CHAPTER 5 Analysis Of Key Topics—Sources, Behavior, And Transport The preceding overviews of national distribution and trends of pesticides in bed sediment and aquatic biota, and of governing factors that affect their concentrations in these media, leaves many specific questions unanswered. The next two chapters draw on information in the literature reviewed to discuss, in detail, several important topics related to pesticides in bed sediment and aquatic biota. Each key topic falls into one of two categories: (1) sources, behavior, and transport (Chapter 5), or (2) environmental significance (Chapter 6). 5.1 EFFECT OF LAND USE ON PESTICIDE CONTAMINATION The terrestrial environment has a strong influence on the water quality of adjacent hydrologic systems. Both natural and anthropogenic characteristics of the terrestrial environment are important. For example, concentrations of major chemical constituents (such as sulfate, calcium, and pH) in a hydrologic system are influenced by geology, and the concentration of suspended sediment is influenced by soil characteristics, topography, and land cover. Land use activities, such as row crop agriculture, pasture, forestry, industry, and urbanization, also can affect adjacent water bodies. Any pesticide associated with a land use can potentially find its way to the hydrologic system and, if the pesticide has persistent and hydrophobic properties (see Section 5.4), it will tend to accumulate in bed sediment and aquatic biota. The following section addresses the observed link between land use and the detection of pesticides in bed sediment and aquatic biota. Four types of land use will be discussed: agriculture, forestry, urban areas and industry, and remote or undeveloped areas. In many cases, forested areas also could be described as remote or undeveloped areas. The critical distinction here, however, is that many forested areas have been managed with the use of pesticides whereas remote and undeveloped areas have not. 5.1.1 AGRICULTURE By far the largest use of most pesticides, both presently and historically, has been in agriculture (Aspelin and others, 1992; Aspelin 1994). The soils of many agricultural areas still © 1999 by CRC Press LLC contain residues of hydrophobic, persistent pesticides that were applied during the 1970s or earlier. This was documented in 1970 for 35 states, mostly east of the Mississippi River (Crockett and others, 1974), in 1985 in California (Mischke and others, 1985), and during 1988–1989 in Washington (Rinella and others, 1993). In the California study (Mischke and others, 1985), only fields with known previous DDT use were targeted. This study obtained 99 soil samples from fields in 32 counties. Every sample analyzed contained residues of total DDT (the sum of DDT and its transformation products). The investigators compared the concentrations of the parent DDT with the concentrations of its transformation products (DDD and DDE) and found that the ratio of the parent DDT to total DDT was 0.49. That is, 49 percent of the total DDT remaining in the soils at least 13 years after use still existed as the parent compound. In the U.S. Environmental Protections Agency’s (USEPA) National Study of Chemical Residues in Fish (NSCRF), which measured fish contaminants at sites in different land-use categories (such as agricultural sites, industrial and urban sites, paper mills using chlorine, other paper mills, and Superfund sites), sites in agricultural areas had the highest mean and median concentrations of p , p ′ -DDE in fish, as well as four of the top five individual fish sample concentrations. Agricultural sites also had the second highest mean concentration of dieldrin in fish (second to Superfund sites), as well as two of the top five individual sample concentrations. Soils containing residues of DDT and similar recalcitrant pesticides from past agricultural use constitute a reservoir for these pesticides today; they have been, and will continue to be, a source of these compounds to hydrologic systems, thus leading to contamination of surface water, bed sediment, and aquatic biota. Those pesticides currently used in agriculture (Table 3.5) are not as persistent as the restricted organochlorine compounds. As discussed in Section 3.3, some moderately hydrophobic, moderately persistent pesticides have been detected in bed sediment and aquatic biota, although at lower detection frequencies than the more persistent organochlorine compounds. It is probable that additional pesticides with moderate water solubilities and persistence may be found in bed sediment or aquatic biota if they are targeted in these media (see Section 5.4), especially in high use areas. A few moderately hydrophobic, moderately persistent compounds were analyzed in fish by the NSCRF (U.S. Environmental Protection Agency, 1992a): dicofol, lindane, α -HCH, and methoxychlor (organochlorine insecticides or insecticide components); chlorpyrifos (organophosphate insecticide); and trifluralin, isopropalin, and nitrofen (herbicides). Of these compounds, several were found in association with agricultural areas. Agricultural sites had the highest mean and maximum concentrations of dicofol and chlorpyrifos, and they had the highest mean concentration of trifluralin, in fish. Moreover, sites with the highest trifluralin residues in fish were in states with the highest agricultural use of tri- fluralin (Arkansas, Illinois, Iowa, Minnesota, Missouri, North Dakota, South Carolina, Tennes- see, and Texas). In California’s Toxic Substance Monitoring Program, which monitored pesti- cides in fish and invertebrates from over 200 water bodies throughout the state, the highest concentrations of several currently used pesticides in fish during 1978–1987 were from two intensively farmed areas (Rasmussen and Blethrow, 1990). These pesticides are the insecticides chlorpyrifos, diazinon, endosulfan, and parathion, and the herbicide dacthal. The highest residues of dacthal in whole fish analyzed by the Fish and Wildlife Service’s (FWS) National Contaminant Biomonitoring Program (NCBP) also occurred in intensively farmed areas (Schmitt and others, 1990). © 1999 by CRC Press LLC 5.1.2 FORESTRY A number of studies monitored one or more pesticides in forest streams or lakes after known application. Most of these studies were at sites in the forests of the southeastern United States (e.g., Yule and Tomlin, 1970; Neary and others, 1983; Bush and others, 1986; Neary and Michael, 1989), northwestern United States (e.g., Sears and Meehan, 1971; Moore and others, 1974), or Canada (e.g., Kingsbury and Kreutzweiser, 1987; Sundaram, 1987; Feng and others, 1990; Kreutzweiser and Wood, 1991; Sundaram and others, 1991). The majority of these studies can be described as field experiments, in which a known amount of a certain pesticide was applied to a section of a watershed, with subsequent sampling of water, bed sediment, or aquatic biota for a period of weeks to years. These studies are considered process and matrix distribution studies and are described in Table 2.3 (if they were conducted in United States streams and they sampled bed sediment or aquatic biota). Few, if any, studies have reported on the ambient concentrations of pesticides in bed sediment or aquatic biota after routine use of pesticides, so little is known about the long-term presence of pesticides in streams from forest applications. On the basis of the reported field experiments and information on pesticide use in forestry, a few conclusions can be drawn. The choice of chemicals used for forest applications has changed over time (Freed, 1984). Before the mid-1940s, the only pesticides that were used were inorganic compounds. Organic pesticides were introduced after World War II. Aerial spraying of pesticides began following the availability of suitable airplanes. The chlorophenoxy acid herbicides, 2,4-D and 2,4,5-T, and the organochlorine insecticide, DDT, were the first of the organic pesticides to be widely used. In subsequent years, a wide variety of herbicides and insecticides were introduced into forestry use. Most of the major classes of herbicides were represented, including triazines, ureas, uracils, and chlorophenoxy acids. The insecticides used included most of the organochlorine compounds and numerous organophosphate, carbamate, and pyrethroid compounds. Since the 1980s, the use of chemical pesticides in forestry has declined (Larson and others, 1997). The chemical insecticides have largely been replaced by biological pesticides. The current use of pesticides in forestry (Section 3.2.2) and forestry as a source of pesticides in surface water systems (Section 4.1.1) were previously discussed. The potential impacts on water quality are covered in more detail in Larson and others (1997). The pesticides used in forestry since the 1940s may have caused some environmental impact at the time of application. However, many of these pesticides do not persist long in forest soils or streams, so are unlikely to have lasting or long-term effects on stream biota after a period of time (days, months, or years, depending on the chemical) has elapsed since application. The exceptions are pesticides that are hydrophobic and recalcitrant (and thus long-lived), such as the organochlorine insecticides. Because of their physical and chemical properties (see Section 5.4), organochlorine insecticides may persist in bed sediment and aquatic biota of forest streams, and forest soils containing organochlorine insecticide residues may be washed into the stream for many years after the period of application. Also, as with pesticides applied in agricultural areas (see Section 3.3.2), there is potential for detection of moderately hydrophobic, moderately persistent silvicultural pesticides in bed sediment or biota, especially in high-use areas. Triclopyr is now the highest-use herbicide on national forest land. The next most commonly used herbicides in national forests in 1992 were 2,4-D, hexazinone, glyphosate, and © 1999 by CRC Press LLC picloram (Larson and others, 1997). Except for Bacillus thuringiensis var. Kurstaki (Bt), car- baryl was the highest-use insecticide in national forests in 1992 (Larson and others, 1997). Of these compounds, only 2,4-D was targeted in sediment or aquatic biota at more than 30 (total) sites in all the monitoring studies reviewed (Tables 3.1 and 3.2). When data from all monitoring studies were combined, 2,4-D was detected in 1 percent of (825 total) sediment samples and in 5 percent of (44 total) biota samples. Of the other recently used pesticides, picloram was detected bed sediment in 2 percent of (53) samples; detection data were not reported for biota. Carbaryl was detected in aquatic biota (11 percent of 27 samples), but not in bed sediment (only 3 samples analyzed). Glyphosate was not detected in any of 19 total bed sediment samples; data for biota were not reported. Triclopyr and hexazinone were not targeted in bed sediment or aquatic biota in any of the monitoring studies reporting detection data. Five process and matrix distribution studies (or field experiments) in forest streams provide some indication of the behavior of organochlorine insecticides, pyrethroid insecticides, and other selected pesticides following application in forestry. In one study in New Brunswick, Canada (Yule and Tomlin, 1970), DDT and its transformation products, DDE and DDD, were studied in water and bed sediment of a stream after application to nearby forests for the control of Spruce budworm. One motivation for this study was that fish-kills occurred following the use of DDT in forests in this area. The stream had high concentrations of DDT in the surface of the water column immediately after application, but these subsided to the background concentration (about 0.7 µ g/L) after a few hours. The deeper stream water (12–18 in. below the surface) did not show the same immediate DDT concentration spike; however, DDT levels there were relatively consistent for 2 years following the application. Twelve months after application, every bed sediment sample (18 total) collected from the vicinity of the site of application to the mouth of the river, about 50 mi downstream, had measurable concentrations of total DDT. The average bed sediment concentration was about 12 percent of the forest soil concentration on a dry weight basis. There was a trend of decreasing concentration downstream, and also a change in the ratio of DDT/total DDT. As the distance from the point of application increased, the transformation products constituted a greater percentage of the total DDT, indicating in-stream transformation. Unfortunately, no time series data were presented for the bed sediment. The authors suggested that DDT persists in forest soils, predominately as the parent compound, and that the long-term transport to streams is through runoff of soil particles. The presence of DDT components in the bed sediment throughout the river system 1 year after application, and the presence of DDT components in the water 2 years after application, suggest that there is long-term storage of DDT in the forest soil and in the bed sediment of the river system, and that the soil and bed sediment constitute a constant source of contaminant to the river water. Prior to its cancellation in the early 1980s, endrin was used in forestry as a coating on aerially applied tree seeds to protect them from seed-eating rodents. One study (Moore and others, 1974) examined the presence of this compound in the water and aquatic biota of two Oregon watersheds after seeding. The actual amount of endrin applied to the watersheds was estimated to be 2.5 to 10 grams a.i. per hectare. Endrin was observed consistently in the stream water for about 9 days (maximum concentration was about 12 ng/L), then was nondetectable until a high flow period about 21 days after application. At this time, it was detected in the water again. This second period of detection suggests that the endrin was stored either in the forest soils or in the bed sediment of the stream and then released with higher streamflow. Fish (coho salmon [ Onchorhynchus kisutch ] and sculpins [family Cottidae]) and various unidentified aquatic © 1999 by CRC Press LLC insects were analyzed for endrin. Because of sample contamination, the results are somewhat ambiguous. The authors did conclude that endrin was present in all biotic samples obtained within days after application. Samples collected 12 and 30 months after the application of endrin did not contain detectable traces of endrin. Bed sediment was not collected during this study. A third example is the study of permethrin in Canadian streams (Kreutzweiser and Wood, 1991; Sundaram, 1991). Permethrin, a synthetic pyrethroid, is known not only for its high insecticidal activity and its ability to control lepidopterous defoliators, but also for its high acute toxicity to fish and strong sorption tendencies. Kreutzweiser and Wood (1991) examined the presence of permethrin in a forest stream after aerial application. They detected the compound in water, bed sediment, and fish. The concentration in water declined with time and distance from application. Permethrin was seldom seen in the bed sediment of the stream (only 8 percent of the samples). Atlantic salmon ( Salmo salar ), brook trout ( Salvelinus fontinalis ), and slimy sculpin ( Cottus cognatus ) were analyzed, and permethrin was detected in about half of the samples during the first 28 days after application. The fish were sampled again 69 to 73 days after application and no traces of permethrin were detected. Sundaram (1991) studied the behavior of permethrin by adding it directly into a forest stream. He found that it was not detected in the stream water near the site of application after 5 hours and that it was seldom detected in the bed sediment of the system, probably because of the low sediment organic carbon content. Sundaram (1991) did detect permethrin in aquatic plants (water arum, Calla palustris ), stream detritus, caged crayfish ( Orconectes propinquus ), and caged brook trout collected during the study (up to 7 to 14 hours after application). No permethrin was detectable in caged stoneflies ( Acroneuria abnormis ) throughout the study duration (14 hours). Permethrin also was detected in invertebrate drift collected 280–1,700 m downstream of the application point. The longer-term presence of permethrin in this system was not studied. In a fourth example, 2,4-D was sprayed on clearcut forested lands in Alaska (Sears and Meehan, 1971). The results of this study show potential for at least initial accumulation in biota. Residues of 2,4-D were detected in river water samples (up to 200 µ g/L), and in a single composite sample of coho salmon fry (500 µ g/kg), collected 3 days after spraying. Unfortunately, later samples were not taken, so no information is provided on dissipation rates. Finally, a dissipation study of the organophosphate pesticide chlorpyrifos-methyl was conducted in a forest stream in New Brunswick, Canada (Szeto and Sundarum, 1981). The results of this study indicate that there is potential for initial accumulation in stream bed sediment and aquatic biota, but that residues are unlikely to persist. After aerial application, chlorpyrifos-methyl residues persisted in balsam fir foliage and forest litter for the duration of the experiment (125 days). Residues in bed sediment (10–180 µ g/kg dry weight) persisted for at least 10 days; at the next sampling time (105 days post-application), residues in sediment were nondetectable (less than 1 µ g/kg wet weight). In stream water, chlorpyrifos-methyl dissipated rapidly within the first 24 hours after application, and it was not detectable in water (less than 0.02 µ g/L) after four days. Residues of up to 46 µ g/kg chlorpyrifos-methyl were detected in fish (slimy sculpin and brook trout); only trace levels (less than 3 µ g/kg wet weight) were detected after 9 days, and chlorpyrifos-methyl was nondetectable (less than 1.5 µ g/kg wet weight) after 47 days. Concentrations in brook trout were consistently higher than in slimy sculpin sampled at the same time. The results from these limited studies suggest that the behavior of pesticides in forested streams are in agreement with their behavior in agricultural streams. DDT and its transformation © 1999 by CRC Press LLC products appear to have the longest residual time in the bed sediment. Endrin, permethrin, and chlorpyrifos-methyl, although persisting for days to months in the bed sediment or biota, gradually dissipated. Carbaryl, 2,4-D, and picloram are moderate in water solubility, but would be expected to degrade in the environment eventually. Moderately hydrophobic, moderately persistent pesticides may be expected to be found in some bed sediment or biota samples, especially in areas of high or repeated use. 5.1.3 URBAN AREAS AND INDUSTRY Another source of pesticides to surface water systems, and thus to bed sediment and aquatic biota, is from urban areas. Pesticides are applied to control pests for public health or aesthetic reasons in and around homes, yards, gardens, public parks, urban forests, golf courses, and public and commercial buildings (Buhler and others, 1973; Racke, 1993). The available data suggest that the patterns of urban pesticide use have changed during the past few decades, much as has pesticide use in agriculture and forestry. Many of the high use organochlorine insecticides have been banned and replaced by organophosphate, carbamate, and pyrethroid insecticides. The use of herbicides in and around homes and gardens has increased, whereas herbicide applications to industry, commercial, and government buildings and land have decreased (Aspelin, 1997). The major pesticides used in and around homes and gardens in 1990 are listed in Table 3.5. An examination of Table 3.5 shows that most of the organochlorine pesticides that are commonly observed in bed sediment (Figure 3.1) and aquatic biota (Figure 3.2) are no longer used in urban areas, with the exception of dicofol, chlordane, heptachlor, lindane, and methoxychlor. The commercial use of existing stocks of chlordane in urban environments was banned in 1988, and homeowner use of existing stocks is likely to have declined since then also. Although the kind of data in Table 3.5 does not exist for the time period of the 1950s through mid-1970s, it is known that many of the organochlorine insecticides had significant urban uses, including aldrin, chlordane, DDT, dieldrin, endosulfan, heptachlor, and lindane (Meister Publishing Company, 1970). In 1970, lindane was used predominantly in the urban environment; there was also considerable urban use of chlordane (Meister Publishing Company, 1970). It seems that endrin was the exception, with little or no urban use. Of the moderately hydrophobic, moderately persistent pesticides that have been observed, when targeted, in sediment or aquatic biota, several are used in and around the home and garden (Table 3.5). These include chlorpyrifos, diazinon, carbaryl, permethrin, and 2,4-D. A number of local-scale studies have monitored pesticides in the sediment or aquatic biota of urban areas. Mattraw (1975) examined the occurrence and distribution of dieldrin and DDT components in the bed sediment of southern Florida. The study area included the urbanized areas on the Atlantic coast (such as Miami and Fort Lauderdale), the Everglades water conservation area and two nearby agricultural areas. Mattraw reported the data as concentration frequency plots, shown in Figures 5.1 and 5.2. In the case of DDD (Figure 5.1), urban areas had a mean concentration and a general distribution between those of the two agricultural areas, and well above those of the undeveloped area. In the case of dieldrin (Figure 5.2), the urban areas had a mean bed sediment concentration and a general concentration distribution greater than all other land use activities. In another example, Kauss (1983) measured 15 different organochlorine insecticides and transformation products in the Niagara River below Buffalo, New York. This is an area with many large chemical production facilities. It is thought that some of the chemicals © 1999 by CRC Press LLC in the sediment of this river are due either to transport from Lake Erie (the source of water for the Niagara River) or to localized inputs. One example of a potential localized input is disposal of 1,700 metric tons of endosulfan at disposal sites in the area. In another urban area study, Thompson (1984) reported DDE, DDE, DDT, dieldrin, heptachlor, methoxychlor, silvex, and 2,4-D in the sediment of the Jordan River in Salt Lake City, Utah. Pariso and others (1984) reported that DDT and chlordane were observed in bed sediment and in various species of fish collected from the Milwaukee Harbor and Green Bay urban areas of Wisconsin, during a study of contaminants in the rivers draining into Lake Michigan. Lau and others (1989) reported the presence of trans -chlordane, DDE, DDD, and DDT in the suspended sediment of the St. Clair and Detroit rivers on the Michigan and Ontario border. Fuhrer (1989) reported the presence of chlordane, DDD, DDE, DDT, and dieldrin in the bed sediment of the Portland, Oregon harbor. Capel and Eisenreich (1990) reported concentrations of α -HCH, DDE, DDD, and DDT in the bed sediment and tissues of mayfly ( Hexagenia ) larvae from the harbor in Lake Superior at Duluth, Minnesota. Crane and Younghaus-Hans (1992) detected oxadiazon residues in fish (red shiner, Cyprinella lutrensis ) and bed sediment from San Diego Creek, California. Oxadiazon was also detected in transplanted clams ( Corbicula fluminea ) in the San Diego Creek and in transplanted mussels ( Mytilis californianus ) in the receiving estuary, Newport Bay. Oxadiazon is widely used in landscape and rights-of-way maintenance in California, and the high residues observed in this study were attributed to its use on golf courses upstream of the study area. Although there have been numerous local-scale studies, there has been no systematic large- scale study of pesticides in the bed sediment or aquatic biota of urban freshwater hydrologic systems. The National Oceanic and Atmospheric Administration’s (NOAA) National Status and Trends (NS&T) Program targeted coastal and estuarine sites near urban population centers, and found a correlation between most organic contaminants in bottom sediment and human population levels (National Oceanic and Atmospheric Administration, 1991). However, there is no comparable nationwide study of pesticides in bed sediment or aquatic biota from rivers in urban areas. In the U.S. Geological Survey (USGS)–USEPA’s Pesticide Monitoring Network (PMN), which sampled bed sediment from major United States rivers, only 10 of about 180 sites sampled between 1975–1980 were in urban areas (Gilliom and others, 1985). Nonetheless, two of these urban sites (Philadelphia, Pennsylvania, and Trenton, New Jersey) were among the 10 sites with the highest frequency of pesticide detection. The only national-scale study of pesticides in rivers near urban centers was the USEPA’s Nationwide Urban Runoff Program (NURP), which analyzed water samples for pesticides in urban areas nationwide during 1980– 1983 (Cole and others, 1983, 1984). The NURP samples were analyzed for the priority pollutants, which include 20 organochlorine insecticides or transformation products, at 61 residential and commercial sites across the United States. Of these 20 organochlorine insecticides, 13 were observed in at least one water sample. The most frequently observed organochlorine insecticides were α -HCH (in 20 percent of samples), endosulfan I (in 19 percent), pentachlorophenol (in 19 percent), chlordane (in 17 percent), and lindane (in 15 percent). During this time period, all of these chemicals were still in active use in urban areas. Because of the hydrophobicity of these compounds, their detection in the water column suggests that they also would have been present at detectable levels in bed sediment and aquatic biota in these urban environments. Although many monitoring studies have reported the frequent detection of organochlorine pesticides in bed sediment, aquatic biota, and water in urban areas, the actual sources of these © 1999 by CRC Press LLC pesticide residues are not completely known. Since the organochlorine insecticides had both extensive urban and agricultural uses, their presence in urban areas could have been derived from either source, since many urban areas are located downstream from agricultural areas. Conversely, some rivers flowing through agricultural areas may be located downstream of urban areas. Examples are the Mississippi River below Minneapolis and St. Paul, Minnesota, and below St. Louis, Missouri. In such cases, residues may derive from urban, as well as from agricultural, origin. It is reasonable to suppose that most pesticides currently in bed sediment and aquatic biota in urban areas are derived from both agricultural and urban uses, although the relative contribution of each of the two sources probably varies by location and compound. 1,000 100 10 1.0 0.1 0.01 0 102030405060 708090 100 Samples that equal or are less than the value indicated, in percent DDD concentration, in ␮g/kg Agricultural area near Everglades Urban area Everglades area Eastern agricultural area Undeveloped Big Cypress watershed Figure 5.1. Concentration frequency plot for DDD in bed sediment from agricultural, urban, and undeveloped areas in southern Florida (1968–1972). Redrawn from Mattraw (1975) with permission of the author. © 1999 by CRC Press LLC 5.1.4 REMOTE OR UNDEVELOPED AREAS Pesticides, particularly the organochlorine insecticides, are often observed in bed sediment and aquatic biota in remote areas of the United States and of the rest of the world. Their presence in remote or undeveloped areas is seldom due to local use, but rather to atmospheric transport and deposition. Majewski and Capel (1995) have reviewed the presence and movement of pesti- cides in the atmosphere and the deposition processes involved in their delivery to remote areas. For some pesticides, particularly the organochlorine insecticides, regional atmospheric transport is common and serves as a mechanism to disperse them throughout the world, particularly toward the polar regions. Agricultural area near Everglades Urban area Everglades area Eastern agricultural area Undeveloped Big Cypress watershed 1,000 100 10 1.0 0.1 0.01 10 20 30 40 50 60 70 80 90 1000 Sample that equal or are less than value indicated, in percent Dieldrin concentration, in ␮g/kg Figure 5.2. Concentration frequency plot for dieldrin in bed sediment from agricultural, urban, and undeveloped areas in southern Florida (1968–1972). Redrawn from Mattraw (1975) with permission of the author. © 1999 by CRC Press LLC Pesticides are introduced into the atmosphere either by volatilization or wind erosion. Once they are in the atmosphere, they can either be deposited locally (in the range of tens of kilo- meters) or move into the upper troposphere and stratosphere for more widespread regional, or possibly global, distribution. Once in the upper atmosphere, the global wind circulation patterns control their long-range transport. The general global longitudinal circulation is a form of ther- mal convection driven by the difference in solar heating between equatorial and polar regions. Over the long-term, upper air masses tend to be carried poleward and descend into the subtrop- ics, subpolar, or polar regions. These air masses are then carried back toward the tropics in the lower atmosphere (Levy, 1990). Once in the atmosphere, the residence time of a pesticide depends on how efficiently it is removed by either deposition or chemical transformation. Atmos- pheric deposition processes can be classified into two categories: those involving precipitation (wet deposition) and those not involving precipitation (dry deposition). The effectiveness of a particular removal process depends on the physical and chemical properties of the pesticide, the meteorological conditions, and the terrestrial or aquatic surface to which deposition is occurring. Risebrough (1990) described the airborne movement of pesticides from their point of applica- tion as a global gas-chromatographic system where pesticide molecules move many times between the vapor-soil-water-vegetation phases, maintaining an equilibrium of chemical poten- tial between these phases. That is, after a pesticide is deposited from the atmosphere to a terres- trial or aquatic surface, it can reenter the atmosphere and be transported and redeposited down- wind repeatedly until it is chemically transformed or globally distributed. Virtually all studies of pesticides in remote areas have been conducted on remote lakes and oceans, rather than on rivers and streams. A few examples of these studies will be presented to illustrate the global nature of atmospheric deposition. One of the earliest reports that attributed the presence of DDT in a remote surface water body to atmospheric deposition was a study by Swain (1978) conducted in the national park in Isle Royale, Michigan. Although this island is in Lake Superior and is removed hundreds of kilometers from agricultural uses of DDT, DDT was found in the water, sediment, and fish (lake trout, Salvelinus namaycush , and lake whitefish, Coregonus clupeaformis ) of Siskitwit Lake on Isle Royale. The probable explanation for this contamination was through atmospheric deposition. Organochlorine contamination of air, snow, water, and aquatic biota in the Arctic has been extensively studied (Hargrave and others, 1988; Patton and others, 1989; Bidleman and others, 1990; Gregor, 1990; Muir and others, 1990) and also is attributed to atmospheric transport. All of the common organochlorine insecticides have been observed in Arctic studies, but the two most prevalent were α -HCH and lindane. These are the two organochlorine insecticides with the highest vapor pressures and their abundance sup- ports the idea of the global gas-chromatographic effect of pesticides being transported to the polar regions described above. Although organochlorine concentrations in the Arctic water are low, these contaminants bioaccumulate in aquatic biota and appear to be magnified in aquatic and terrestrial food webs, reaching quite elevated levels in the Arctic mammals. Addison and Zinck (1986) found that the DDT concentration in the Arctic ringed seal ( Phoca hispida ) did not decrease significantly between 1969 and 1981, while the concentration of polychlorinated biphe- nyls (PCB) did decline. They attributed this to continued atmospheric deposition of DDT from its use in areas of eastern Europe during this time, compared with declining global PCB use. © 1999 by CRC Press LLC [...]... chain The dotted line indicates a fugacity ratio of 1 The compounds plotted are: cis-chlordane; trans-chlordane; p,p′-DDE; α-HCH; 2,4 ,5, 2′,4′ ,5 -hexachlorobiphenyl; lindane; mirex; octachlorostyrene; 2,4 ,5, 2′ ,5 -pentachlorobiphenyl; 2,3,4 ,5, 6-pentachlorotoluene; 1,2,3,4-tetrachlorobenzene; 2,3,2′,3′-tetrachlorobiphenyl; 2 ,5, 2′ ,5 -tetrachlorobiphenyl; 1,2,4-trichlorobenzene; 2 ,5, 2′-trichlorobiphenyl... heptachlor epoxide; lind, lindane; me clpy, chlorpyrifos-methyl; meth, 2,2-bis-(4-methylphenyl )-1 ,1,1-trichloroethane; metrib, metribuzin; µg/L, microgram per liter; nitroben, nitrobenzene; parat, parathion; 3-PCB, 2 ,5, 2′-trichlorobiphenyl; 4-PCB, 2 ,5, 2′ ,5 -tetrachlorobiphenyl; 5- PCB, 2,4 ,5, 2′ ,5 pentachlorobiphenyl; PCP, pentachlorophenol; prop, propoxur; tox, toxaphene; triflu, trifluralin Redrawn from... Boese and others, 1996 Clam (Macoma nasuta) Laboratory PCB- 153 HCB Boese and others, 19 95 Clam (M nasuta) Laboratory PCB congeners HCB McFarland and others, Fish (Japanese medaka) 1994 Laboratory PCB -5 2 Pruell and others, 1993 Clam (M nasuta) Laboratory PCB- 153 2,3,7,8-TCDD 31.8 30.9 Laboratory PCB- 153 2,3,7,8-TCDD 31.4 Laboratory PCB- 153 2,3,7,8-TCDD 32.1 30.7 Muir and others, 1992 Mussel (Anodonta grandis)... (location) some laboratory Chemical and field Mean BSAF Field (Rhode Island) trans-Chlordane cis-Chlordane p,p′-DDD Aroclor 1 254 5. 9 4.2 4.2 4 Field (Rhode Island) trans-Chlordane cis-Chlordane p,p′-DDD Aroclor 1 254 4 .5 4 4 3.3 1Mean value for all PCB congeners in fine sediment 3Estimated by eye from figure in reference 4 Exposed for 10–24 days 5Carcass (minus gills and gastrointestinal tract) 6Range of mean... Weininger, 1978; Thomann, 1981; Biddinger and Gloss, 1984) and the importance of dietary sources (Thomann and Connolly, 1984; Thomann, 1989) The food chain model developed by Thomann (1989) contains four trophic levels (above phytoplankton), and assumes steady-state conditions and uptake from water and food For PeCB HCB 1,2,4 , 5- TeCB 1,2,3,4-TeCB Log BCF 4 1,3 , 5- TCB 1,2,4-TCB 1,2,3-TCB 3 1,3-DCB 1,4-DCB... congeners PCB- 153 Ferraro and others, 1990 Clam (M nasuta) Laboratory p,p′-DDE p,p′-DDD Aroclor 1 254 70.7–2.8 70 .52 –1.0 70 .5 1.8 Lake and others, 1990 Clam (Mercenaria mercenaria) and polychaete (N incisa) Field (New York, Massachusetts, and Rhode Island) Aroclor 1 254 PCB- 153 81.6 84.6 McElroy and Means, 1988 Bivalve (Yoldia limatula) Laboratory Hexachloro-PCB Polychaete (N incisa) Laboratory Hexachloro-PCB... include the following: dieldrin in aquatic invertebrates in the Rocky River, South Carolina (Wallace and Brady, 1971); PCBs in cod (Gadus morhua) (livers and fillets) and prey organisms from the western Baltic Sea (Schneider, 1982); and organochlorine residues in amphipods and other stream animals from Swedish streams (Sodergren and others, 1972) The lack in finding any food chain effects has been attributed... argentatus, herring gull (immature) (c) 5. 53 25 62 13 Sterna albifrons, least tern (b) 6.40 17 68 15 Sterna hirundo, common tern (five abandoned eggs) 7.13 23 50 27 Larus argentatus, herring gull (d) 7 .53 19 70 11 9.60 22 71 7 13.8 15 64 21 18 .5 30 56 14 22.8 28 65 7 Phalacrocorax auritus, double-crested cormorant (immature) 26.4 12 75 13 Larus delawarensis, ring-billed gull (immature) 75. 5 15 71 14 Larus... trophic levels In a Long Island (New York) salt marsh, DDT residues in marine organisms increased with increasing organism size and increasing trophic level (Woodwell and others, 1967) Total DDT residues ranged over three orders of magnitude, from 40 µg/kg wet weight in plankton to 2,070 µg/kg in a carnivorous fish (the Atlantic needlefish, Strongylura marina) to 75, 500 µ g/kg in ring-billed gulls (Larus... benthic fish, and Weddell seals (Leptonychotes weddelli) in the Antarctic Ocean (Hidaka and others, 1983); kepone in the James River food chain (Connolly and Tonelli, 19 85) ; PCBs in the lake trout food chain in Lake Michigan (Thomann and Connolly, 1984); PCBs in the yellow perch (Perca flavescens) food chain in the Ottawa River (Norstrom and others, 1976); organochlorine compounds in micro- and macrozooplankton . 5. 4), organochlorine insecticides may persist in bed sediment and aquatic biota of forest streams, and forest soils containing organochlorine insecticide residues may be washed into the stream for many. applications to industry, commercial, and government buildings and land have decreased (Aspelin, 1997). The major pesticides used in and around homes and gardens in 1990 are listed in Table 3 .5. An examination. targeted, in sediment or aquatic biota, several are used in and around the home and garden (Table 3 .5) . These include chlorpyrifos, diazinon, carbaryl, permethrin, and 2,4-D. A number of local-scale