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ANNEX 1
PRIORITISING OF POP-CANDIDATES
Pentabromo diphenyl ether (PentaBDE, PeBDE)
CAS No. 32534-81-9 (commercial mixture)
PeBDE refers here to the commercial mixture of Pentabromo diphenyl ether that consists of
Polybrominated diphenyl ethers with three to eight bromine atoms per molecule. The penta
brominated molecules dominate in the mixture and contribute with more than 50% to the total amount
whereas tetra- and hexa brominated molecules contribute with 24-28 % and 4-12 % respectively. The
summary below is mainly based upon a Nordic report (1).
PeBDE is a brominated flame retardant mainly used in different polyurethane (PUR) applications as
for furniture and upholstery in automotive industry and domestic furnishing. Other possible minor uses
are in rigid polyurethane elastomers (e.g., in instrument casings), in epoxy resins and phenol resins
(electric and electronic appliances). The global consumption of PeBDE in 1999 has been estimated to
8 500 tonnes of which only 210 tonnes was used in Europe.
Persistence
According to a standard, OECD 301B ready biodegradability test, with aerobic activated sludge
sewage treatment plant organisms, PeBDE is not readily biodegradable. Nevertheless, according to the
results from a study with decabromo diphenyl ether, photolysis resulting in reductive debromination
might be a possible pathway for abiotic degradation. The total (biotic and abiotic) half-lives of one
tetra- and one penta brominated diphenyl ether in aerobic sediment has been estimated to 600 days and
150 days in water and soil. Findings of PeBDE related substances in remote regions also indicate high
persistency. Thus, the criterion for persistency, seems to be met.
Bioaccumulation
All components of PeBDE as well as of commercial PeBDE have a log Kow greater than 5, suggesting
that they have potential to bio-accumulate. The bio-concentration factor (BCF) for commercial PeBDE
in carp was estimated to 27 400, which is well above the criterion limit. In fish, BDE-47 is taken up
more efficiently than CB-153, the PCB congener with the highest concentrations in biota. But both
BDE-99 uptake and BDE-153 uptake in fish seems to be similar to those of other PCBs studied (-31, 52, -77 and -118). In mammals, the main components of PeBDE are taken up efficiently and excreted
slowly by both rats and mice. The excretion in mammals is mainly faecal and uptake efficiency and
elimination time correlates negatively with the degree of bromination. The presence of PeBDE in biota
as fish, birds (guillemot and peregrine falcon) and human food as pork, lamb and beef (2) also support
bioaccumulation. There are also findings indicating that PeBDE might biomagnify. The criterion for
bio-accumulation is fulfilled.
Toxicity
In vitro studies of PeBDE have shown i.a., an ability to activate the Ah-receptor and possible
genotoxicity (intragenic recombination). Immunotoxicity for the major PeBDE congeners has been
shown in mice but not in rats. In vitro, several PeBDE congeners have been shown give rise to
antiestrogenic response.
In vivo, studies of rats indicate that liver is the main target organ affected by PeBDE with a NOAEL of
1 mg/kg/d. Other studies have revealed i.a. developmental neurotoxicity in mice after a single dose of
0.8 mg/kg BDE-99 to 10 days old mouse pups.
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Studies of algae and invertebrates indicate effects at exposure levels as low as 3-6 µg/l and larval
development for a copepod Acartia tonsa showed disturbances after 5 days at 13 µg/l. The criterion
for adverse effects is met.
Potential for long-range transport
PeBDE components have very low volatility. Vapour pressure between 9.6 x 10-8 - 4.7 x 10-5 Pa have
been reported and the corresponding water solubility between 2 – 13 µg/l. The estimated Henry’s Law
Constants (3) nevertheless suggest that at least the lower brominated components can also be
volatilised in significant amounts from aqueous solutions. Vapour pressure and water solubility
decreases with increasing bromination.
PeBDE congeners have been found in Arctic air at remote sites in Canada and Russia. Total
concentrations were <1-28 pg/m3. At another remote Arctic area in Pallas, Finland, BDE-47 and BDE99 concentrations were measured between 0.3-2 pg/m3. The same congeners were also observed at two
Swedish sites, Ammarnäs and Hoburgen, remote from point sources. The sumPBDE concentration in
the air varied in this study generally between 1 and 10 pg/m3. Also according to the atmospheric halflife estimates from SAR modelling, PeBDE has long-range transport potential in the atmosphere as the
half-life I estimated to 10-20 days, which is well above criterion limit at 2 days.
The group “Polybrominated diphenyl ethers” is listed among priority substances under the EU water
framework directive (4, 5). In this directive a mean concentration of PeBDE congeners in European
sediment samples at 1.25 µg/kg, based on 16 samples (all positive) from 16 sites.
Consequences
As the European market at this very day has adjusted its assortment of flame retardants to a product
mix where PeBDE constitutes less than 1% , it is not likely that a regulation of this substance would
result in any major conflicts. A ban within the EU is under way (6). As many countries have different
standards for flame protection for different materials this might cause problems to regulate the current
substances in e.g. the US, as their present use of PeBDE is 15% of the total consumption and still
increasing.
References
1. Pentabromodiphenyl ether as a global POP. Tema Nord 2000:XX. Document can be found at:
http://www.norden.org/miljoe/uk/PeBDEfinal.pdf.
2. Darnerud et al (2000) Organiska miljökontaminanter i Svenska livsmedel, Sakrapport till
naturvårdsverkets miljöövervakningsnämnd. Report and data can be found at:
http://www.imm.ki.se/national/.
3. COM, 2000. Risk Assessment of Diphenyl Ether, Pentabromoderivative (Pentabromodiphenyl
Ether). CAS Number: 32534-81-9, EINECS Number: 251-084-2. Final Report of August 2000,
Commission of the European Communities. Rapporteur: United Kingdom.
4. EU parliament and council. Water framework directive 2001/C 154 E/11, 29.5.2001 and Directive
2000/60/EG.
5. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
6. Commission Recommendation 2001/194/EC.
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Polychlorinated naphtalenes (PCN)
CAS No. 70776-03-3 (a mixture)
PCN was introduced on the market already in the 1920’s. The technical PCN is a mixture of congeners
containing 1-8 chlorines per molecule. It has similar use areas as the more well-known PCBs, and they
include cable insulation, wood preservation, engine oil additive, in electroplating, in dye production, in
capacitors and in oils for refraction index measurements. Because of many cases of poisoning,
especially in cattle, the use of PCN decreased in the 1970’s. Today, there is no known use in the
industrialised world (1). Considering that the production is simple, one can’t rule out continuing
manufacture in developing countries. Another important source of PCN, is the unintentional formation
of PCN when chlorine-containing material is combusted under poor conditions (a high formation of
PCN was observed when fly ash was heated at 300oC). The relative contribution from waste
incineration, in relation to active use of PCN, seems to increase with time.
Persistence
There are few reliable degradation studies on PCN. The available data indicate that the persistence
increases with increasing degree of chlorination, and that PCNs with more than one chlorine would
fulfil the persistence criterion (2). The ubiquitous presence of PCN in biota even today, and then
mainly in mammals including humans (3), supports a very high persistence of highly chlorinated PCN.
Bioaccumulation
For PCNs containing 2-5 chlorines, there are bioconcentration studies in fish showing a
bioconcentration well above the criterion (BCF>5000) (1). The highest values (33 000) are reported
for tetraCN, but the congeners with the highest potential (penta and hexaCN), as indicated from their
presence in predatory animals, are not studied.
Toxicity
PCN is very toxic to most organisms, probably because of the structural resemblance with the
chlorinated dioxines. In fish and crustaceans, LC50 values of 0.4-2.8 mg/l have been reported for
mono-and dichlorinated PCNs. LC50 values of 0.008-0.44 mg/l have been reported for highly
chlorinated technical mixtures in two fish species (1).
There are many cases of exposure of cattle to PCN (2), which indicate a high toxicity in mammals.
Systemic effects were observed at exposure to 1 mg PCN/kg body weight/day. Effects were first
observed on the skin, followed by anemia, liver damage, and finally even death (1, 2). One reason for
the toxicity seems to be an imbalance in the vitamine A homeostasis, making it possible to classify
PCN as a potential endocrine disruptor.
Potential for long-range transport
Photolysis experiments have shown a halflife of 2.7 days in air for dichlorinated PCN (1). QSARmodels indicate a halflife of 8-437 days for highly chlorinated PCNs (1). The criterion for long-range
transport thus seems to be fulfilled, which is also supported by the presence of PCN in arctic air ( 49
pg/m3) (4). It should be noted that the air concentration of PCN is just a few times lower than the
concentration of PCB in some air samples from the Arctic (4) and the UK (5).
Consequences
OSPAR has prioritised PCN as of ”very high concern”, because of POP-like characteristics, but there
is no known use in the EU. A decreasing global use of PCN is indicated by decreasing concentrations
in the environment. A global regulation should in the first place be directed at making sure that any
potential manufacturing in developing countries is stopped, as there are many alternative chemicals.
Another purpose would be to encourage and support environmentally sound disposal of products
containing PCN, as presently for the PCBs. Such disposal is presently required for PCB in the POPconventions, but even if no new plants would be required, controlled destruction is costly. The biggest
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problem in fulfilling such a requirement for PCN-containing products still appears to be how to
identify products that contain PCN. Those measures that are required by the conventions to reduce the
unintentional formation of dioxines/furanes in combustion processes, may be sufficient to also reduce
unintentional formation of PCN, and further measures may not be needed for that particular source.
References
1. Risk profile polychlorinated naphtalenes, Preliminary risk profile prepared for Ministry of
Housing, Physical Planning and the Environment (VROM, the Netherlands) in the framework of
the project Risk Profiles III, March 2002.
2. Environmental hazard assessment: Halogenated naphtalenes. Toxic Substances Division,
Directorate for Air, Climate and Toxic Substances, Department of the Environment, UK, 1993
3. Norén, K. and D. Meironyté, Certain organochlorine and organobromine contaminants in Swedish
human milk in perspective of past 20-30 years. Chemosphere 2000, 40, 1111-1123.
4. Harner et al, Polychlorinated naphtalenes and coplanar polychlorinated biphenyls in arctic air,
Environ. Sci. Technol. 1998, 32, 3257-3265.
5. Harner et al, Polychlorinated naphtalenes in the atmosphere of the United Kingdom, Environ. Sci.
Technol. 2000, 34, 3137-3142.
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Hexachlorocyclohexane (HCH, including -HCH, lindane)
Technical HCH is a mixture of different isomeric forms (-, -, -, -, -) where the hydrogen and
chlorine atoms have different spatial orientation on the carbon atoms of the hexane ring. Lindane
contains >99% -HCH.
CAS No.: 608-73-1 (HCH), 319-84-6 (-HCH), 319-85-7 (-HCH), 319-86-8 (-HCH), 58-89-9 (HCH, lindane).
This short summary is mainly based on data compiled in two WHO documents (1,2) and on a draft
produced within the context of the OSPAR Convention (3). In addition, information has been taken
from reports on monitoring and from reports generated within the EU review of active ingredients in
plant protection products (4). It should be noted that a background document on lindane is in
preparation for UN ECE in the context of the Convention on Long Range Transboundary Air Pollution
(LRTAP) (3).
Only -HCH shows a significant insecticidal activity. Purification of lindane from HCH produces the
remaining isomers (mainly - and -) which are used as intermediates in the production of
trichlorobenzene, hydrochloric acid and other chemicals (1). Isomerization of lindane does not seem to
occur in the environment, whereas slow isomerization of -HCH occurs (2).
HCH has been produced commercially since 1949 (1). Until the end of the 1970's isomeric mixtures of
-, - and -HCH were commonly used as insecticides in agriculture and as wood preservatives,
however, in most countries where HCH is still used, the use is restricted to -HCH. EU Member States
put an end to the use of technical HCH in 1979 by Directive 79/117/EEC. It appears that technical
HCH is still used in some Eastern European countries (3). Still in 1986-87, approximately 27 000
tonnes of technical HCH was used in India (1). In 1970, the usage of -HCH was estimated to be
25 000 on an European basis, while the usage in 1996 was only 366 tonnes (3). The major part of the
remaining use of -HCH in 1996 has (however with uncertainty) been attributed to use in Eastern
Europe (3). A similar decrease has been observed for -HCH.
Lindane has a wide use in agriculture and forestry (for seed treatment and soil application), in
household biocidal products (e.g. treatment of animals, ornamentals and turfs), as wood and textile
preservative, and also in medical control of ectoparasites on humans and animals (3). The world
production of lindane was estimated to approximately 38 000 tonnes in 1986 (3); but to only about
3 200 tonnes per year during the period 1990-1995 (3). Within the OSPAR Convention area the nonagricultural use has been judged as insignificant (3). However, according to a questionnaire in 1997,
some non-agricultural use of lindane was important in the United Kingdom and in Belgium (3).
Persistence
The results from screening tests on biodegradability of lindane are highly variable, but it is notable
that lack of transformation was demonstrated in a few tests even after adaptation of the
microorganisms (3). Reported half-lives for lindane in aquatic systems also vary considerably, from a
few days to >700 days (3). From laboratory studies in soil, half-lives of 260-708 days (3) and 133-980
days (4) have been reported. From field studies maximum half-lives of >300 days have been reported
(3). In the report produced under EU's Plant Protection Directive (91/414/EEC), it is concluded that
DT90 after incorporation into soil (commonly used within the EU) in most cases is <1 year, but that the
persistency can be significantly prolonged under certain conditions (4). The first dehydrochlorination
step to pentachlorocyclohexene is considered as a rate limiting step for degradation. Metabolites has
been found in very low amounts in laboratory studies (4). Data suggests that degradation of lindane is
accelerated under anaerobic conditions (4).
Since substantial volatilization of lindane is expected, many of the reported half-lives are considered to
be very uncertain estimates. By use of a level III multimedia model, the overall persistence of lindane
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in source areas or in remote areas (e.g. assuming 100% emission to air, or 20% emission to air, 80% to
soil etc.) was calculated to 613- 873 days (3).
For -HCH data are conflicting: this isomer is reported to be less persistent than lindane (3), or it is
reported to more slowly biodegraded than lindane (2). -HCH seems to be the most persistent isomer
(2).
The expected persistence is confirmed by identification of HCH (-, - and -) in air, rain water,
plants, and in aquatic and terrestrial animals, including man. Therefore, the criterion for persistency is
considered to be met for HCH, despite the considerable variation in reported half-lives.
Bioaccumulation
Reported values for log Kow for lindane varies between 3.2 and 3.7. Bioconcentration factors (BCF)
for lindane in a study on four different fish species were 12 800-15 400, however, usually the reported
BCFs are a factor of 10 lower than these values (3). In the EU report on lindane (4), a BCF of 1 300
was reported, with 15% of the radioactivity remaining in the fish after 14 days in clean water. The
report concludes that the detection of residues in wild birds and mammals indicates that organisms
consuming fish are at risk.
- as well as -HCH have log Kow values of 3.8 (2). BCF for -HCH in fish varies from 313 to 1 216
(2), however, based on measured amounts in muscle and fat in bream collected in the River Elbe, BCF
has been calculated to 10 000-50 000 (2). For -HCH, BCF in fish of 250-1 500 was reported however
it has also been concluded that the bioconcentration is higher and the elimination is slower for -HCH
than for the other isomers (2).
Recent data from the Swedish Museum of Natural History indicates that the concentrations of HCHs
in biota from the Baltic Sea as well as from the Swedish west coast are decreasing by a rate of 10% or
more per year since the end of the 1980's (6). -HCH is in general decreasing faster than lindane. The
ratio lindane/-HCH was found to be higher in fish from the Kattegatt compared to the Baltic. This
could reflect that in the former east-bloc countries technical HCH has been used whereas the use of
lindane has been more common in western countries. The concentrations of lindane (-HCH) varies
from 5 to 30 µg/kg lipid in fish and mussel and the - and -isomers are detected at similar
concentrations.
Also within the OSPAR Convention area concentrations of lindane in fish and mussel tissues has
generally been decreasing during the period 1990 to 1995, especially in relatively polluted regions (3).
In contrast, a significant upward trend was observed in dab muscle from southern Norway (3). In biota
in the Arctic Ocean the following concentrations of HCH were reported for the first half of the 1990's:
mussels <0.5-0.82 µg/kg ww; fish liver <0.6-153 µg/kg ww; seabird liver 0.2-25 µg/kg ww; marine
mammals blubber 17-473 µg/kg ww; marine mammals fat n.d.-1 150 µg/kg ww (3).
From end of 1980's to mid-1990's, mean concentrations of approximately 160-600 µg HCH/kg lipid in
Arctic polar bears were reported (9).
Lindane is also found in animal food items; the substance was found in 10-50% of meat samples and
in more than 50% of fish, crab and mollusc samples in Germany in the period 1995 to 1998 (3). Other
HCH isomers were found less often. HCHs are also found in human breast milk. As an example, the
mean concentration of lindane in more than 7 000 samples of breast milk samples collected in
Germany between 1969-1984 was 0.01-0.11 mg/kg on a fat basis (1). A slow decrease was observed
during the last years. From Swedish measurements (10) of -HCH, concentrations of 5.2-127 µg/kg
fat have been reported (mean concentration 14.8 µg/kg fat, n=31). Lindane has been found in serum
samples from Swedish women at <2-13.4 µg/kg serum fat; -HCH at <2-7.4 µg/kg (10).
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In the OSPAR report (3) it is suggested that also a not so lipophilic substance like lindane may be a
candidate for biomagnification. It has been suggested that not only the degree of lipophilicity but also
the degree and position of chlorination and particularly the elimination pathways determine the
potential for biomagnification (3). Some data (such as measured levels in fisheating tuna fish and
dolphins in relation to concentrations in fish, and measured levels in eggs of pelicans in relation to
concentration in pelican's food items) suggests a potential for biomagnification (3). In contrast, others
have shown that even though HCH occurs in Arctic air, snow and sea water and is efficiently
accumulated by species at low trophic levels, the biomagnification potential is low at the upper end of
the food web, with HCHs only in 10% or less of the samples (3).
As an overall conclusion, the frequency at which the HCH isomers are being detected in biota
indicates that the criterion for bioaccumulation is met, despite the fact that some of the reported BCF
values are relatively low. It is worth noting that the measured concentrations in various environmental
compartments show a decreasing trend.
Toxicity
LC/EC50 values in short-term studies on toxicity studies of lindane to fish and daphnia are 0.022-0.063
mg/l and 1.6-2.6 mg/l, respectively (4). Reported NOECs following long-term exposure are 0.00290.054 mg/l (4).
The lowest NOAEL determined in standard toxicity tests on mammals for lindane is 0.47 mg/kg bw/d
based on effects on the liver (4). A range of different effects caused by endocrine disruption, have
been indicated in studies on different mammalian species. One example is reduced ovulation rate seen
in rabbits at a dose of 0.8 mg/kg bw/d (4). In the rat, effects related to hormonal disruption as well as
increased foetal mortality occurred at 0.5 mg/kg bw/d (3). The International Agency for Research on
Cancer (IARC) has concluded that lindane is a "possible human carcinogen" (Class 2B), however the
substance is not considered to pose a mutagenic risk (3). Within the EU, lindane has not been
classified in relation to criteria for carcinogenicity, reproductive toxicity or mutagenicity. Based on
presence in mother's milk the substance is classified with R 64 ("May cause harm to breastfed
babies"). The criterion for adverse effects is considered to be met for lindane.
The other isomers are less toxic than lindane; LC/EC50 for - and -HCH in fish and aquatic
invertebrates are of the order of 1 mg/l (2). In long-term study on daphnids however NOEL was as low
as 0.05 mg/l for -HCH; from a long-term study on -HCH in fish, NOEL was 0.03 mg/l (2). -HCH
has been shown to cause a clear increase in the activity of liver enzymes at 5 mg/kg diet, equivalent to
0.25 mg/kg bw. A weak estrogenic effect of -HCH has been described (2). The criterion for adverse
effects is considered to be met also for these isomers.
Potential for long-range transport
The vapour pressure of lindane is 1.2-7.4 x 10-3 Pa (20-25C) (3). Henry's Law Constant has been
calculated to the order of 1 x 10-1 Pa x m3 x mol-1. Estimated half-life in air is 4.6 days (4) or 43 days
(3), however, there is also other data available which indicates that the substance is persistent in air
with estimated half-lives of >11 000 days (4). Monitoring data seem to support that the substance is
relatively persistent in air. Laboratory as well as field studies indicate a substantial distribution to air;
90% loss from soil surfaces was observed within 24 hours, while loss from plant surfaces was even
faster, 86% within 6 hours (4). Soil incorporation reduces the distribution to air; 13% loss was
observed within 24 hours at the laboratory, while field studies have indicated presence of lindane
above background levels still 6 months after soil incorporation (4).
Vapour pressure of -HCH is 2.7 Pa (20) (2) and the calculated Henry's Law Constant is 390 Pa x m3
x mol-1. This isomer is therefore considerably more volatile than lindane. Monitoring data supports the
high potential for volatilization. -HCH has a slightly lower potential for volatility, 0.67 Pa at 20 (2).
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The estimated Characteristic Travel Distance (CTD) describes the long-range transport potential as the
distance at which the initial concentration drops to 37% (1/e). For the standard scenario (100%
emission to air), the CTD for lindane is 7 400 km (3).
The atmospheric background concentration of lindane has been reported to be in the range of 0.0150.3 ng/m3 (3). A relationship has been found between global use of technical HCH and air
concentrations of -HCH in the Arctic air between 1979 and 1994 (3). Elevated levels of persistent
organic pollutants (including HCH) are positively correlated with long-range transport episodes from
use areas in the mid-latitudes of North America, Europe and Asia (3). Such correlation between levels
of PCBs and HCHs and episodes with air masses originating mainly in Europe has been reported from
Svalbard (3). Measurements in Sweden (Rörvik station) show 0.010-0.050 ng/m3 of each one of the and -isomers, and a deposition of 0.2-0.7 ng -HCH/m2 per day, 0.2-0.9 ng lindan/m2 per day (10).
In Swedish measurements in rain water (three stations in Skåne, Uppland and Lappland) during 199092, - and -HCH were detected in nearly 100% of the samples (5). Highest concentrations of lindane
were detected in the south of Sweden (median concentration 13 ng/l). In contrast, the levels of -HCH
showed little variation between stations and season. The ratio /-HCH therefore increases from the
south to the north (5). The results indicate that atmospheric transport of lindane occurs despite the fact
that European usage mainly has been limited to treatment of seeds which are incorporated in soil.
For -HCH a long time trend in muscle of pike from lake Storvindeln, near the Arctic circle in
Sweden, has been presented. As there is little agricultural activity in this remote area, atmospheric
transport and deposition is expected to be the only significant source of -HCH (3).
The emission of lindane to the atmosphere from Europe (38 countries) was estimated to 1 310 tonnes
per year in 1990, to 765 tonnes/year in 1997 (3). From the 15 Contracting Parties of the OSPAR
Convention, the emission was estimated to 417 tonnes/year in 1990, and to 733 tonnes/year in 1997
(3).
Lindane is also being transported from the application areas via water. A total of 800-940 kg was
estimated to reach the North Sea in 1998, mainly by riverine input but also from direct discharges (3).
HCH is included in The list of priority substances in the field of water policy, established under
Directive 2000/60/EC of the European Parliament and of the Council, establishing a framework for
Community action in the field of water policy (7). In the context of the Water Framework Directive,
the following mean concentrations have been reported for European surface waters (8):
0.017 µg/l
(11 666 samples from 546 stations; 8 260 above determination limit)
-HCH:
0.0094
µg/l
(1 974 samples from 77 stations; 1 190 above determination limit)
-HCH:
0.013 µg/l
(1 226 samples from 44 stations; 751 above determination limit)
-HCH:
0.0092 µg/l
(208 samples from 18 stations; 106 above determination limit)
-HCH:
For the sediment phase, the following mean concentrations were reported (8):
9.15 µg/kg
(953 samples from 53 stations; 689 above determination limit)
-HCH:
19.4 µg/kg
(594 samples from 27 stations; 398 above determination limit)
-HCH:
42.3 µg/kg
(822 samples from 27 stations; 528 above determination limit)
-HCH:
Lindane has also been measured in groundwater samples in Germany, in a few cases as > 0.1 µg/l (3).
The global marine background value for lindane has been reported to about 0.6 ng/l, with values
ranging from 0.016 ng/l in lower reaches of the Atlantic to a maximum of 4.4 ng/l off the coast of
Iceland in the Arctic (3). Slightly more recent (1990-92), values are 0.021-0.075 ng/l (3). Since sea
water acts as a source of atmospheric lindane, it has been estimated that the accumulated levels in the
sea can support the air concentration levels during 10-15 years (3).
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The global transport and distribution of HCHs exemplifies the "cold condensation" effect with
volatilization soon after application, especially in the tropics, and partitioning from the air in colder
environments, such as the Arctic (3,9).
According to calculations of the overall HCH budget for the Arctic (9), the Arctic Ocean is in steady
state with respect to lindane (input being approximately equal to output), while for -HCH there is a
net export, mainly via ocean current advection and to a lesser extent by atmospheric processes. For
HCH (- and -) the net output from the Arctic Ocean was calculated to about 200 tonnes/year in the
early 1990's. By comparison, the net loading was calculated to about 80 tonnes/year in the 1980's, 80%
of which was gas-phase deposition (9).
Due to the widespread occurrence, it is concluded that all HCH isomers meet the criterion for longrange transport.
Consequences
Lindane is included in the list of "Substances Scheduled for Restrictions in Use", set up by the UN EC
LRTAP. However, the importance of technical HCH as insecticide in general and lindane in particular
has been declining in Europe (3). While in a period from 1970 to 1979 HCH made up around 13% of
the insecticides used in Europe, it was reduced to <5% in the period 1991-96 (3). Furthermore, the use
of lindane has decreased markedly since the middle of the 1990's (3). However, there is a large
regional variance depending on the latitudes and the crops grown, e.g. maize, rape, rice and cotton
require more insecticides than cereals and soya (3). In accordance with Commission Decision
2000/801/EC, plant protection products containing lindane can no longer be authorized within the EU.
All use of lindane in plants protection products within the EU must have ceased by June 2002. Outside
western Europe, inclusion in the Stockholm Convention may have negative effect on pest control.
However, the large decline in world production of lindane (see above) indicates that there is a
decreasing need for this particular substance.
In the context of the OSPAR Convention, some alternatives to lindane have been listed: e.g.
azaconazole, boric acid, chromic aid, copper/chrome/arsenic and sodium fluoride for wood
preservative use; organophosphates and pyrethroids for insecticidal use in agriculture and for medical
use (human and veterinary) and; acute poisons and anticoagulants for rodenticidal use (3).
Should the use of HCH be stopped, identification of HCH in different environmental compartments
would still be expected for a long time due to the accumulated concentrations in sea water and in the
Arctic.
References
1. WHO-IPCS (1991) Environmental health Criteria 124 Lindane. Geneva, World Health
Organization.
2. WHO-IPCS (1992) Environmental health Criteria 123 Alpha- and beta-hexachlorocyclohexanes.
Geneva, World Health Organization.
3. Final Draft OSPAR Background Document on Hazardous Substances Identified for Priority
Action - Lindane (-HCH) - Presented by Germany. OSPAR 02/7/10-E. To Meeting of the
OSPAR Commission, Amsterdam, 24-28 June 2002. OSPAR Convention for the Protection of the
Marine Environment of the North-East Atlantic.
4. - European Commission Peer Review Programme. Draft Assessment Report prepared in the
context of the possible inclusion of the following active substance in Annex I of Council Directive
91/414/EEC: Lindane. Volumes 3. Rapporteur Member State: Austria.
- European Commission Co-Operation. Concise Outline Report of ECCO meeting 85: Lindane.
5. Nordiska Ministerrådet (1994) Pesticides in precipitation and surface water. TemaNord 1995:558.
Copenhagen, Nordic Council of Ministers.
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6. Swedish Museum of Natural History (2000) Comments concerning the National Swedish
contaminant monitoring programme in marine biota. Compiled by Bignert A, Contaminant
Research Group at the Swedish Museum of Natural History. Stockholm, 2000-04-25.
(http://www.nrm.se/mg/mcom.pdf).
7. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001
establishing the list of priority substances in the field of water policy and amending directive
2000/60/EC. Official Journal of the European Communities L331/1, 15.12.2001.
8. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
9. de March BGE, de Wit CA and Muir DCG (1998) Persistent Organic Pollutants. Chapter 6 in:
AMAP (1998) AMAP Assessment Report: Arctic Pollution Issues. Oslo, Arctic Monitoring and
Assessment Programme.
10. Personal communication. Britta Hedlund, Swedish Environmental Protection Agency.
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Perfluorooctane Sulfonates (PFOS)
C8F17SO3
PFOS belongs to a group of organic compounds which have in common that they contain fluorine
atoms, i.e. all or some of the hydrogen atoms, bounded to the carbon chain, have been substituted with
fluorine atoms. To the end of the carbon chain there is a reactive sulfonate group which could be
associated with metal ions and other positive charged substances All other positions on the carbon
chain are occupied by fluorine atoms. Perfluorooctane sulfonate ion itself has no Cas No.. In the table
below there is a list of some selected perfluorooctane sulfonate compounds and their derivatives.
All of them dissociate in water at neutral pH. A major part of the PFOS produced is incorporated into
polymeric chains. PFOS has been used mainly as a surface treatment agent for carpets, fabrics,
furniture, paper and leather.
PFOS derivatives and salts
Acid
Ammonium salt
Diethanolamine salt
Lithium salt
Potassium salt
CAS No.
1763-23-1
29081-56-9
70225-14-8
29457-72-5
2795-39-3
Persistence
PFOS does not hydrolyse, photolyse or biodegrade under environmental conditions. This stability is
typical for perflourinated compounds and the reason for that is that the carbon chain, which is the
backbone of the molecule, is totally surrounded by fluorine atoms with strong bindings to all positions
of the carbon chain. These C – F bindings are probably the strongest bindings in nature. PFOS is
therefore assessed to be very persistent in the environment, and to fulfil the criterion for persistence.
The occurrences of PFOS in the environment confirm this statement through findings in blood plasma
(up to 1 047 ppb) in birds (eagles, albatross) and fish. PFOS has also been detected in polar bears and
seals in the arctic, dolphins in the Mediterranean Sea and Ganges and turtles in Mississippi.
Bioaccumulation
Due to the fact that PFOS is an extreme surface-active agent, it is impossible to measure log Kow
since the test ends up with three phases. Reported Bioconcentration Factor (BCF) in fish is 980, but it
is possible that the physical/chemical properties of PFOS may influence the validity of this test.
Experiment on rats indicates that PFOS do not accumulate in adipose tissue but on blood proteins.
PFOS is well-absorbed following ingestion. The half-life for elimination from plasma after single oral
dose in male rats was 7.5 days. A study on humans which were exposed over a long time, retirees who
worked in a factory producing PFOS, shows a very long half-life for elimination: between 1 and 4
years. The occurrences of PFOS in birds of prey, fish and polar bears are also an indication on
accumulation in these organisms and this bioaccumulation is not connected to adipose tissue.
Toxicity
During chronic toxicity studies on fish (Pimephales promelas) adverse effects has been observed such
as growth inhibition and decreased survival at 0.3 mg/L of the potassium salt of PFOS after 42 days.
At almost the same level (0.25 mg/L) there are also growth inhibition and impact on the reproduction
to the saltwater species Mysidopsis bahia (Mysid shrimp) at a 35-day test with the potassium salt of
PFOS.
Postnatal deaths and other developmental effects were reported at low doses in offspring in a 2generation reproductive toxicity study in rats. The NOAEL and LOAEL for the second-generation
offspring (F2 pups) were 0.1 mg/kg/day and 0.4 mg/kg/day, respectively, based on reduction in pup
body weight. In a 6 month study of cynomolgus monkeys, deaths were observed at doses as low as
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0.75 mg/kg/day. Thus, the criterion for adverse effects is considered to be met.
Potential for long-range transport
PFOS is not a volatile compound and PFOS is likely to resist in the atmosphere due to the persistency
of PFOS. Findings in polar bears and seals in the Arctic indicate long-range transport of PFOS.
Consequences
The largest producer of PFOS, 3M, is going to phase out PFOS due to voluntarily agreements. This is
a result of negotiations between 3M and USEPA. Obviously there are alternative to PFOS and these
alternatives are soon going to be put on the market. USEPA has during spring 2002 declared new
regulations which regulates new uses of PFOS. Within the EU, a risk management program is
currently initiated in the EU, with UK as a lead country.
References
1. Renner, R.,(2001) Environmental Science and Technology, vol 35, p. 155 – 160.
2. Sulfonated Perfluorochemicals in the Environment, Sources, Dispersion, Fate and Effects, 3M,
(2000)
3. Draft OECD report
4. Federal Register, USEPA, Monday, March 11, 2002, part 3, 40 CFR part 721, Perfluoroalkyl
Sulfonates; Significant New use Rule; Final Rule and Supplemental Proposed Rule.
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Short-chained chlorinated paraffins (Alkanes, C10-13, chloro, SCCP)
CAS No 85535-84-8
Short-chained chlorinated paraffins (SCCP) is the group of chlorinated paraffins (short-, medium- and
long-chained) which exhibit properties, that qualify the group as POP. Two producer exist within EU,
according to the EU RAR under the Council Regulation (EEC) No 793/93 of 23 March 1993 on the
evaluation and control of the risks of existing substances. The main uses are in metal working fluids,
as plasticiser in paints, coatings and sealants, as flame retardant in rubbers and textiles, and in leather
processing (fat liquoring) (1).
Persistence
SCCP is recognized as very persistent. SCCP does not hydrolyse in water and is neither readily
biodegradable nor inherently biodegradable (1). SCCP is expected to fulfil the criterion for
persistence.
Bioaccumulation
The log Kow for SCCP is 4.4-8.7. High bioaccumulation factors between 1000 and up to 50 000 have
been noted i fresh- and seawater organisms. In “pure water”, the elimination half time is between 9
and 20 days (1). SCCP is expected to fulfil the criterion for bioaccumulation.
Toxicity
SCCP exhibits a high toxicity towards aquatic organisms: NOEC for fish is in the order of 0,04 mg/l,
for Daphnia magna 0.005 mg/l and for algae 0.012 mg/l. Reproduction test with mallard exhibits a
NOAEL of 166 mg/kg. SCCP is further classified as to IARC group 3 (not classifiable as to human
carcinogenicity) (1). SCCP is expected to fulfil the criterion for toxicity.
Potential for long-range transport
Emissions of SCCP to the atmosphere are likely to be very low. Estimated levels exhibit a small but
measurable volatility. Vapour pressure at 40°C is 0.0123 Pa. The half life time in air has been
estimated to 1.9-7.2 days. Recently performed investigations exhibit high levels of SCCP in biota from
the Artic region, up to 1.4 mg/kg blubber in white whales (Beluga). This may indicate that these
substances are transported over long distances (1). Some uncertainties still remain if SCCP can fulfil
the criterion for long-range transportation.
Consequences
In 1994 ca. 15 000 tonnes of SCCP were produced within EU. Different voluntary actions within EU
resulted in a drastically reduced use, and EuroChlor estimated the use for 1998 to ca. 4 000 tonnes (1).
In the context of a possible long-range transportation, SCCP is discussed as a possible candidate for
inclusion in the LRTAP POP-protocol. On 25 June 2002 the EC agreed on Directive 2002/45/EC, as
20th amendment to Directive 76/769/EC, banning the use of SCCPs in metal working fluids and leather
fattening liquors (2). SCCP is further selected for priority action within OSPAR (3), and classified as a
priority hazardous substance within the Water Framework Directive (2000/60/EC) (4).
The consequence of a regulation within EU is expected to be moderate (5). Medium- (MCCP) and
long- (LCCP) chained chlorinated paraffins are used as substitutes. These substances are, however,
presently investigated according to their human and environmental effects. Since SCCP also is
produced and used globally, the total global consequence is difficult to judge.
References
1. OJ of the European Communities, Commission Recommendation of 12 October 1999 for:
Alkanes, C10-13, chloro, CAS#: 85535-84-8, EINECS#: 287-476-5.
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2. Directive 2002/45/EC of the European Parliament and of the Council of 25 June 2002 amending
for the twentieth time Council Directive 76/769/EEC relating to restrictions on the marketing and
use of certain dangerous substances and preparations (short-chain chlorinated paraffins)
3. OSPAR Background Document on Short Chain Chlorinated Paraffins, OSPAR 01/4/8, 2001
4. DECISION No 2455/2001/EC OF THE EUROPEAN PARLIAMENT AND OF THE COUNCIL
of 20 November 2001 establishing the list of priority substances in the field of water policy and
amending Directive 2000/60/EC.
5. Socio-Economic Impacts of the Identification of Priority Hazardous Substances under the Water
Framework Directive, Final Report prepared for European Commission Directorate-General
Environment, RPA, December 2000.
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Chlordecone (Kepone)
1,1a,3,3a,4,5,5,5a,5b,6-decachlorooctahydro-1,3,4-metheno-2H-cyclobuta(cd)pentalen-2-one.
CAS No. 143-50-0
Chlordecone is a pesticide (insecticide and fungicide) that has been used mainly in the US, Africa, and
South America. It has not been produced after 1976 in the EU or North America. There was an
approved, but very limited, use of chlordecone in Sweden 1973-1978. It has also been used as a
starting material in the production of the pesticide Kelevan. There is no information indicating use or
production of chlordecone for, at least, the last 10 years (1). All information below is extracted from
IPCS Environmental Health Criteria 43, Chlordecone, 1984 (2).
Persistence
We have not found any degradation studies. However, a high persistence is indicated by widespread
occurence of chlordecone in soil and birds in areas where chlordecone has been used, and by an
estimated half-life of 2-5 months in humans. thus, the criterion for persistence seems to be met.
Bioaccumulation
Chlordecone has a relatively high lipophilicity (log K ow 4.5). A fish study from 1982 has generated a
BCF-value of 16 600 in fathead minnow. Studies in marine fish species have given BCF-values
between 1800 and 7100. The presence in birds (<13 mg/kg) and human breast milk supports that the
criterion for bioaccumulation is met.
Toxicity
The acute toxicity of chlordecone is high in algea (7-days EC50 <1 µg/l in four species) and fishes (96
h LC50  70 µg/l in four species). The acute toxicity is also relatively high in mammals, with LD 50
values just below 100 mg/kg in rats and rabbits. Repeated exposure to 1-25 mg/kg food may cause
neurotoxicity, liver toxicity, and morphological changes in endocrine organs, such as the adrenal, the
thyroid, and the testis. Hormonal effects may underlie the reproductive toxicity observed at exposure
to >1 mg/kg body weight/day. Liver tumours are observed both in mice and rats, and the substance is
classified as a possible (2B) carcinogen by IARC. The criteriron for adverse effect is met.
Potential for long-distance transport
Chlordecone is not expected to be degraded by sunlight. Half-lives > 10 days has been measured in the
presence of ethylendiamin. The criteria for long-distance transport is probably met, but there is no
clear evidence.
Consequences
Chlordecone is included in the ECE-LRTAP treaty. If there is production of chlordecone, for which
there is no indications today, it is relevant to include it in the SC. Numerous alternative pesticides are
available today. Possible economic consequences depends on whether there is still any production and
use of chlordecone.
References
1. Pesticide manual
2. WHO-IPCS (1984) Environmental Health Criteria 43 Chlordecone. Geneva, World Health
Organization.
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Pentachlorobenzene
CAS No. 608-93-5
This short summary is mainly based on an IPCS EHC document from 1991 (1) and on a “Preliminary
Risk Profile” document prepared in the Netherlands containing pertinent data on pentachlorobenzene
as a possible candidate for the POP-protocol in UN ECE LRTAP (2).
Pentachlorobenzene was formerly used as a fungicide and as a flame retardant. If these uses still can
be found somewhere in the world is unknown. Probably there is no production any longer.
Pentachlorobenzene is an impurity of up to 2% in hexachlorobenzene and has previously been an
impurity in the pesticide quintozene. The substance is today not registered in any products in the
Swedish product register.(2)
Persistence
Information on the degradability is primarily based on tests examining several chlorobenzenes as a
group. Half-lives of hundred up to several hundreds of days are reported for pentachlorobenzene in
sediment. The ”Preliminary Risk Profile” document reaches the conclusion that the criterion for
persistence in UN ECE LRTAP is met (2). This can also be supported by the presence of
pentachlorobenzene in arctic samples of biota (2).
Bioaccumulation
The log Kow values available vary between 4.8 and 5.2 (2), indicating a high potential for
bioaccumulation, borderline of meeting the criterion of the Stockholm convention. There are also
studies on bioaccumulation performed that confirm a high level of bioaccumulation of
pentachlorobenzene. The measured BCF values vary between 3 400 and 13 000 (2). In several cases
the BCF value exceeds the limit of 5 000 in the criteria of the Stockholm Convention.
Toxicity
Aquatic toxicity has been tested on several species representing different groups of organisms and has
been shown to be very high. The lowest acute LC50 value available for freshwater organisms is 0.25
mg/l for fish and the lowest NOEC value is 0.01 mg/l for crustaceans.(1,2)
In mammals, effects have been observed in different organs, e.g. liver, kidney and the thyroid gland.
The EHC document refers to a subchronic dietary study where NOEL based on histopathological
lesions in male and female rats is reported to be approximately 2.0 and 21.5 mg/kg bodyweight,
respectively (1). The corresponding NOEL in female mice is reported to be 18.3 mg/kg bodyweight. In
male mice no NOEL could be established (1). In the ”Preliminary Risk Profile” document, NOEL
from a subchronic study is reported to be 12.5 mg/kg bodyweight (2). No chronic study is reported.
Available tests indicate pentachlorobenzene to be teratogenic but the evidence is considered to be
insufficient (1,2).
The ”Preliminary Risk Profile” document concludes that pentachlorobenzene meets the criterion for
toxicity in UN ECE LRTAP (2).
Potential for long-range transport
Pentachlorobenzene is expected to degrade very slowly in air, with half-lives estimated to hundreds of
days (2). Vapour pressure is 2.2 Pa at 25 ºC (2). This indicates that pentachlorobenzene possess a
potential for long-range transport. This is also supported by the findings of pentachlorobenzene in
arctic biota.
Pentachlorobenzene is included in the list of priority substances in the field of water policy,
established under Directive 2000/60/EC of the European Parliament and of the Council, establishing a
framework for Community action in the field of water policy (3). In the context of the Water
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Framework Directive, 0.9 ng/l have been reported as the mean concentration for European surface
waters based on 179 samples (83 positive samples) from 7 stations (4). In sediments 14.6 µg/kg have
been reported as the mean concentration based on 459 samples (375 positive samples) from 22 stations
(4). These monitoring data either indicate the substance still to be in use or to be very persistent.
Consequences
Since production and use of pentachlorobenzene probably is nonexistent or insignificant at present, the
impact on society of further restrictive measures will be very small. Provided that future use is
prevented and sources of emissions are identified and eliminated, levels in the environment should
also decrease in the long run. The substance is a ”priority chemical” in the work of OSPAR and has
been selected through the DYNAMEC-process. It is assigned to ”selection box” group E; substances
with PBT properties but which are heavily regulated or withdrawn from the market.
References
1. WHO-IPCS (1991) Environmental health Criteria 128 Chlorobenzenes other than
Pentachlorobenzene. Geneva, World Health Organization.
2. Risk profile polychlorinated pentachlorobenzene, Preliminary risk profile prepared for Ministry of
Housing, Physical Planning and the Environment (VROM, the Netherlands) in the framework of
the project Risk Profiles III, October 2001.
3. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001
establishing the list of priority substances in the field of water policy and amending directive
2000/60/EC. Official Journal of the European Communities L331/1, 15.12.2001.
4. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
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Polybrominated biphenyls (PBB)
CAS No. 59536-65-1 and 67774-32-7 (HexaBB), 61288-13-9 (OctaBB), and 13654-09-6 (DecaBB)
The group polybrominated biphenyls (PBB) contains three technical products, hexabromobiphenyl
(HBB), octabromobiphenyl (OBB) och decabromobiphenyl (DBB). These products are brominated
flame retardants containing 5-7, 7-9, and 9-10 bromine per molecule, respectively. The production of
DBB ceased in 2000, and the production pf HBB and OBB in the 1970’s-1980’s. Thus, there is no
known production of them today, but they (especially DBB) may still be present in old articles. The
production of OBB never exceeded a few percent of the total production of PBBs, and is therefore not
commented further below (1).
Persistence
The persistence of PBB is high in all media, including biota, with half-lives in the order of weeks to
several years (1). Presence of HBB in different environmental samples supports that the criterion is
met for HBB.
Bioaccumulation
The lipophilicity is very high (log Kow = 7 and 8.6 for HBB and DBB, respectively). However, the
bioconcentration potential for these substances differs considerably, with BCF-values well above 5000
for HBB wheras one study on DBB indicate a BCF-value below 5 (1,2). For HBB, there are also data
showing biomagnification in mink (BMF = 60), and presence in both breast milk and cow’s milk
(Germany, 1988, 2 and 0.05 ng/g fat, respectively) (1). With increasing bromination, the size of the
molecule increases, which probably decreases the uptake into organisms. The big size of DBB may
thus explain the low potential for bioaccumulation. Wheras HBB fulfils the criterion, DBB does not.
Toxicity
HBB is a structural analogue to the chlorinated dioxins, and thus very toxic to most organisms, but
perhaps especially to mammals after repeated exposure. Reproductive toxicity in mink and monkeys is
evident at daily exposure to 1 and 0.3 mg HBB/kg feed, respectively. Other effects include
developmental toxicity, immunotoxicity, toxicity to the liver, thyroid and skin, and finally weight loss
and death. The NOAEL for HBB is below 1 mg/kg/day in many species. HBB is classified as a
possible carcinogen (IARC group 2B) (1). DBB is less toxic than HBB, but effects on the liver has
been seen after repeated exposure (NOAEL 35 mg/kg/day). Nonobromobiphenyl has caused liver
tumours in experimental animals (1). It is not clear whether DBB can be considered to fulfill the
criterion for adverse effects.
Potential for long-range transport
Low concentrations of HBB has been found in air samples from indutrial areas of the US, but there is
no data on the presence in air from remote areas. The presence of HBB in marine biota may support
some potential for long-range transport (Atlantic dolphins (20 ng/g fat), and seals from the Baltic (26
ng/g fat), Spitsbergen (1.9 ng/g fat) and Svalbard (0.4 ng/g fat) (1). There is no evidence for DBB
fulfilling the criteria.
Consequences
DBB may only fulfill one of the criteria, questioning whether these conventions are the proper fora for
regulation of DBB. HBB is already included in the UN/ECE LRTAP convention. One reason to
incorporate HBB in the SC as well is to make sure that production will not be restarted anywhere.
Since there is no known production of HBB, there would be no cost for including HBB in the SC.
References
1. WHO-IPCS (1994) Environmental Health Criteria 152 Polybrominated biphenyls. Geneva, World
Health Organization.
2. Elf Atochem, Risk Assessment Decabrombiphenyl, August 1998.
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Dicofol
(CAS No. 115-32-2)
Dicofol is an organochlorine pesticide manufactured from DDT via DDE. The technical product
consists of approximately 80% p, p’-dicofol and 20% o, p’-dicofol. Dicofol - sometimes also referred
to one of its trade names “Kelhane” - is a miticidal pesticide and acaricide used on a wide variety of
fruit, vegetables, ornamental and field crops. Commercial and domestic use was withdrawn in 1997 in
Sweden. The summary below is mainly based upon a Dutch preliminary Risk Profile on Dicofol (1).
Persistence
Degradation of Dicofol in water is pH-dependent. Hydrolysis, to dichlorobenzophenones, is fast under
alkalic conditions (2): the p, p’-isomer hydrolyses with a t1/2 of 85, 4 and 0.02 days at pH 5, 7 and 9,
respectively. The t1/2 for the o, p’-isomer is 47, 0.3 and 0.006 days at the same pH conditions. Several
field dissipation studies have been carried out with Dicofol indicating t ½ in soil to be in the range 3060 days, where the longer t½ values refer to the p, p’ isomer. No leaching from soil was observed
beyond the two to three inches of the soil top layer. This confirms the results from adsorption studies
in which high Koc values were determined and leaching studies in which only 0.3-1.0% of Dicofol
applied was recovered in the leachate. On persistency and mobility in soil, the following is stated in
the EPA RED file:
“Photolysis on soil is not an important route of degradation for Dicofol, possibly due to binding on
the soil and lack of solubility in soil water. o, p'-dicofol degraded with a half-life of 30 days while p,
p’-dicofol degraded with a half-life of 21-30 days on silt loam soil irradiated with artificial light that
does not simulate natural sunlight (MRIDs 40042036 and 40042037). The major degradates identified
in the studies were the o, p' and p, p' isomers of DCBP.”
Dicofol is degraded in both water and soil but the most common isomer p, p’ is more persistent than
the o, p’-isomer. In water, the p, p’-isomer fulfilles the criterion t½ > 2 months only at pH 5. In soil,
the t½ is slightly shorter than the stipulated 6 months. Therefore, it is uncertain if Dicofol if could
meet the criteria for persistency. However the persistency in sediment is still unknown as well as the
significance of cold climate on persistency in water and soil.
Bioaccumulation
Log KOW is reported to be in the range of 4.08 to 5.02. There are two studies of bioaccumulation in
fish (Pimephales promelas and Lepomis macrochirus) that gave BCF-values at 8050 and 13500
respectively. These results are well above the criterion limit. It is therefore concluded that Dicofol
does meet the criterion for bioaccumulation of 5000.
Toxicity
The toxicity of Dicofol has been studied in several animals. The acute toxicity in mammals is
moderate with LD50 in the range 0.4-4.3 g/kg. Chronic and sub-chronic studies reveal enzyme
induction and other changes in the liver, adrenal gland and urinary bladder at doses of 2.5 mg/kg. In a
study, over two years, NOAEL was determined to 0.22 mg/kg and day. There are no results indicating
that Dicofol should be carcinogenic in rats fed 38-47 mg/kg and day for 78 weeks. In mice, however,
increased incidence of liver tumours has been reported at levels from 13.2 mg/kg and day. IARC has
classified Dicofol in category 3 (Not classifiable as a human carcinogen).
Studies of ecotoxicity have revealed high acute toxicity in aquatic environments. LC50 for eastern
oyster has been reported to be as low as 0.015 ppm and the corresponding value for rainbow trout to
be around 0.12 ppm. Dicofol has been shown to affect eggshell quality and a NOEC at 2.5 ppm in feed
has been established, while hatchability was affected at 40 ppm. In falcons, feminised embryos from
females given 5 mg/kg have been reported.
Dicofol is moderately toxic to mammals and not carcinogenic. In wildlife it is reported to be
reprotoxic. In birds, Dicofol may reduce the eggshell quality. Based on the acute toxicity tests, Dicofol
is very toxic to the aquatic environment.
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Potential for long-range transport
The vapour pressure of Dicofol is low, being less than 5,3x10-5 Pa. The calculated t½ in air is 3.1 days.
No experimental data are available.
Based on the vapour pressure, Dicofol is expected to partition between the gas and particle phases in
the atmosphere and is likely to exist largely in the particle phase. The average half-life time for
particles is estimated to be about 3,5 – 10 days and the average lifetime for particles is estimated to be
about 5 - 15 days.
Analyses of Dicofol in connection to sites with high use indicate that the loss is small from these
areas. There is however one study where Dicofol has been analysed in water and sediment downstream an application site. In the water the levels were too low to be detected but in sediment Dicofol
was found at levels ranging from 6.8 to 23.7 ng/L.
There is no information on analyses of Dicofol from Arctic regions. The California Air Toxics
Program has published a Toxic Air Contaminant Fact Sheet on Dicofol where it is stated: “In 1970
787 samples were taken from 14 states and Dicofol was not detected. In 1971 667 samples taken from
16 states showed 0.15 percent positive results for Dicofol, with an average concentration of 9.5 ng/m3.
In 1972 1025 samples were taken from 16 states and Dicofol was not detected”.
Based on vapour pressure and an atmospheric half-life of >2 days, Dicofol meets the criteria for longrange atmospheric transport
Consequences
Dicofol is listed in the Commission Regulation 1490/2002 laying down the detailed rules for the third
stage of EU review program of active substances in plant protection products, which means that
Dicofol will be assessed within the next few years.
References
1. van de Plassche E.J, Schwegler, A.M.G.R, Rasenberg, M.H.C and Balk F. Dicofol. Preliminary
risk profile prepared for the Ministry of Housing, Physical Planning and the Environment
(VROM) in the framework of the project Risk Profiles III. October 2002.
2. European Chemicals Bureau (2000). IUCLID Dataset.
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Endosulfan
(1,4,5,6,7,7-hexachloro-8,9,10-trinorborn-5-en-2,3-ylenbismethylene) sulfite.
CAS No.: 959-98-8 (-isomer), 33213-65-9 (-isomer), 115-29-7 (technical material, consisting of a
2:1 mixture of the - and -isomer). Both isomers are biologically active. The CAS No. for one of the
breakdown products of environmental importance, endosulfan sulfate, is 1031-07-8.
This short summary is mainly based on a draft produced within the context of the OSPAR Convention
(1), and on material presented in reports produced within the EU review of active ingredients in plant
protection products (2). No decision has been taken on endosulfan with respect to inclusion in Annex I
to the "Plant Protection Products Directive" (91/414/EEC), and some required data has not yet been
presented within the EU review. The conclusions of the reports generated within that program are thus
regarded as preliminary only.
Endosulfan is an insecticide which has been used for more than 40 years (1). In addition to use in
agriculture, horticulture and forestry, it is also used to control termites and tsetse fly (1). A minor use
as a wood preservative has also been reported (1). Global annual production was reported to be
12 000-13 000 tonnes in the mid-1990's (1). Consumption within western Europe (EU plus
Switzerland) was reported to 840-1 030 tonnes per year during 1994-1996, to 470-590 tonnes in 19971999, reflecting a reduction in use, mainly in the northern countries (1). In Sweden, pesticides
containing endosulfan as active ingredient were registered until 1995.
Persistence
No hydrolysis of endosulfan occurs under acidic conditions, while hydrolytic half-lives have been
determined to a few weeks in neutral medium, and to about 1 day in alkaline medium (1,2). Thus,
hydrolysis is important especially in marine environments (1). Endosulfan diol is the main metabolite
(1,2). In study of biotic transformation in water/sediment system, endosulfan sulfate and endosulfan
hydrocarboxylic acid were identified as main metabolites, but reliable half-lives have not been
presented as yet (2).
Endosulfan does not seem to exceed the criteria for persistence in itself. The -isomer appears to be
more rapidly degraded than the -isomer (1,2), however, most results are given for the sum of isomers.
Half-lives determined in laboratory studies in soil at 20C (+) range from 26 to 128 days, with a
mean value of 79 days (2). From field studies on dissipation in soil, carried out in Germany and the
US (Georgia and California), DT50s of 16-93 days are reported (for +), with a mean value of 63 days
(2). The main metabolite in soil is endosulfan sulfate (1,2). From a field study in Germany half-lives
of this metabolite were reported to 655 days (planted soil) and 2 210 days (bare ground), while field
studies in Spain and Greece resulted in shorter half-lives; 75 and 47-161 days, respectively (1). Thus,
the criterion for persistency is considered to be met for endosulfan when also considering endosulfan
sulfate.
Bioaccumulation
Reported log Kow values for endosulfan are 2.2-4.8 (1) and 4.9 at pH 4, 4.6-4.8 at pH 7 and 5.6 at pH
10 (2). Thus, a high potential for bioconcentration is indicated. This is confirmed by at least some
studies on bioconcentration in fish. Reported bioconcentration factors (BCFs) range from 350 to
11 580 (1,2). Half-life for depuration in fish has been reported as 2-4 days. In mussel, a BCF of 600
was reported, with a half-life for clearance of about 1 day (1).
Endosulfan and endosulfan residues have been found in numerous food products such as vegetables
(0.5-13 µg/kg), seafood (0.0002-1.7 µg/kg), milk and tobacco (1). In 1995-98 endosulfan and
endosulfan sulfate were found in 0.1-9.9% of meat samples and in fish, crab and mollusc samples in
Germany (1). In recent samples of blue mussel from the south Swedish coast, 0.03-0.06 mg endosulfan/kg lipid was measured (lipid content of the mussels being approx. 1 %) (5).
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- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
In mammals, excretion rate seems to be rapid after a single oral dose. Following repeated dosing in rat,
the residues reached maximum values after about 3 weeks, with the highest concentrations in kidneys
(2). These residues consisted of polar compounds, whereas endosulfan sulfate was the major
component in the residues in fat. Following cessation of dosing, the residues in all tissues fell
significantly over the next 5 days to levels which for most tissues were similar to those seen 24 hours
after a single oral dose (2). Residue levels in reproductive organs were not higher than in general
organs. In mammals, endosulfan is converted to the following metabolites: endosulfan sulfate,
endosulfan diol, endosulfan ether, endosulfan hydroxyether, endosulfan lactone, and a number of polar
metabolites which probably are the conjugates of the metabolites (2).
Although some of the reported BCF values exceed 5000, it is not clear whether endosulfan meet the
criterion for bioaccumulation. It is clear that varying degree of bioaccumulation occurs in organisms
having relatively low metabolic capacity. In mammals however bioaccumulation or biomagnification
is not likely to occur because metabolism in mammals seems to be fairly rapid. Also, no data on
measured concentrations in mammals was found.
Toxicity
Endosulfan is highly toxic for aquatic organisms, with several EC/LC50 values in the µg/l range
reported for fish and aquatic invertebrates (1,2). Lowest LC50 value from acute toxicity test was 0.04
µg/l for the marine crustacean Penaeus duorarum (1,2). The -isomer seems to be more toxic than the
-isomer. In carp, LC50 (96 h) was 0.75 µg/l for -endosulfan, > 3.1 µg/l for -endosulfan (2). In
daphnids, EC50 (48 h) was 224 µg/l for -endosulfan, 528 µg/l for -endosulfan (2). In a pond study,
fish mortality was observed at water concentrations of 0.4 and 1 µg/l (2).
Some NOEC values reported from long-term studies on aquatic organisms are: 21-d and 28-d NOEC
in fish 0.05 µg/l (1,2); 21-d NOEC in daphnids 63 µg/l (2) and 14-d NOEC in daphnids 49 µg/l (1).
Short-term toxicity of metabolites have also been investigated in fish (carp) and daphnids (2):
endosulfan sulfate: LC50 (96 h) fish 2.2 µg/l, EC50 (48 h) daphnids 300 µg/l,
endosulfan lactone: LC50 (96 h) fish 570 µg/l, EC50 (48 h) daphnids >1 300 µg/l,
endosulfan ether: LC50 (96 h) fish > 1 650 µg/l, EC50 (48 h) daphnids 580 µg/l,
endosulfan hydroxyether: LC50 (96 h) fish 2 300 µg/l, EC50 (48 h) daphnids 1 600 µg/l.
Endosulfan is also toxic to birds; Acute LD50 was 28 mg/kg bw; Short-term dietary LC50 was 805 ppm,
and NOEC for effects on reproduction was as low as 30 ppm (2).
To mammals, endosulfan is highly toxic via the oral route (1), with LD50 values of 10-23 mg/kg bw
reported for female rat (2). Lowest relevant NOAEL has been established as 0.6 mg/kg bw/d for
neurotoxic effects observed in a 1-year study in dog (2). In rat and mouse, 90-d NOAEL was 3.8
mg/kg bw/d (male rat) and 2.3 mg/kg bw/d (mouse) (2). Stimulation of the central nervous system is
the major characteristic of endosulfan poisoning (1). Sheep and pigs grazing on fields sprayed with
endosulfan showed lack of muscle coordination and blindness; reversible blindness has also been
reported for cows grazing in contaminated fields (1).
Endosulfan has no carcinogenic potential (1,2). The substance is not mutagenic in vitro and in vivo in
somatic ells, but some positive results have been reported in in vivo studies in germ cells (2). There are
no evidence of teratogenicity or effects on reproduction in mammals (1,2). Several data suggest that
endosulfan as well as endosulfan sulfate are endocrine disrupters (estrogenic action), however, as yet it
seems as if there is no clear consensus on this topic (1).
The toxicity of metabolites has also been tested in mammals (2):
For endosulfan sulfate an LD50 of 568 mg/kg bw was reported for male rat, 25-50 mg/kg bw for
females. For endosulfane lactone, male rats were more sensitive than females, with LD50 < 200 mg/kg
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
bw. Acute oral toxicity for endosulfan hydroxyether was 1750 mg/kg bw in female rat. A low acute
oral toxicity was reported for endosulfan ether and endosulfan diol. For endosulfan sulfate, lowest 90d NOAEL was reported for dog, 0.75 mg/kg bw/d (males and females). For endosulfan diol 90-d
NOAEL was about 8-9 mg/kg bw/d in rat and dog.
From all these data it is clear that endosulfan as well as some of its metabolites meet the criterion for
adverse effects.
Potential for long-range transport
The vapour pressure of -endosulfan is 1.9 x 10-3 Pa, of -endosulfan 9.2 x 10-5 Pa (1). The value of
Henry's Law Constant indicates a potential for volatilisation from moist surfaces, being 1.1 Pa x m3 x
mol-1 for -endosulfan and 0.2 Pa x m3 x mol-1 for -endosulfan (2). Studies on leaf surfaces and soil
also show a high loss rate to the atmosphere, and a higher loss of - than of -endosulfan (2). The
half-life in air has been calculated to 8.5-27 days (2). Endosulfan has been detected in samples of
Arctic air (1,2) and Arctic sea water (1). Reported mean concentrations in Arctic air are 3.0-8.3 pg/m3
(1). From this information, the criterion for long-range transport is considered to be met.
In European rivers, the - and -isomers as well as the breakdown product endosulfan sulfate been
detected (1). The concentrations are low but considering the high toxicity to aquatic organisms they
are not negligible. In addition, the presence in river water points to a route of transport from the areas
of use to, e.g., sea water.
Endosulfan is included in The list of priority substances in the field of water policy, established under
Directive 2000/60/EC of the European Parliament and of the Council, establishing a framework for
Community action in the field of water policy (3). In the context of the Water Framework Directive,
the following mean concentrations have been reported for European surface waters (4):
0.017 µg/l
(191 samples from 27 stations; 93 above determination limit)
-isomer:
0.0088 µg/l
(180 samples from 25 stations; 82 above determination limit)
-isomer:
endosulfan sulfate: 0.0094 µg/l
(246 samples from 37 stations; 126 above determination limit)
For the sediment phase, the following mean concentrations were reported (4):
37.8 µg/kg
(75 samples from 20 stations; 45 above determination limit)
-isomer:
Consequences
The reported reduction of European consumption, mainly achieved in northern countries (see above),
points to a possibility to reduce the usage also globally without too severe impacts on agriculture,
horticulture and forestry. However, it is most likely that countries suffering from a higher insect pest
pressure would encounter significantly larger difficulties. In comparison with potential chemical
substitutes for endosulfan, it is important to consider also that target organisms are less likely to
develop resistance against endosulfan, and also that endosulfan show lower toxicity to beneficial
insects than some of the alternatives (1).
Within the work under the OSPAR Convention, endosulfan has been included in the List of Chemicals
for Priority Action under the OSPAR Strategy with regard to Hazardous Substances, with the ultimate
aim to achieve concentrations in the marine environment close to zero.
References
1. Final Draft OSPAR Background Document on Hazardous Substances Identified for Priority
Action - Endosulphan - Presented by Germany. OSPAR 02/7/9-E. To Meeting of the OSPAR
Commission, Amsterdam, 24-28 June 2002. OSPAR Convention for the Protection of the Marine
Environment of the North-East Atlantic.
2. - European Commission Peer Review Programme. Draft assessment report prepared in the context
of the possible inclusion of the following active substance in Annex I of Council Directive
91/414/EEC: Endosulfan. Volumes 1 and 3, December 1999. Addendum to Volume 3, May 2001,
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
October 2001 and January 2002, respectively. Rapporteur Member State: Spain.
- European Commission Co-Operation. Concise outline Reports of ECCO meetings 102, 103 and
105, 2001: Endosulfan.
3. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001
establishing the list of priority substances in the field of water policy and amending directive
2000/60/EC. Official Journal of the European Communities L331/1, 15.12.2001.
4. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
5. Svenska Naturvårdsverket. Redovisning från nationell miljöövervakning 2002. Endosulfan. C
Esbjörnson, examensarbete vid Karolinska Institutet.
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
Methoxychlor
1,1,1-trichloro-2,2-bis(4-methoxyphenyl)ethane
CAS No. 72-43-5
Metoxychlor has been used as a pesticide (insecticide), as a biocide and as a veterinary product. There
are some structural similarities to DDT. Based on information in the draft OSPAR background
document on methoxychlor, produced by Finland (3), there is no current production or use of
metoxychlor in the OSPAR countries (note that the situation in Spain is unclear because of missing
information, and that there are some EU-countries which are not contracting parties to the OSPAR
convention). The production of methoxychlor in the USA was in the beginning of 1990 about 150 to
300 tonnes. The substance has been used in 19 different products in Sweden but since 1991 it is not
allowed as a pesticide anymore.
Persistence
The degradation of methoxychlor is slow during aerobic conditions but much faster during anaerobic
conditions. The half-life has been measured to < 30 days for anaerobic degradation in sediment and >
100 days during aerobic degradation in sediment. This degradation pattern is the same in soil. The
half-life has been measured to 46 days, which indicate that methoxychlor is less persistent than DDT.
The degradation products of methoxychlor are suspected to be endocrine disruptors (7). It is unclear if
methoxychlor meets the criterion for persistency.
Bioaccumulation
Log Kow is 4.7 – 5.1. Bioconcentration factors (BCFs) in three different fish species are (1): 113-264
(sheepshead minnow), 1 500 (western mosquito fish, test duration only 72 hours), and 8 300 (fathead
minnow, flow-through system). Methoxychlor seems to meet the criterion for bioaccumulation.
Toxicity
Methoxychlor is extremely toxic to aquatic organisms. The acute toxicity (LC50) for fish is 52 µg/l and
67 µg/l for rainbow trout and bluegill sunfish, respectively. The acute toxicity (LC50) for daphnids is
as low as 0.8 µg/l. Methoxychlor is an endocrine disrupting chemical (2). The criterion for adverse
effects is considered to be met.
Potential for long-range transport
Methoxychlor has a low vapour pressure (1.9 x 10-4 Pa) but Henry's Law Constant is 1.6 Pa x m3 x
mol-1 which indicates a potential for volatilisation. The half-life in air is reported to be only 4 to 6.8
hours, which does not meet the criterion for long-range transport. However, methoxychlor has been
detected in rain and snow from remote areas in Canada, which indicate that methoxychlor may be
persistent in the atmosphere and undergo long-range transport. (4)
Consequences
Since methoxychlor has not been notified under the "Plant Protection Produts Directive" 91/414/EEC
(i.e., there is no stakeholder which will produce a dossier for the substance), any plant protection
products containing this substance must be withdrawn from the EU market by July 2003. This
suggests that the economical consequences of inclusion in the Stockholm Convention will be
insignificant within the EU.
The global use of methoxychlor is unknown and it is therefore to early to predict the global
consequences for restrictions. Due to the fact that the demand for methoxychlor within the EU seems
not to exist anymore, and with the knowledge that there are other insecticides on the market, it is
reasonable to assume that acceptable alternatives exist, at least for some areas of use.
The substance is a ”priority chemical” in the work of OSPAR and has been selected through the
DYNAMEC-process. It appears in the ”selection box” group A which means that that the substance
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
probably has POP-like properties and there are some indication on production, use or occurrence in the
environment.
References
1. Aquire (2001), ecotoxicological database on the internet, established by USEPA
2. Cummings, A. M. (1997) Methoxychlor as a model for environmental estrogens. Critical Reviews
in Toxicology 27 (4): 367 – 379.
3. Draft OSPAR Background Document on Methoxychlor, Presented by Finland, Meeting of the
Working Group on Priority Substances (SPS) in Arona 15-19 October 2001.
4. Howard, P. H., (1991) Handbook of Environmental Fate and Exposure data for Organic
Chemicals, vol 3, p 502 – 507.
5. Keith L. H., (1998) Environmental Endocrine Disruptors, A handbook of property data. p 802 –
830.
6. Metabolic Pathways of Agrochemicals (1999) Insecticides and Fungicides, part two, p 181 – 185
7. Toxicological profile for Methoxychlor, Atlanta (1994), GA, Agency for Toxic Substances and
Disease Registry.
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
Hexabromocyclododecane (HBCD)
CAS No. 25637-99-4
HBCD is a brominated flame retardant with its main use in insulation material of polystyren, e.g., for
buildings. There is also some use in textiles. HBCD is a high production volume chemical both in the
EU and in the US, and there is production in Japan as well. HBCD is currently being risk assessed in
the EU.
Persistence
Based on acceptable laboratory studies, HBCD is not readily biodegradable. One study may indicate
some biodegradation (1). Simulation half life studies are in progress in the EU ESR program to allow a
conclusion whether the criterion is met or not. The presence of HBCD in different environmental
samples (sewage treatment plants, soil, fish, birds, and seals) supports a relatively high persistence.
Bioacumulation
Two studies in fish have given BCF-values well above the criteria cut-off (9000-18 000) (1). The
presence of HBCD in fish (e.g. Baltic herring), birds (guillemot and peregrine falcon egg), Baltic seals
(100 ng/g fat) human food stuff (e.g. meat) supports that bioaccumulation may exist (2,3).
Toxicity
At water-soluble concentrations of HBCD (a few µg/l), HBCD affects the growth of algea
(Skeletonema costatum, EC50 (72 h) 11 µg/l) and growth, reproduction, and survival of Daphnia
magna (1). The NOEC for daphnia magna is 3 µg HBCD/litre. The toxicity to fish is low (1).
In mammals, the liver and thyroid system is affected after repeated exposure, but no conclusive
NOAEL can be set. Based on the ecotoxicity, the criterion for adverse effects is met.
Potential for long-range transport
HBCD is not very volatile, but there is data indicating its presence in air of Scandinavian background
areas (0.002-0.28 ng/m3) and deposition to soil (1.6 - 13 ng/m2 and day) (1,3). Potential for long-range
transport is supported by QSAR-modeling (hydroxyl-mediated degradation in air), giving a half-life of
1.8 days. The criterion is 2 days, but considering the uncertainties in the modelling, 1.8 days is close
enough to support a potential for long-range transport.
Consequences
According to Industry, there are no alternative flame retardants that can be used in polystyrene. The
incorporation of flame retardants into polystyrene is mandatory in many countries due to strict fire
safety standards. A potential regulation of HBCD may thus be preceeded by development of other
flame retardants or changes in the fire safety standards.
References
1. EU Risk Assessment Report on Hexabromocyclododecane, CAS-No. 25637-94-4, draft of 2002.
2. Abstracts from The Second International Workshop on Brominated Flame Retardants – BFR 2001
Stockholm May 14-16, 2001.
3. IVL (2001) HBCD i Sverige - screening av ett bromerat flamskyddsmedel. Sternbeck J et al. IVL
rapport B1434. Stockholm, IVL Svenska Miljöinstitutet AB, november 2001.
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
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Hexachloro-1,3-butadiene
CAS No. 87-68-3
This short summary is mainly based on an IPCS EHC document from 1994 (1) and on a “Preliminary
Risk Profile” document prepared in the Netherlands containing pertinent data on hexachlorobutadiene
as a possible candidate for the POP-protocol in UN ECE LRTAP (2).
Hexachlorobutadiene is formed mainly as a by-product during the manufacture of certain chlorinated
hydrocarbons. The global annual production was estimated to be 10 000 tonnes in 1982 (1). The use is,
inter alia, as a chemical intermediate in the manufacture of rubber compounds, and in lesser amounts
as a solvent, heat transfer liquid, and hydraulic fluid. Hexachlorobutadiene has also been used as a
fumigant and may still be used as a fumigant in some countries (2). The substance is today not
registered in any products in the Swedish product register.
Persistence
Test data regarding degradability is scarce and the estimations reported in the “Preliminary Risk
Profile” document cannot clarify the picture. Accordingly, that document reaches the conclusion that
insufficient evidence is available on the persistence of hexachlorobutadiene but that the substance
probably is recalcitrant to biodegradation under aerobic conditions (2). This assumption could also be
supported by the relative abundance of monitoring data as regards hexachlorobutadiene. The substance
can be found in different compartments of the environment including biota, primarily in industrial
regions but has also been detected in biota in northern Canada (2).
Bioaccumulation
The log Kow values available vary between 3.7 and 4.9 (1,2). The ”Preliminary Risk Profile”
document recommends the value 4.9 to be used, which indicates a high potential for bioaccumulation
although just below the limit of the screening criterion of the Stockholm Convention (2). Several
bioaccumulation studies confirm a high level of bioaccumulation of hexachlorobutadiene, showing
BCF values of up to 19 000, clearly fulfilling the bioaccumulation criterion (2).
Toxicity
Aquatic toxicity has been tested on species representing different groups of organisms and has been
shown to be very high. The lowest acute LC50 value available is 0.032 mg/l for crustaceans and the
lowest NOEC value is 0.0065 mg/l for fish.(1,2)
One reliable study with birds as test animal is available. In this 90 day toxicity test with Japanese quail
a NOAEL of 3 mg/kg was found.(2)
In short-term and long-term diet studies with rats and mice the kidney has shown to be the major target
organ. The NOAEL for renal toxicity in rats in a 2-year study was 0.2 mg/kg body weight per day.
Based on this 2-year diet study the International Agency for Research on Cancer (IARC) has found
limited evidence for carcinogenicity in animals and insufficient evidence in humans. IARC has placed
hexachlorobutadiene in Group 3 (not classifiable as to human carcinogenicity).(1,2)
The ”Preliminary Risk Profile” document concludes that the criterion for toxicity in UN ECE LRTAP
is met (2).
Potential for long-range transport
Hexachlorobutadiene is expected to degrade very slowly in air, with half-lives hundreds of days in
different estimations (1,2). Vapour pressure is about 20 Pa at 20 ºC and Henry`s Law Constant is about
1000 Pa m3/mol (1,2). This indicates that hexachlorobutadiene possess a potential for long-range
transport. This is also supported by the findings of hexachlorobutadiene in biota in northern Canada.
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
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Hexachlorobutadiene is included in the list of priority substances in the field of water policy,
established under Directive 2000/60/EC of the European Parliament and of the Council, establishing a
framework for Community action in the field of water policy (3). In the context of the Water
Framework Directive, 0.0093 µg/l have been reported as the mean concentration for European surface
waters based on 1391 samples (1154 positive samples) from 68 stations (4). These monitoring data,
though low concentrations, support the persistency as well as the potential for long-range transport of
the substance.
Consequences
The available information on production and use is old and probably does not reflect the situation
today. Possible use is probably limited and impact on society of restrictive measures aimed at direct
production will be very small. However, measures intended for the cessation of the emissions of, and
the formation of, hexachlorobutadiene as a by-product during the manufacture of other chlorinated
hydrocarbons may have a larger economic impact. Provided that future use is prevented and sources of
emissions are identified and eliminated, concentrations in the environment should also decrease in the
long run. The substance is a ”priority chemical” in the work of OSPAR and has been selected through
the DYNAMEC-process. It is assigned to ”selection box” group E; substances with PBT properties but
which are heavily regulated or withdrawn from the market.
References
1. WHO-IPCS (1994) Environmental health Criteria 156 Hexachlorobutadiene. Geneva, World
Health Organization.
2. Risk profile polychlorinated hexachlorobutadiene, Preliminary risk profile prepared for Ministry
of Housing, Physical Planning and the Environment (VROM, the Netherlands) in the framework
of the project Risk Profiles III, October 2001.
3. Decision No 2455/2001/EC of the European Parliament and of the Council of 20 November 2001
establishing the list of priority substances in the field of water policy and amending directive
2000/60/EC. Official Journal of the European Communities L331/1, 15.12.2001.
4. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
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Polycyclic aromatic hydrocarbon (PAH)
PAH is a large group of substances, which consist of molecules with 2 to 3 or more aggregated
benzene rings. Most of them are generated in connection with different thermal processes and are
emitted to the environment via both point and diffuse sources. Some of them have, as individual
substances, a commercial use. Some PAH-substances are further toxic, some bioconcentrate in
invertebrates in the aquatic environment, in fact to some extent they exhibit properties which are POPlike. PAH metabolise in vertebrates, but the metabolites are reactive and some are known to be
carcinogenic.
In order to take a closer look on PAH-substances, six substances were selected, the so called Borneff
6: benzo[a]pyrene (CAS No. 50-32-8), benzo[ghi]perylene (CAS No. 191-24-2), indeno[1,2,3cd]pyrene (CAS No. 193-39-5), benzo[b]floranthene (CAS No. 205-99-2), benzo[k]fluoranthene
(CAS No. 207-08-9) och fluoranthene (CAS No. 206-44-0) (1).
Persistence
All selected PAH-substances fulfil the criteria for peristence, that is half-lifes in water exceeding 60
days, and in soil 180 days. This is not a fact for all PAH-susbstances (1,2). It is, thereby, doubtful if
PAH-substances fulfil the criterion for persistence.
Bioaccumulation
All selected PAH-substances have log Kow exceeding 5 and most of them have BCF-values exceeding
5 000. The BCF values are related to organisms at lower tropic-levels. Higher organisms do
metabolise PAH. However, many metabolites are toxic (1,2). PAH-substances, thereby, bioconcentrate
in lower organisms, but not in higher organisms.
It is doubtful if PAH-substances fulfil the criterion for bioaccumulation.
Toxicity
All selected PAH-substances are genotoxic, which is known also for many other PAH-substances. All,
except benzo(ghi)perylene, have been proven to be carcinogenic, which is also known for other
individual PAH-substances. Indivual PAH-substances are classified in IARC-groups 2A (probably
carcinogenic to humans) and 2B (possibly carcinogenic to humans), but none in group 1 (carcinogenic
to humans). Complex mixtures containing different PAH, such as tar, soot, smoke from aluminum
production and tobacco smoke, are classified in group 1.
Low molecular PAH are toxic towards several aquatic organisms with EC50-values below 0.001 mg/l.
Reproduction disturbances and mutagenicity/carcinogenicity in aquatic organisms have been observed
even for some high-molecular substances (1,2). PAH-substances are expected to fulfil the criterion for
adverse effects.
Potential for long-range transport
Most PAH-substances, with the exception for some low molecular substances, have low volatility and
water solubility. In the atmosphere they are primarily adsorbed on particles and can by that be
transported long-range. Transport in this way has been reported to the Artic region, but it may be
problematic to define the origin of the substances. Most of the risks related to PAH in the environment
are, however, more associated to the local levels than the levels in remote areas (3).
PAH-substances are already included in the LRTAP POP-protocol, targeted for a reduction of the total
discharges mainly through the application of best available technology (BAT) at some large point
sources, such as the production of coke, anode and aluminium (3).
The PAH-group is treated within the Water Framework Directive 2000/60/EC, where polycyclic
aromatic hydrocarbons, and some individual substances within the group, are included in the list of
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Governmental commission, National Cemicals Inspectorate/Swedish EPA
- Priority list for chemicals to LRTAP and SC Annex 1 to the Interim Report
priority substances (4). For some PAHs1 levels of 0.0091-0.036 µg/l have been reported in European
surface waters. For sediments, levels of 381-742 µg/kg for the same substances have been reported
(5).
Consequences
A decision for a total cessation of discharges of PAH-substances would end all thermal processes, as
would the use of automobiles and energy production from fossil fuels. The costs would be enormous.
Even phasing out the use of creosote should be in the order of 220 million Euro/year, only through
shorter intervals in exchanging electricity distribution- and telephone poles, depending on the lower
efficiency of the substitutes. In developing countries such a decision would further forbid processes
like house-warming and cooking with the aid of wood combustion (6).
The eventual inclusion of PAH in the Stocholm Convention can be discussed considering:
 their possibilities to fulfil the criteria
 the possibility that conventions can be weakened if substances are included, where there is
little or no possibility for elimination
 effects of PAH-substances are more considered as a local or a regional problem, not a problem
caused by long-range transported PAH
References
1. PAH as a POP, Sara Edlund, Thesis, Internationella miljöinstitutet, Lund, September 2001.
2. OSPAR Draft Background Document on Polycyclic Hydrocarbons, OSPAR 2001
3. 1979 Convention on Long-Range Transboundary Air Pollution, 1998 Protocol on Persistent
Organic Pollutants, United Nations, ECE/EB.Air/66, 1999
4. DECISION No 2455/2001/EC OF THE EUROPEAN PARLIAMENT AND OF THE COUNCIL
of 20 November 2001 establishing the list of priority substances in the field of water policy and
amending Directive 2000/60/EC.
5. Fraunhofer-Institut (1999) Revised Proposal for a List of Priority Substances in the Context of the
Water Framework Directive (COMMPS Procedure). Draft Final Report. Declaration ref.:
98/788/3040/DEB/E1. Schmallenberg, Fraunhofer Institut Umweltchemie und Ökotoxikologie.
6. Socio-Economic Impacts of the Identification of Priority Hazardous Substances under the Water
Framework Directive, Final Report prepared for European Commission Directorate-General
Environment, RPA, December 2000
1
benzo-a-anthracene, benzo-a-pyrene, benzo-b-fluoroanthene, benzo-g,h,i-perylene, benzo-k-fluoranthene, samt
indeno(1,2,3-cd)pyrene.
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