Several indicators of the health of an ecosystem may be used to assess the hazard of a particular stressor. These indicators include rates of primary production; trophic structure; survival of sensitive species; species diversity; and population of shellfish, fish, and endangered species of birds and mammals. Specific measures include abundance counts, growth rate, survival, reproduction and recruitment (Chapman 1992). Responses to toxic chemical stresses can take place at the cellular, organismal, population or the community/ecosystem level. For aquatic ecosystems, change in community structure is an important ecological concern and appears to be sensitive to toxic chemical response (National Research Council 1993). Changes in community structure have been shown to be sensitive to pollutants (La Point et al. 1984, Giddings et al. 1985). Community function endpoints provide insight into how community metabolism responds to contaminant inputs and how changes in community structure may affect community function. Although structure may change, redundancy mechanisms for maintaining community function commonly exist. Too many biomonitoring studies have resulted in collections of data with little forethought into how the data will be utilized. Harwell and Kelly (1986) devised a set of criteria for selecting indicators of ecosystem effect. The rationale for selecting indicators, not necessarily mutually exclusive, includes 1) intrinsic importance, such as endangered or economically important species; 2) early warning indicators when a rapid response to stress is desired; 3) sensitive indicators, known to predictably respond to a stress; and 4) process indicators, when a community function (i.e. primary production ) is known to respond to a stress. A variety of toxic agents predictably cause changes, with loss of sensitive species (e.g. amphipods ) which may result in domination by opportunistic species (e.g. capitellid worms). Species diversity usually decreases, along with the loss of sensitive species.
Ecosystems vary in their sensitivity to stress by type and region as determined by physical and ecological parameters. Due to dilution, coastal systems may be less sensitive to effects caused by toxic substances, although bioaccumulation may occur (National Research Council 1993). Regional differences in community structure will also make a difference. Marine ecosystems already subject to natural stresses may be more resistant to toxic chemical stress. Areas receiving only small nutrient and sediment inputs and having few major storms that affect the bottom are most sensitive to new toxic chemical stresses.
Trace amounts of metals are naturally present in surface waters from the weathering of rocks and soil. Natural waters normally contain low concentrations of metals (0.1-0.001µg/l) (Chapman 1992 ). Anthropogenic sources increase the concentrations above natural levels. The toxicity of metals in water depends upon the form. As a rule, the ionic form of a metal is the most toxic (Chapman 1992). Toxicity is reduced if the ions are bound into organic or mineral complexes. However, some organic complexes such as methylmercury are actually more toxic than the ionic form. Metals exist in dissolved, colloidal and suspended forms. The distribution of the forms between these phases differ between metals and water bodies depending on pH, buffering capacity, suspended matter and organic content. Water quality assessment programs generally evaluate aluminum, cadmium, chromium, copper, iron, mercury, manganese, nickel, lead, zinc (arsenic and selenium are also usually included, though not strictly metals) as being associated with pollution (Chapman 1992).
Mineral oil and petroleum products are widespread pollutants often responsible for ecological damage in water bodies (Chapman 1992). More than 800 individual compounds have been identified in mineral oils; low and high molecular weight aliphatic, aromatic and naphthenic hydrocarbons, high molecular weight unsaturated heterocyclic compounds as well as numerous others. Oil is distributed in water in different forms: dissolved, film, emulsion, and sorbed fractions (Chapman 1992). Interactions between these fractions are complicated and diverse, and depend upon specific gravities, boiling points, surface tensions, viscosities, solubilities and sorption capabilities. Phenols enter water bodies through industrial waste streams. They are normally easily oxidized chemically, biochemically or photochemically, but can produce toxic effects in fish at concentrations as small as 0.01 mg/l (Chapman 1992).
Most currently used pesticides are synthetic organic compounds which have toxic effects on many different living organisms. There are over 10,000 different pesticides now available which include commonly used fungicides, insecticides, herbicides, and rodenticides. The mode of action for a pesticide is determined by its chemical structure (Chapman 1992). They may be grouped into classes which have similar structures: organochlorine pesticides (OCP), organophosphorous pesticides (OPP), carbamate pesticides, triazine herbicides and chlorphenolic acids to name a few.
Organochlorine pesticides (DDT, DDD, DDE, heptachlor, HCH) were introduced during World War II and were thought to be safer and more effective than the inorganic pesticides they were replacing. They have low water solubility but a high affinity for hydrocarbons and fats and a strong tendency to attach to soil particles. As their use expanded, it was found that they accumulate in the environment and bioaccumulate to toxic concentrations within the food web. Bottom sediments play a significant role in storage and transformation of organochlorine pesticides (Chapman 1992). The reader is referred to the Soil section for information about the role bottom sediments play in adsorption and accumulation of pesticides. Environmental levels of organochlorine pesticides tend to be higher than other pesticides due to their widespread application and long-term stability. Use of many of these pesticides has been restricted or banned; however, environmental concentrations of some, such as DDT, remain high.
Organophosphorus pesticides such as malathion, diazinon and parathion are replacing organochlorine compounds. Although they are highly toxic to humans and non-target organisms, they break down more rapidly in the environment. Breakdown depends upon conditions and half-lives are generally on the scale of a month; however, for many of the organophosphate pesticides, the daughter products, oxon and oxide forms, can be as toxic as the the parent OP. Organophosphate pesticides are readily adsorbed onto suspended matter. Photolysis, hydrolysis, oxidation and enzyme decay processes are the principle mechanisms of decay that result in degradation.
Accumulating evidence suggests that anthropogenic discharges of chemicals and complex mixtures are capable of eliciting endocrine disrupting effects that adversely affect the health of humans and wildlife. Current studies are finding that numerous species experience compromised reproductive fitness and increases in hormone-dependent cancers. Researchers are also concerned that these endocrine disruptors can have other effects such as altered immunity and decreased disease resistance, especially during embryonic and fetal development. These chemicals can alter hormone metabolism resulting in elevated levels of androgens, which may lead to pseudohermaphrodism, developmental abnormalities, and reproductive impairment. Chemicals known or suspected to cause these effects and act as synthetic estrogens include organochlorine pesticides, polychlorinated biphenyls, dioxins and other synthetic chemicals.
Examples of endocrine disruptor effects include abnormalities in embryonic and fetal development in alligators, sex determination, hormonal development and reproductive success of fish, disruption of normal hormonal signaling in birds, sex determination in turtles and infertility in sharks. The extent of effects upon wildlife and human health is just beginning to be studied. As more data become available, the impacts of endocrine disruptors may become one of the most serious environmental problems humans have faced. Attempts to protect species diversity within the remaining natural habitats of the world will become difficult if the health and reproductive capabilities of wildlife are seriously affected by widespread exposure to endocrine disruptors .
Endocrine disrupting effects on wildlife may also be elicited from the anthropogenic concentration and discharge of natural estrogenic compounds (Thompson 1997). With the increasingly high concentration of human populations along water bodies, large amounts of endogenous (natural) estrogens are discharged from point sources into the environment. Estrogenic compounds such as estradiol, estrone, and equilenin are excreted with animal (human) wastes and are found in effects-level concentrations within municipal wastewaters, animal feed lot discharges, food processing plants, and grain processing facilities. Studies are beginning to correlate abnormalities in wildlife embryonic development, sex ratios and reproductive capabilities with exposure to wastewaters containing concentrations of these naturally emitted estrogens (Thompson 1997).
The relatively low level of human development, combined with the large percentage of land area still in a natural state suggests levels of toxic pollutants within the surface waters should be low. However, there are areas of concern. The state of South Carolina has issued a fish consumption advisory for parts of the Combahee, South Edisto, North Edisto and Dawho Rivers due to mercury. The source of mercury contamination is uncertain, although natural occurrence may be partially responsible. Air deposition of mercury, a byproduct of fossil fuel combustion, may be the major source. Mercury discharges cannot be traced to any industry, and some areas of these rivers have a naturally low pH which increases mercury mobility.
The EPA Toxic Release Inventory (TRI) for Colleton County showed the following priority pollutants were discharged into air, water and land within the ACE Basin area environment between 1988 and 1994 (http://www.epa.gov/tri/ ):
The total reported releases of priority pollutants within Colleton County in 1994 were 2,573 kg (5,673 pounds) compared to 31,911 kg (70,351 pounds) in 1988.
Aquatic life uses were not fully supported on the Coosaw River due to zinc concentrations in excess of acute standards for aquatic life. In the Dawho River, aquatic life uses may be threatened due to a very high concentration of zinc combined with a declining trend in pH (SCDHEC 1995b). South Edisto River water samples have also had high concentrations of zinc (SCDHEC 1995a). The compound Diethyl phthalate was detected within water samples from the Combahee River (SCDHEC 1997). USGS National Water Quality Assessment database listed no sites within the ACE Basin with water column pesticide concentrations above detection limits (USGS 1998). Point source discharges may account for most of the non-pesticide contaminants found in these cited cases of concern. Urban and agricultural runoff would be the suspected sources if pesticide contamination was detected. Quarterly sampling by SCDHEC has shown that, in general, heavy metals and organic concentrations in the water column are below detection limits except for the few instances discussed (SCDHEC 1997).
Effects of endocrine disrupting chemicals upon wildlife within the ACE Basin have not been studied. The relatively low overall concentration of endocrine disrupting chemicals in water samples from the area would suggest a low probability of detecting effects. Rhodes (1998) surveyed the South Carolina alligator population for evidence of endocrine disrupting chemical influences. He concluded that reproductive success and juvenile survival does not seem to be influenced by abnormal factors such as those acting on Lake Apopka alligators in Florida. In a study by Wood (1994), alligator eggs from the Bear Island Wildlife Management Area in the ACE Basin were found to have PCB concentrations lower (0.34µg/g, n=19) than concentrations found in eggs from the Yawkey Wildlife Center in Georgetown, SC (2.4 µg/g, n=19) (Rhodes 1998). The mean concentrations of PCBs from both sites were lower than concentrations found in other reptilian eggs studied. Mercury has been shown to affect growth and development, metabolism, and reproduction in fish, wildlife and humans (Rhodes 1998). Fifteen alligator samples from coastal South Carolina were analyzed for the presence of mercury. All samples were found to have mercury concentrations ranging from 0.049 ppm to 3.44 ppm (Rhodes 1998). The U.S. Food and Drug Administration action level for mercury in fish flesh consumption is 1 ppm, but no action level has been established for alligators. Twenty percent of the alligator samples exceeded the action level for fish.
Overall, the presence of toxic substances within ACE Basin area waters appears to be minimal, but potential exists for levels to increase significantly in those areas now undergoing urbanization. Agricultural runoff may be a concern for pesticide pollutants in some areas of the ACE Basin.
Best management practices, as discussed in the Land Use Synthesis Module, and least toxic pest control for agricultural lands may be the best means for controlling the pesticides found in water bodies. Point source discharges of other toxicants are limited by the Clean Water Act mandated regulations. NPDES permitting requires monitoring and treatment of wastewaters containing priority pollutants to levels needed to maintain water quality standards in stream. Endocrine disruption effects caused by persistent estrogen mimicking compounds and natural estrogens must be studied to develop an understanding of needed conservation actions.
Chapman, D. 1992. Water quality assessments. Chapman and Hall, New York, NY.
Giddings, J. M., P. J. Franco, S. M. Bartell, R. M. Cushman, and S. C. Herbes. 1984. Effects of contaminants on aquatic ecosystems: Experiments with microcosms and outdoor ponds, a synthesis report. Oak Ridge National Laboratory Publication No. 2381. Oak Ridge National Laboratory. Oak Ridge, TN.
Harwell, M. A. and C. C. Kelly. 1986. Scientific responsibility in environmental decision-making. Sea Technology 27:28-30.
La Point, T. W., S. M. Melancon, and M. K. Morris. 1984. Relationships among observed metal concentrations, criteria and benthic community structure. Journal Water Pollution Control Federation 56:1030-1038.
National Research Council. 1993. Managing wastewater in coastal urban areas. Committee on Wastewater Management for Coastal Urban Areas, National Academy of Science. National Academy Press, Washington, DC.
Rhodes, W. E. 1998. The health of alligator populations in South Carolina. In R. Kendall. R. Dickerson, J. Giesy, and W. Suk (eds): Principles and processes in endocrine disruptions. Society of Environmental Toxicology and Chemistry 1997 conference proceedings. SETAC Press, Pensacola, FL.
South Carolina Department of Health and Environmental Control. 1997. Watershed water quality management strategy: Savannah and Salkehatchie River basins. Bureau of Water, SC Department of Health and Environmental Control, Columbia, SC. Technical report No. 003-97.
South Carolina Department of Health and Environmental Control. 1995a. Summary of heavy metals concentrations in South Carolina waters and sediments January 1, 1990 to December 31, 1994. Bureau of Water, SC Department of Health and Environmental Control, Columbia, SC. Technical report No. 001-95.
South Carolina Department of Health and Environmental Control. 1995b. Watershed water quality management strategy: Saluda and Edisto River basins. Bureau of Water, SC Department of Health and Environmental Control, Columbia, SC. Technical report No. 003-95.
Thompson, M. 1997. A case for natural estrogens as environmental pollutants. Unpublished presentation. Environmental Studies Program, Medical University of South Carolina, Charleston, SC.
United States Geological Survey. 1998. Pesticides in surface waters of the Santee River basin and coastal drainages of South and North Carolina. Santee River Basin and Coastal Drainages Study Unit, National Water Quality Assessment Program. http://wwwsc.er.usgs.gov/nawqa/pest.html. Accessed October 1998.
Wood, P. D. 1994. Determination of PCBs in eggs and chorioallantoic membranes of loggerhead sea turtles (Caretta caretta) and American alligators (Alligator mississippiensis): Assessment of CAM as an indicator tissue in reptiles. M.A. Thesis. Clemson University, Clemson, SC.