Endosulfan is a broad spectrum, non-systemic, contact and stomach acting insecticide and acaricide. It is sometimes seen as a replacement for other more persistent organochlorine insecticides (e.g. DDT, drins) but it is not widely used in agriculture.

The technical endosulfan product (typically 96% active ingredient (a.i.)) is a mixture of two isomers, known as a (or A or I) and b (or B or II), in the ratio of 70-80% a to 30-20% b.

In 1984, Dequinze et al (1984) estimated EC production capacity was 6,700 tonnes/year (t/y). World production was estimated at approximately 10,000 t/y (WHO 1984).

The main route of entry into the aquatic environment in the UK is from diffuse sources associated with its use as a pesticide, such as run-off from land and spray-drift.

A survey of 80 UK surface water sites involving 160 samples during the winter of 1988/89 (SAC 1989) revealed only two samples with positive, i.e. >10 ng/l, results for endosulfan: 11 ng l^-1 on the River Torridge and 14 ng l^-1 in the Forth estuary. In both samples, only a-endosulfan was detected. Both sites were sampled twice but only one positive result was obtained at each site.

Additional data on concentrations reported in the marine environment are presented in Appendix D.

Information summarised by Crane and Jones (1991) suggests that removal of endosulfan from the aqueous environment may occur by photolysis, hydrolysis, oxidation, volatilisation, biodegradation and sorption under certain conditions. However, the relative importance of the different processes is likely to be difficult to predict for a particular circumstance.

Sorption is an important fate for endosulfan in aquatic systems. Greve and Wit (1971) found that more than 75% of the endosulfan in the River Rhine was associated with particulate matter (mud or silt).

The degradation of endosulfan in marine microcosms was investigated by Cotham and Bidleman (1989). The results seemed to indicate virtually no degradation under the non-sterile conditions. Degradation in unsterilised sediment-water mixtures was also studied. Sediment was taken from a creek on the South Carolina coast where a number of fish kills had occurred. The half-lives of the two isomers were 22 days for the alpha and 8.3 days for the beta forms. The system was spiked by adding the pesticide to the overlying water and it was not until day 4 that the majority of the substance was found adsorbed in the sediment layer. The greater volatility of the a-isomer was demonstrated with most of the remaining a-endosulfan being found in the polyurethane plug used to seal the flasks by day 20 of the experiment. Endosulfan diol was the only metabolite identified in these studies.

An exhaustive literature review on the toxicity of endosulfan to marine organisms has not been carried out for the purposes of this profile. The information provided in this section is taken from existing review documents (Crane and Jones 1991). The most sensitive groups of organisms have been identified.

Crane and Jones (1991) summarised information on the aquatic toxicity of endosulfan and the more significant data are presented below.

Thursby et al (1985) conducted experiments to determine the effects of technical endosulfan on the growth and reproduction of the marine red macroalga (seaweed) Champia parvula. Growth of female and tetrasporophyte structures was significantly reduced after 14 days exposure to 47 µg l^-1 (lowest concentration tested) and 130 µg l^-1, respectively.

For molluscs, effects of endosulfan at concentrations of less than 100 µg l^-1 were only observed in the test on the inhibition of shell growth in the eastern oyster Crassostrea virginica reported by Butler (1964). However, this test was performed at 28 °C, which is above the recommended (US EPA 1982) temperature range for the species. At 19 °C, the EC50 for shell growth was six times larger at 380 µg l^-1. The effect was temporary with recovery periods in clean water of seven weeks at 19 °C and two weeks at 28 °C.

The toxicity of the a- and b-isomers of endosulfan to the common mussel Mytilus edulis was assessed by Roberts (1975) who measured the effects of the chemicals on the development of the byssal threads used by bivalves to anchor themselves. The b-isomer was found to be more toxic than the a-isomer with reductions in byssal thread attachment after 48 hours exposure to 200 µg l^-1 of 85% for b and 35% for a-. In experiments with an emulsifiable formulation of endosulfan, the toxicity was greater for smaller mussels and at higher temperatures.

Short-term LC50s for a variety of species of shrimp vary between 0.04 and 17 µg l^-1.

The lowest 96-hour LC50 for a crustacean species is 0.04 µg l^-1 for Pennies duorarum (Schimmel et al 1977). Crangon septemspinosa had an LC50 value of 0.2 µg l^-1 (McLeese and Metcalfe 1980), whereas the LC50s for all other crustaceans were greater than 0.4 µg l^-1. The lowest fish LC50s were 0.09 µg l^-1 for Leistomus xanthurus (Schimmel et al 1977) and 0.1 µg l^-1 for Morone saxatilis (Korn and Earnest 1974). The chronic values for the shrimp Mysidopsis bahia and the sheepshead minnow Cyprinodon variegatus are higher than the acute LC50s cited above.

The results of a six-laboratory ring-test on the toxicity of technical endosulfan to the polychaete worm Neanthes arenaceodentata were reported by Pesch and Hoffman (1983). The worms were exposed in flow-through systems and sand was provided as a sediment in which they could burrow. After exposure for 96 hours, 10 days and 28 days LC50 values were 195 µg l^-1, 158 µg l^-1 and 106 µg l^-1, respectively. Values for EC50s, based on the numbers of test animals which did not burrow, were almost identical to the corresponding LC50s.

In another experiment with a polychaete worm, McLeese et al (1982) investigated the toxicity of endosulfan to the ragworm Nereis virens with and without sediment in the test vessel. The preparation of the test solutions and dosed sediments was unusual in that solutions of the toxicant in a volatile solvent were evaporated in the test vessels before water or water and sediment were added. The exposure regime was semi-static, with aqueous test solutions being changed every 48 hours and sediment-water mixtures every 96 hours. The sediment, consisting of silt and clay (83%) and sand (17%), was 30 mm deep and covered with 15 mm of water. The 12-day LC50 values for worms exposed to seawater only and to seawater in the presence of sediment were 100 µg l^-1. Stressed worms in the test with sediment emerged from the sediment and subsequently did not burrow, even after the sediment was changed. The LC50 in the sediment-water experiment expressed in terms of the concentration of endosulfan in the sediment was 340 µg kg^-1.

The bioaccumulation of endosulfan has also been summarised in Crane and Jones (1991).

A maximum BCF of 22.5 was reported for mussels exposed to 100 µg l^-1 endosulfan for 70 days; the BCF decreased to 17 after 112 days (Roberts 1972). Exposure to concentrations of 500 and 1,000 µg l^-1 resulted in greater tissue levels but bioaccumulation factors (BCFs) of only 11 and 8.1 after 112 days. Ernst (1977) also tested mussels but used a-endosulfan in a mixture of pesticides and worked at much lower concentrations. Ernst reported a BCF of 600 for mussels at 10 °C in water initially containing 2.05 µg l^-1 a-endosulfan. The BCF was calculated using the steady-state water concentration of 0.14 µg l^-1 and tissue concentration of 84 µg kg^-1 wet weight obtained within 50 hours. The paper reports a half-life of 34 hours for a-endosulfan in mussels based on a one-compartment model. However, on considering the data presented, it appeared that more than 50% of the accumulated pesticide is lost after only 9 hours in clean water.

Haya and Burridge (1988) exposed the polychaete worm Nereis virens to solutions of endosulfan in aquaria containing seawater and sediments. The worms were exposed to concentrations of 60 µg l^-1 under hypoxic (12% saturated) and normoxic (presumably close to air-saturation) conditions at 7 °C for four days, with the test solutions being renewed after two days. After four days, the animals were transferred to clean water for the depuration phase. Uptake of endosulfan appeared to be linear under both hypoxic and normoxic conditions, although the bioaccumulation rate was nearly three times faster in the oxygen-deficient conditions. The maximum concentrations in the worms under hypoxic and normoxic conditions were about 4.4 and 1.7 mg/g lipid, respectively, and were recorded at the end of the exposure period and in neither case was equilibrium reached. The half-life for elimination of the endosulfan was approximately 60 hours.

During their investigation of the toxicity of endosulfan to two species of shrimp and three species of fish, Schimmel et al (1977) investigated the uptake of endosulfan by the test animals. In all cases where measurable (10 µg/kg wet tissue) residues were found after exposure for 96 hours to a technical mixture of a- and b-endosulfan, the predominant form in the tissue was endosulfan sulphate. The pink shrimp, although extremely sensitive to the acute toxic effects of endosulfan, does not appear to accumulate the chemical. Even when exposed to the highest test concentration of 0.089 µg l^-1, no residue was detected in the shrimp tissues. Bioaccumulation factors of 81 to 245 were calculated for the grass shrimp based on measured concentrations of 0.16 to 1.75 µg l^-1. The highest concentration in this test gave 65% mortality and residues of endosulfan of 78, 42 and 360 µg/kg (a, b and sulphate respectively). BCFs reported for pinfish Lagodon rhomboides, spot and striped mullet Mugil cephalus after exposure for 96 hours reached 1,299, 895 and 1,344, respectively.

In the same paper, the authors also reported that, during the course of a 28-day experiment, juvenile striped mullet exposed to an endosulfan concentration of 0.035 µg l^-1 reached a BCF of 1,000 after 96 hours and 2,755 after 28 days with tissue concentrations still increasing at the end of the test. The concentration of endosulfan sulphate in the fish was, at 80 µg/kg, nearly five times greater than that of b-endosulfan (17 µg/kg), whereas a-endosulfan was below the detection limit (10 µg/kg). The endosulfan was totally eliminated after only 48 hours in clean water. At the nominal concentration of 0.008 µg l^-1 in the water (which could not be measured accurately) no residues were detectable in the fish.

Potential effects include:

• toxic effects to algae and invertebrates (particularly crustaceans) at concentrations above the EQS of 0.003 µg l^-1 in the water column;
• sediment-dwelling organisms, especially crustaceans, may be at risk because the ultimate fate of endosulfan, its metabolites and degradation products is not known;
• identification as an endocrine disrupting substance.

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