Water: Monitoring & Assessment
5.0 Wetland Algae
Impacts on Quality of Inland Wetlands of the United States:
A Survey of Indicators, Techniques, and Applications of Community Level Biomonitoring Data
Excerpts from Report #EPA/600/3-90/073
(now out of print)
This discussion concerns wetland communities containing phytoplankton, metaphyton, benthic algae, periphyton, and epiphytic algae. Wetlands may contain algal communities that differ from other surface waters, or that indirectly influence community composition of algae in receiving waters. For example, acidic wetland waters commonly are rich in desmid species and acid-tolerant diatoms, such as Eunotia, Frustulia, and Pinnularia (Flensburg and Spalding 1973, Graffius 1958, Patrick 1977). Marshes may become dominated by Nostoc pruniforme, Microcoleus paludosus, Vaucheria sessilis , and sometimes Aphanothece stagnina (Prescott 1968). In a study of the effect on periphyton in a river above and below a marsh, Perdue et al. (1981) found some species of Navicula were common upriver of a marsh but almost non-existent below the marsh; several Nitzschia spp. and Fragilaria spp. were common below but rare above the marsh. Fragilaria construens was abundant in both areas.
5.1 Use as Indicators
As with microbial communities, algal communities in wetlands have most often been measured indirectly, in the pursuit of estimates of photosynthesis, respiration, and productivity. Few studies have quantified algal community structure in wetlands, or identified particular wetland algal species as indicators of wetland ecological condition. However, paleoecological studies of several peatlands have been undertaken. These use diatoms and pollen from peat cores as indicators of ancient environmental conditions (e.g., Agbeti and Dickman 1989, Battarbee and Charles 1987).
Following are discussions of algal community responses to various stressors.
Enrichment/eutrophication and Organic loading . Algal blooms are synonymous with eutrophication, so algae (particularly blue-green forms) are obvious indicators of trophic state, at least in lakes (Hecky and Kilham 1988). As concentrations of phosphorus in flowing water begin to exceed 0.020 mg/L, or 0.015 mg/L (and frequently less) in standing water, significant changes in algal communities can begin to occur (e.g., Traaen 1978), particularly if flow-adjusted loads are greater than 0.22 g/m3 (Craig and Day 1977). Florida regulations for discharge of treated wastewater to forested wetlands specify that, on an annual average basis, waters entering the wetland contain less than 3 mg/L nitrogen and less than 1 mg/L phosphorus; the monthly average for total ammonia must be less than 2.0 mg/L. Enriched conditions can be associated with either increased (e.g., Morgan 1987) or decreased (e.g., Hooper 1982, Schindler and Turner 1982) species richness of algal communities, depending on whether algae are mostly epiphytic or benthic, the pH, water regime, original state of the system, and other factors. Few studies have used algal community composition to classify the trophic state of wetlands. In other shallow surface waters, taxa such as the following (for example) have become dominant in response to fertilization (Mulligan et al. 1976, Patrick 1977, Prescott 1968):
In New Jersey streams exposed to residential and agricultural runoff, Morgan (1987) reported a shift from species characteristic of the region to species that had been geographically peripheral to the region. Algal community structure in some cases might be capable of reflecting the form of enrichment; based on experiments in a Michigan bog, chlorophytean species responded particularly to ammonium, whereas blue-green (cyanobacteria) species dominated when phosphate was added (Hooper 1982). Euglenophytes (one-celled, mobile 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 (Hutchinson 1975). Near a wastewater-disposal pipeline in a Michigan bog, several algal species bloomed-- Cladophera glomerata, Microspora, Euglena, and Spirogyra (Richardson and Schwegler 1986); algal growth rates were faster at the outfall site than at the control and at various distances away from the outfall.
Contaminant Toxicity. Numerous studies have demonstrated adverse effects of heavy metals (Whitton 1971), herbicides, synthetic organics, oil, and/or heavy metals on freshwater algae. Most such studies have been conducted in laboratories or non-wetland mesocosms, and/or have generally not examined community structure. Several (e.g., Hurlbert et al. 1972) report major algal blooms occurring after insecticide application due to temporary suppression of grazing by aquatic invertebrates. Herbicides have been shown to cause a shift in community composition from large filamentous chlorophyes (green algae) to smaller diatom species and blue-green algal species, particularly those of the order Chaemaesiphonales (Goldsborough and Robinson 1986, Gurney and Robinson 1989, Hamilton et al. 1987, Herman et al. 1986).
Following application of phenol to a shallow pond mesocosm, Giddings et al. (1984, 1985) found and indirectly-caused increase in the dominance of the taxa Euglena, Phacus, Gonium , Coleochaeta, and Scenedesmus. Oil was predicted by Werner et al. (1985) to shift community composition from algae to heterotrophic microbes. In other studies, tolerance to high arsenic levels was demonstrated by Chlorella vulgaris (Maeda et al. 1983) and in a lake contaminated with copper, lead, and zinc, Rhizosoenia eriensis bloomed while other species declined (Deniseger et al. 1990). Algal assays using highway runoff have demonstrated chronic toxicity in several cases, probably due to combined effects of heavy metals, road salt, and sediment (FHWA 1988).
Acidification. Algal responses to acidification in lakes are summarized by Stokes (1981, 1984). Algal species richness can decline in acidified akes, particularly in the presence of heavy metals (Dillon et al. 1979). Filamentous algae typically show a proportionate increase, and the genus Mougeotia has been reported to be a useful indicator of acidification. Nonetheless, algal production can be relatively high in some naturally acidic wetlands (e.g., Bricker and Gannon 1976).
Thermal Alteration. From knowledge of algal responses in other surface waters (e.g., Squires et al. 1979), it appears likely that algae in wetlands would respond dramatically to thermal effluents, and that suitable assemblages of "most-sensitive species" could eventually be identified.
Dehydration/Inundation. Drawdown of wetland water levels often concentrates nutrients and mobilizes nutrients locked up in exposed peat. This can cause algal blooms in remaining surface water (Schlosser and Karr 1981, Schoenberg and Oliver 1988). Inundation may have the opposite effect, diluting nutrients, reducing nutrient mobilization via oxidation, 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 algae, and provides increased opportunities for dispersal of some algal taxa into and out of a wetland. In some Prairie pothole wetlands, metaphyton (unattached, filamentous algae that float in a visible mat) and periphyton (attached algae) increase, while phytoplankton decreases, as higher water levels reduce the density of vascular plants and increase light penetration (Hosseini 1986).
Other Human Disturbance. In other surface waters, species suggestive of "clean" water includeMelosira islandica and Cyclotella ocellata. Algal or microbial species that can indicate "contaminated" water include Chlamydomonas, Euglena viridis, Nitzschia palea, Microcystis aeruginos, Oscillatoria tenuis, O. limosa, Stigeocloneum tenue, and Aphanizomenon flos-aquae (Prescott 1968, APHA 1980).
Salinization; Sedimentation/Burial; Vegetation Removal; Fragmentation of Habitat. We found no explicit information on algal indicators or algal community response to these stressors in wetlands. From knowledge of algal responses in other surface waters (e.g., Dickman and Gochnauer 1978), it appears likely that algae in wetlands would respond dramatically to many of these stressors, and that suitable assemblages of "most-sensitive species" could be identified.
5.2 Sampling Equipment and Methods
Factors that could be important to standardize (if possible) among collections of algal communities include:
- age of wetland (successional status)
- light penetration (water depth, turbidity, shade)
- hydraulic residence time
- conductivity and baseline chemistry of waters
- current velocity
- leaf surface area and stand density of associated vascular plants
- density of grazing aquatic invertebrates
- typical duration and frequency of wetland inundataion
- time elapsed since last runoff or inundation event.
Standard protocols for algal monitoring are available, although uncertainty exists concerning their applicability to wetlands. One is presented by the manual of Britton and Greeson (1988).
Replication requirements in wetland algal studies are significant, due to large spatial and temporal variability. Some investigators have recommended that samples that will be assumed to come from the same time period should be sampled within a time period less than the hydraulic residence time of the wetland. Rapid succession in dominant flagellate species was typical of shallow, eutrophic ponds where conditions fluctuate quickly (Estep and Remsen 1985).
Sampling can occur at any season, but algal biomass is often greatest during the mid to late growing season (e.g., Crumpton 1989, Hooper 1978, Hooper-Reid and Robinson 1978a, b). In deeper waters, it may be advisable to sample phytoplankton at mid-day, due to vertical movements at other times (Estep and Remsen 1985). The pigment, chlorophyll-a is sometimes sampled from the water column as an indicator of algal biomass, but yields little information on community structure. Rabe and Gibson (1984) found greater phytoplankton density in a shallow vegetated pond than at nonvegetated sites, but species composition was similar. In contrast, Seelbach and McDiffett (1983) found that a pond with submerged vegetation had more taxa but lower population density than an open-water pond.
Algal communities in wetlands are generally collected from sediment samples, water column samples, artificial substrates, or natural organic substrates. Methods are described as follows.
Sediment sampling. Algae can be sampled from sediment surfaces in all types of wetlands. Piston corers, plastic syringes, or other suction devices are typically used.
Water column sampling. Any wetland types that have surface water permanently or seasonally can be sampled. Samples from surface waters commonly involve use of volumetric containers or fine-mesh nets. Vertically-integrating, automated samplers can be used (e.g., Schoenberg and Oliver 1988). Surface microlayers (top 250-440 micrometers) can be sampled using fine nets or screens mounted on a frame (e.g., Estep and Remsen 1985). In flowing-water wetlands, fine nets can be mounted to intercept algae carried by currents.
Artificial substrates. Artificial substrates (initially sterile materials placed in a wetland and subjected to natural colonization) may integrate algal assemblages from a large variety of microhabitats. As with microbial communities, algal communities can be monitored by installing plexiglass plates or similar inert, sterile surfaces in any wetlands that have surface water permanently or seasonally, and allowing them to be colonized by attached algae over a period of several weeks. Substrates are then retrieved and community structure is analyzed (e.g., Hooper-Reid 1978).
Natural substrates. Natural organic substrates, particularly those in shallow water, may contain a great biomass of algae. Epiphytic and epibenthic algae are often sampled using a quadrat approach, in which a frame is placed over a standard-sized area of bottom or a standard volume of the water column is enclosed. Frame sizes of 10 x 10 cm (Atchue et al. 1983) and 1-2 m2 (Schoenberg and Oliver 1988) have been used. If algal density is to be estimated accurately, 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.
5.3 Spatial and Temporal Variability, Data Gaps
In no region of the country, and in no wetland type, have data on algal community structure been uniformly collected from a series of statistically representative wetlands. Thus, it is currently impossible to state what are "normal" levels for parameters such as seasonal density, species richness, and their temporal and spatial variability.
Studies that have compared algal community structure among wetlands (spatial variation) apparently include only Hern et al. (1978) who studied the Atchafalaya system in Louisiana, and Sykora(1984), who reported a range of 9 to 21 phytoplankton taxa per ml (mean=9, S.D.=2.3) from a series of six West Virginia wetlands. Phytoplankton density (cells per ml) ranged from 19 to 2581 (mean=203, S.D.=126). Atchue et al. (1982) found 56 taxa of phytoplankton in 8 springtime collections from a one-hectare temporary swamp pool in Virginia. We encountered no journal papers that quantified measurement errors or year-to-year variation in microbial community structure in U.S. inland wetlands.
Even qualitatively, lists of "expected" wetland algal taxa appear not to have been compiled for any region or wetland type. Limited qualitative information may be available by wetland type from the "community profile" publication series of the U.S. Fish and Wildlife Service (USFWS)(Appendix C).