Entry into the marine environment
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
Recorded levels in the marine
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.
Fate and behaviour in the marine
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
Effects on the marine environment
Toxicity to marine organisms
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,
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
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 on interest
features of European marine sites
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
- sediment-dwelling organisms, especially crustaceans,
may be at risk because the ultimate fate of endosulfan,
its metabolites and degradation products is not
- identification as an endocrine disrupting substance.