restored wetlands (range 24-52 per wetland, mean = 26.9). There were 75 species (48 of them "wetland" species) that occurred only in the natural wetlands, and 25 species (10 of them "wetland" species) that occurred only in the restored wetlands; most of the latter species were submersed plants or species planted for erosion control. Between 1 and 22 species were in the driest parts of each wetland, 7 and 49 species in the sedge meadow zone, and 7 and 19 species in the shallow emergent zone. Seedbanks of the natural wetlands contained 15 species, as compared with eight species in the restored wetlands.

Number of species in the Delta Marsh seed bank varied from 8 to 20/m², depending on the zone and depth from which the samples were collected (van der Valk and Davis 1979, van der Valk 1986). Over 90% of the seeds in most seed banks consists of fewer species than this (Wienhold and van der Valk 1989). However, germination of all seed bank samples from the Delta Marsh resulted in a cumulative total of over 40 species. Up to 14 species were found in seed banks in 11 wetlands in Iowa (Wienhold and van der Valk 1989) but in Iowa's Eagle Lake wetlands, van der Valk and Davis (1978) found 45 species. Cumulatively, from a of eight smaller Iowa wetlands, van der Valk and Davis (1976) found 29 species. One seedbank survey of 35 Iowa wetlands found an average of 16 species per wetland (Clambey 1975) and another survey of four Iowa wetlands found 24 species per wetland (LaGrange and Dinsmore 1989b).

Within individual semipermanent wetlands, richness is generally less in the wetter marsh zones than in drier meadow zones upslope (Nelson and Anderson 1983). From a survey of 246 stands of wetland vegetation in southern Saskatchewan, Walker and Coupland (1970) found the greatest richness in the wet meadow and marsh meadow portions of wetlands that were slightly saline and lightly grazed and mowed (43 species in nine wet meadow stands, 37 species in seven marsh meadow stands). Across a gradient of moisture, salinity, and disturbance from grazing and mowing, the stands that had the most unique floras were those that were the wettest, the most saline, and the most disturbed.

Biomass, Density, Cover Ratio

Biomass also varies greatly by species and spatially. In the Canadian prairies, aboveground biomass varied from 425 g/m² for one emergent species (Scirpus lacustris) to 1750 g/m² for another (Typha latifolia) (Shay and Shay 1986). For just a single species (cat-tail), aboveground biomass among several prairie wetlands can range 0 to 2106 g/m² (van der Valk and Davis 1978, Shay and Shay 1986), and up to 2400 g/m² under conditions of artificial enrichment (Neill 1990). Among seven communities of emergent wetland vegetation in eastern North Dakota, the net production was reported to range from 0.30 to 0.97 g/m²/day (Hadley and Buccos 1967). Submersed and floating-leaved species tend to be less productive than emergent species.

In describing their seed bank analyses, van der Valk and Davis (1978) noted, "Among replicate samples within a vegetation type, there [is] a great deal of variation in the number of individuals of a given species ... standard deviations are larger than the means in many cases." Moreover, some of the species currently present in a wetland will be absent from seed banks, whereas others may have seed densities in the sediment of several thousand per square meter. In the Delta Marsh, total seed densities in 250 sediment samples averaged 4582/m², and ranged from 140/m² in open water areas, to 2230/m² in common reed stands and 5810/m² in cat-tail stands (Pederson 1981). However, van der Valk and Davis (1979) found the number of viable seeds in the upper 5 cm of the Delta Marsh seed bank to vary from 7363 to 56,289/m², depending on the zone and sediment depth from which the samples were collected (analyses of cores extending from the surface down to 35 cm revealed a maximum of 255,000 seeds/m²). In a North Dakota wetland, seed densities across wetland zones varied by a factor of about 8; a maximum of 9370/m² was found (Poiani and Johnson 1989). In Iowa a of 10 natural (undrained, seasonally or semipermanently flooded) wetlands had an average seed density of 7369/m² whereas 10 nearby restored wetlands (previously drained, now permanently flooded) wetlands had an average of 3019/m² (Galatowitsch 1993). In contrast, in Minnesota a series of 30 undrained wetlands had an average seed density of 3600/m², whereas five recently drained wetlands had an average as high as 8000/m² (Wienhold and van der Valk 1989).

Cover ratio varies tremendously among prairie wetlands, but few estimates of spatial variability are available.

Germination Rate

Even within a single species, germination rates can vary considerably, e.g., 14-51% for cat-tail (Weller 1975).

3.6.2 Temporal Variability

Species Richness

Over a 7-year period, plant species richness in a single, 1.7-ha, semipermanent wetland in Iowa varied from 7 to 19 species per year (Weller and Voigts 1983). Only one species was present all seven years of the wet-dry cycle; two of 23 common species occurred during only 1 or two years. Over an 85-year period, vascular plant richness in an Iowa prairie lake varied from about 11 to 29 taxa (Niemeier and Hubert 1986). Apparently only one of 52 taxa was found throughout the period.

Biomass and Cover Ratio

Over a five-year period, the biomass of one plant species (Scirpus validus) in an Iowa wetland varied from 0 to 486 g/m²; that of another (Sparganium eurycarpum) varied from 271 to 543 g/m²; and that of a third (Typha glauca) varied from 772 to 1075 g/m² (van der Valk and Davis 1980). Within a single year, the aboveground biomass of Carex rostrata in a single Minnesota wetland varied from 114 to 852 g/m², whereas the belowground biomass varied from 150 g/m² to 328 g/m². Production peaked at 11 g/m²/day (Bernard 1974).

High interannual variation in the cover ratio typifies prairie wetlands. In a 10-year study of 71 Manitoba wetlands, Millar (1973) found that aerial cover changed in 32 (46%) of the wetlands, and of these, a complete conversion to open water occurred in 17 wetlands. In a North Dakota prairie wetland, seed densities in sediments varied from 2840/m² one year to 9370/m² the next (Poiani and Johnson 1989).

3.7 Collection of Ancillary Data

It is easier to separate the anthropogenic from the natural causes of impairment of community structure if data are collected or inferred simultaneously on the following variables of particular importance to vascular plants:

age of wetland and its successional status, light penetration (particularly for submersed species), water or saturation depth, conductivity and general chemistry of waters and sediments, abundance of herbivores (particularly muskrat, geese, grazing cattle, crayfish), sediment type, and the duration, frequency, and seasonal timing of regular inundation, as well as time elapsed since the last severe inundation, drought, or fire.

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; introduction of exotic species; and water regime modification.

3.8 Sampling Design and Required Level of Sampling Effort

3.8.1 General Considerations

If the sole objective is to inventory the presence or absence of plant species within wetlands, then a timed search method that covers all the obviously recognizable zones may be appropriate (e.g., Pip 1979, 1987a,b,c), whereas if vegetation cover and dominance are to be quantified, decisions must be made concerning the physical layout of transects and/or quadrats within a wetland. This task can pose a particular challenge if the objective is to characterize a wetland as a whole unit, rather than describing ecological relationships within just one zone or biological community. Choices involve deciding whether to place quadrats in a purely random manner, or randomly along transects, or at regular intervals along transects, or randomly/regularly within delimited vegetation zones. If transects are used, there are further choices regarding whether to locate the transects randomly (Gurney and Robinson 1988, LaGrange and Dinsmore 1989b, van der Valk and Davis 1978), evenly (gridded), or in relation to vegetation zones ⁽³⁾. The alignment of transects is usually perpendicular to shore (Wienhold and van der Valk 1989), but can be parallel to the long axis of the wetland (Niemeier and Hubert 1986), or follow the four compass axes (e.g., Poiani and Johnson 1988). When transects are used, vegetation can be enumerated using point counts (plants intercepted by the transect at specified intervals are counted), intervals (simple presence/absence of plants intercepted by the transect at specified intervals is noted), or intercepts (all plants intercepted are counted). Based on data from one prairie wetland, Weller and Voigts (1983) concluded that the intercepts method was least cost-effective, and the other two methods gave similar results. The EMAP effort has characterized species composition, relative dominance, and richness of prairie wetlands by randomly locating five 0.25-m² quadrats within each vegetation zone of each wetland while walking through the center of the zone parallel to its long axis. The rationale given for choice of this method is the fact that vegetation in prairie wetlands tends to occur as concentric zones, which would not be well-sampled by transects perpendicular to the shore or center of the wetland. Species not found by this method, but observed while walking between quadrats, are also recorded.

To characterize vegetation at sampling points, investigators (e.g., Wienhold and van der Valk 1989) have used the Releve method (Mueller-Dombois and Ellenberg 1974), Daubenmire approach (Daubenmire 1959), or the Braun-Blanquet (1932) approach (e.g., LaGrange and Dinsmore 1989a). In some situations the midpoint values of cover classes can be averaged to obtain mean cover values for each species within a zone (vegetation community), and these values can be weighted by area to obtain total cover values for species within zones.

If the sole purpose is to assess the habitat value of vegetation as shelter for wildlife (i.e., cover of individual species does not need to be determined), then visual obstruction readings (Robel et al. 1970) and profile board methods (Jones 1968) can be used (Nudds 1982, Duebbert and Lokemoen 1976, Higgins 1986, Barker et al. 1990). Cover estimates can be made once at the peak of vegetation development and/or during bird nesting, but additional estimates may be desirable earlier in springtime (to estimate residual matter) and at other dates if livestock are grazing the wetland. Cover estimations made by the EMAP effort featured the Daubenmire approach; cover was estimated as the portion of water surface (for emergents) or bottom (for submersed plants) shaded by each species.

One factor that affects sample costs is the desired level of taxonomic identification. Identifying plants to the species level usually allows the investigator to make more refined statements about the condition of a wetland, but can increase the required time and requires advanced training and experience. There are no data to indicate whether, and under what conditions, identification of plants only to the genus or family level would be sufficient to define the ecological integrity of a prairie wetland.

Sampling costs are determined not only by the time required to identify plants, but also by the number of quadrats examined. This number should depend on expected variability (coefficient of variation), wetland size, and the desired precision. Larger wetlands require more transects or quadrats, usually spaced farther apart, to accurately characterize overall community composition. More linear wetlands (e.g., narrow fringe marshes along lakes) usually require more tightly spaced sampling points, as do ecotone areas along transects. In a Wisconsin lake, the number of samples needed to adequately quantify the biomass of the entire submersed plant community within various plant community zones, given a goal of maintaining a probability of 95% of being within 15% of the mean, ranged from seven to 200, depending on the zone (Nichols 1984a).

3.8.2 Asymptotic Richness: Results of Analyses

If the goal is not just to quantify species richness within samples, but for the whole wetland (or complex), then considerably more samples are required. The number will be determined not through examination of coefficients of variation, but rather, by plotting species accumulation curves (the cumulative number of species vs. number of samples, or vs. wetland area). Based on their Saskatchewan wetland data, Walker and Coupland (1968) reported that up to 30 quadrats (0.5 m x 0.5 m) per vegetation zone were necessary to characterize the species composition of the zone, i.e., to level off the species accumulation curve. Apparently no similar analyses have been prepared for other prairie wetlands.

For this document, we analyzed plant taxonomic richness from two data sets from prairie wetlands⁽⁴⁾ . One was from quadrats along transects in a series of 20 semipermanent wetlands in Iowa (Galatowitsch 1993). Data from the quadrats and transects were pooled into a single list for each wetland. Our calculations of asymptotic richness indicated that, for example, only 11 wetlands would need to be sampled to capture 99% of the taxa present in all 20 wetlands (Appendix O). The statistical approach used to determine this was described in Section 1.5.

The 20 Iowan wetlands consisted of 10 restored and 10 natural wetlands. When the restored wetlands were compared with the natural wetlands, restored wetlands were found to accumulate taxa at a slightly more rapid rate. This was likely due to the greater homogeneity of species composition among the restored wetlands. Thus, for monitoring designs and conditions similar to those of Galatowitsch (1993), restored wetlands would not have to sampled quite as extensively as natural wetlands. About eight wetlands would be a sufficient number to detect 90% of the plants in the combined population of 20 wetlands.

The other data set consisted of vegetation quadrats from the Marsh Ecology Research Program (MERP) experimental units of Delta Marsh, Manitoba (Squires 1991). Data involving multiple quadrat collections per year were combined from two wetlands to create a single taxa list for each of four treatments:

DD1LOW: One year of drawdown, followed by reflooding to the original depth (n = 385 quadrats covering four post-treatment years)

DD2LOW: Two years of drawdown, followed by reflooding to the original depth (n = 362 quadrats covering four post-treatment years)

DD2MED: One year of drawdown, followed by reflooding to an intermediate depth (n = 387 quadrats covering four post-treatment years)

DD2HIGH: Two years of drawdown, followed by reflooding to a high (deep) depth (n = 351 quadrats covering four post-treatment years)

Our calculations of asymptotic richness revealed the following ordering of species accumulation rates by treatment (Appendix O):

DD2MED (fastest) > DD1LOW > DD2HIGH > DD2 LOW

The differences in accumulation rate among treatments were not great, and the results suggest species composition was most homogeneous among the quadrats in the DD2MED treatment. In most cases an average of about 37 quadrats was adequate to detect 99% of the taxa that were cumulatively present in the full 351-387 quadrats, although for any particular treatment, as few as 20 and as many as 40 quadrats could be needed, depending on the sequence in which the quadrats are collected. Had all 3281 quadrats been considered regardless of treatment (a situation that more closely approximates sampling of natural wetlands with varied water regimes), about 448 quadrats would be needed to detect 99% of the taxa found in all 3281 quadrats.

3.8.3 Power of Detection: Results of Analyses

The Components of Variance approach, as described in Section 1.5, was applied to just the Squires data set. As tabulated in Appendix M, the mean species richness per quadrat levels off at a sample size of about 20 quadrats, in wetland situations and monitoring designs similar to those of Squires. To distinguish interannual (or between-wetland) changes of (for example) two plant species, a total of between 9 (optimistically) and 12 (pessimistically) quadrats would need to be collected. Conversely, if budget or other considerations limited sample size to five quadrats per wetland, then one would be able to detect a mean difference of 2.9 to 3.6 species between wetlands, years, or whatever. In both instances, we are assuming there is an 80% certainty of being correct at the 5% level.

From the same data set, if the objective is to estimate mean seedling density per quadrat, then more than about 12 quadrats are needed in wetland situations and monitoring designs similar to those of Squires. To distinguish interannual (or between-wetland) changes in seedling density of (say) 60 seedlings per quadrat, one would need to sample using only about four quadrats, but to detect a change of only 20 seedlings one would need to use 25-30 quadrats.

3.9 Summary

The species composition of vascular plant communities, and to a lesser degree their species richness, can indicate changes in prairie wetland salinity, water regime, and (among submersed species) sedimentation/turbidity (Table 2). Thresholds for responses are well-documented with extensive published field data from the region. Vascular plants also respond sensitively to changing nutrient levels, grazing, and presence of some contaminants, but existing information is too limited and confounding effects are too prevalent to currently allow widespread use of vascular plants to diagnose impairment of prairie wetlands from these stressors.

Vascular plant communities are exceptionally valuable indicators of conditions in individual prairie wetlands because they are immobile and because they integrate stresses that have occurred intermittently or chronically over months and years. This is especially true when wetland seed banks can be analyzed. Analysis of seed banks, although time-consuming and containing some biases, also can provide one piece of data useful towards defining appropriate "reference" conditions for a wetland resource. Compared with use of other indicators, monitoring of plant species composition causes minimal disturbance, involves (for most taxa) relatively simple identification procedures, and does not require frequent sampling or tight scheduling within a season. Where access to private property is restricted, the gross spatial patterns of vascular plants within a wetland can be characterized using widely available aerial photography, and then used cautiously to infer wetland ecological condition.

Individual prairie wetlands generally contain about 40-60 species of vascular plants, whereas the cumulative total of wetland plant species in the prairie region exceeds 900. There are relatively few quantified, published estimates of interwetland and interannual variability of vascular plant richness and biomass in prairie wetlands. Plant species composition varies considerably among otherwise apparently similar wetlands, and varies among years as well according to a distinctive wet-dry cycle.

Some additional research is needed to improve the potential for using vascular plants as indicators of prairie wetland condition. Descriptive investigations of life histories of many species are needed so that the daunting number of species can be ed into fewer functional s (such as defined by van der Valk 1981 or Boutin and Keddy 1993). Responses to various stressors of such streamlined ings can then be investigated more efficiently (and the results can be generalized more accurately) than if responses of purported indicator species were the sole focus of research and extrapolation. Research is also particularly needed to document the threshold responses of different vascular plant seeds, seedlings, and mature plants to sedimentation.

Table 2. Summary Evaluations of Vascular Plant 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 stressors	Species Composition Community Zone Locations Richness (mature plants) Richness (seed banks) Biomass, Cover Ratio Seed Density Germination Rate	GOOD GOOD FAIR (CE) FAIR (CE) GOOD FAIR (CE) POOR (CE)
Changes in vegetative cover conditions	Species Composition Community Zone Locations Richness Biomass, Cover Ratio	GOOD GOOD POOR (CE) GOOD
Salinity	Species Composition Richness Biomass, Cover Ratio Germination Rate Biomarkers	GOOD FAIR (CE) FAIR (FD) FAIR (CE) FAIR (FD)
Sedimentation & turbidity	Species Composition Richness Biomass, Cover Ratio Germination Rate	GOOD POOR (FD) FAIR (CE) POOR (CE)
Excessive nutrients & anoxia	Species Composition Richness Biomass, Cover Ratio Biomarkers	GOOD POOR (CE) FAIR (CE) POOR (CE, FD)
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	FAIR (CE) POOR (FD) POOR (FD) POOR (FD)

1. ¹     Prairie wetland species with the most information on hydrologic tolerences include Typha spp. (>30 references), Scolochloa festucacea (21 references), Scirpus validus (19 references), and Phragmites australis (18 references). See Appendix 5 for lists of references, by species, on water depth tolerences in prairie wetlands.

2. ²     These are: Acorus calamus, Agropyron repens, Artemisia biennis, Bidens cernua, Cirsium arvense, Cirsium floodmannii, Echinochloa crusgalli, Echinochloa muricata, Glaux maritima, Glyceria maxima, Kochia scoparia, Lysimachia thyrsiflora, Lythrum salicaria, Plantago major, Sonchus arvensis, Spirodela polyrhiza, and Stachys palustris.

3. ³     Statistical methods are available for defining zones somewhat objectively, as demonstrated in prairie wetlands by Johnson et al. (1987).

4. ⁴     The database descriptions that follow are generalized. For a detailed description of monitoring design and data structure of each data set, see Appendix 12.