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Concise International Chemical Assessment Document 68

First draft prepared by Mr Peter Watts, Toxicology Advice & Consulting Ltd, Sutton, United Kingdom

Published under the joint sponsorship of the United Nations Environment Programme, the International Labour Organization, and the World Health Organization, and produced within the framework of the Inter-Organization Programme for the Sound Management of Chemicals.

The International Programme on Chemical Safety (IPCS), established in 1980, is a joint venture of the United Nations Environment Programme (UNEP), the International Labour Organization (ILO), and the World Health Organization (WHO). The overall objectives of the IPCS are to establish the scientific basis for assessment of the risk to human health and the environment from exposure to chemicals, through international peer review processes, as a prerequisite for the promotion of chemical safety, and to provide technical assistance in strengthening national capacities for the sound management of chemicals.

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WHO Library Cataloguing-in-Publication Data

Tetrachloroethene.

(Concise international chemical assessment document ; 68)

Draft prepared by Peter Watts; edited by Marla Sheffer.

1. Tetrachloroethylene - adverse effects. 2. Tetrachloroethylene - toxicity.

3. Environmental exposure. 4. Risk assessment. I. Watts, Peter. II. Sheffer, Marla.

III. International Programme on Chemical Safety. IV. World Health Organization. V. Series.

ISBN 92 4 153068 5          (NLM Classification: QV 633)
ISBN 978 92 4 153068 2
ISSN 1020-6167

©World Health Organization 2006

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Risk assessment activities of the International Programme on Chemical Safety, including the production of Concise International Chemical Assessment Documents, are supported financially by the Department of Health and Department for Environment, Food & Rural Affairs, UK, Environmental Protection Agency, Food and Drug Administration, and National Institute of Environmental Health Sciences, USA, European Commission, German Federal Ministry of Environment, Nature Conservation and Nuclear Safety, Health Canada, Japanese Ministry of Health, Labour and Welfare, and Swiss Agency for Environment, Forests and Landscape.

Technically and linguistically edited by Marla Sheffer, Ottawa, Canada, and printed by Wissenchaftliche Verlagsgesellschaft mbH, Stuttgart, Germany

FOREWORD

Concise International Chemical Assessment Documents (CICADs) are published by the International Programme on Chemical Safety (IPCS) — a cooperative programme of the World Health Organization (WHO), the International Labour Organization (ILO), and the United Nations Environment Programme (UNEP). CICADs have been developed from the Environmental Health Criteria documents (EHCs), more than 200 of which have been published since 1976 as authoritative documents on the risk assessment of chemicals.

International Chemical Safety Cards on the relevant chemical(s) are attached at the end of the CICAD, to provide the reader with concise information on the protection of human health and on emergency action. They are produced in a separate peer-reviewed procedure at IPCS. They may be complemented by information from IPCS Poison Information Monographs (PIM), similarly produced separately from the CICAD process.

CICADs are concise documents that provide summaries of the relevant scientific information concerning the potential effects of chemicals upon human health and/or the environment. They are usually based on selected national or regional evaluation documents or on existing EHCs. Before acceptance for publication as CICADs by IPCS, these documents undergo extensive peer review by internationally selected experts to ensure their completeness, accuracy in the way in which the original data are represented, and the validity of the conclusions drawn.

The primary objective of CICADs is characterization of hazard and dose–response from exposure to a chemical. CICADs are not a summary of all available data on a particular chemical; rather, they include only that information considered critical for characterization of the risk posed by the chemical. The critical studies are, however, presented in sufficient detail to support the conclusions drawn. For additional information, the reader should consult the identified source documents upon which the CICAD has been based.

Risks to human health and the environment will vary considerably depending upon the type and extent of exposure. Responsible authorities are strongly encouraged to characterize risk on the basis of locally measured or predicted exposure scenarios. To assist the reader, examples of exposure estimation and risk characterization are provided in CICADs, whenever possible. These examples cannot be considered as representing all possible exposure situations, but are provided as guidance only. The reader is referred to EHC 170.¹

While every effort is made to ensure that CICADs represent the current status of knowledge, new information is being developed constantly. Unless otherwise stated, CICADs are based on a search of the scientific literature to the date shown in the executive summary. In the event that a reader becomes aware of new information that would change the conclusions drawn in a CICAD, the reader is requested to contact IPCS to inform it of the new information.

Procedures

The flow chart shows the procedures followed to produce a CICAD. These procedures are designed to take advantage of the expertise that exists around the world — expertise that is required to produce the high-quality evaluations of toxicological, exposure, and other data that are necessary for assessing risks to human health and/or the environment. The IPCS Risk Assessment Steering Group advises the Coordinator, IPCS, on the selection of chemicals for an IPCS risk assessment based on the following criteria:

• there is the probability of exposure; and/or

• there is significant toxicity/ecotoxicity.

Thus, it is typical of a priority chemical that:

• it is of transboundary concern;

• it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management;

• there is significant international trade;

• it has high production volume;

• it has dispersive use.

Advice from Risk Assessment Steering Group Criteria of priority: there is the probability of exposure; and/or there is significant toxicity/ ecotoxicity. Thus, it is typical of a priority chemical that it is of transboundary concern; it is of concern to a range of countries (developed, developing, and those with economies in transition) for possible risk management; there is significant international trade; the production volume is high; the use is dispersive. Special emphasis is placed on avoiding duplication of effort by WHO and other international organizations. A prerequisite of the production of a CICAD is the availability of a recent high-quality national/regional risk assessment document = source document. The source document and the CICAD may be produced in parallel. If the source document does not contain an environmental section, this may be produced de novo, provided it is not controversial. If no source document is available, IPCS may produce a de novo risk assessment document if the cost is justified. Depending on the complexity and extent of controversy of the issues involved, the steering group may advise on different levels of peer review: standard IPCS Contact Points above + specialized experts above + consultative group

The Steering Group will also advise IPCS on the appropriate form of the document (i.e. a standard CICAD or a de novo CICAD) and which institution bears the responsibility of the document production, as well as on the type and extent of the international peer review.

The first draft is usually based on an existing national, regional, or international review. When no appropriate source document is available, a CICAD may be produced de novo. Authors of the first draft are usually, but not necessarily, from the institution that developed the original review. A standard outline has been developed to encourage consistency in form. The first draft undergoes primary review by IPCS to ensure that it meets the specified criteria for CICADs.

The second stage involves international peer review by scientists known for their particular expertise and by scientists selected from an international roster compiled by IPCS through recommendations from IPCS national Contact Points and from IPCS Participating Institutions. Adequate time is allowed for the selected experts to undertake a thorough review. Authors are required to take reviewers’ comments into account and revise their draft, if necessary. The resulting second draft is submitted to a Final Review Board together with the reviewers’ comments. At any stage in the international review process, a consultative group may be necessary to address specific areas of the science. When a CICAD is prepared de novo, a consultative group is normally convened.

The CICAD Final Review Board has several important functions:

• to ensure that each CICAD has been subjected to an appropriate and thorough peer review;
• to verify that the peer reviewers’ comments have been addressed appropriately;
• to provide guidance to those responsible for the preparation of CICADs on how to resolve any remaining issues if, in the opinion of the Board, the author has not adequately addressed all comments of the reviewers; and
• to approve CICADs as international assessments.

Board members serve in their personal capacity, not as representatives of any organization, government, or industry. They are selected because of their expertise in human and environmental toxicology or because of their experience in the regulation of chemicals. Boards are chosen according to the range of expertise required for a meeting and the need for balanced geographic representation.

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This CICAD² on tetrachloroethene was drafted by Toxicology Advice & Consulting Ltd based on four source documents. A report produced jointly by the Nordic Expert Group for Criteria Documentation of Health Risks from Chemicals and the Dutch Expert Committee on Occupational Standards (de Raat, 2003) and the IARC evaluation of the carcinogenicity of tetrachloroethene (IARC, 1995) were used to draft most of the human health sections, and a USEPA (2003) discussion paper on neurotoxicity was used as the basis for sections on neurotoxicity. The environmental sections were drafted using the final draft EU Risk Assessment Report (Environment) (EC, 2001).³ Data identified as of 2001 (EC, 2001), 1995 (IARC, 1995), 2002 (USEPA, 2003), and 2002 (de Raat, 2003) were considered in the source documents. A comprehensive literature search of several online databases was conducted in May 2004 to identify any references published subsequent to those incorporated in the source documents. Information on the nature of the peer review and the availability of the source documents is presented in Appendix 2. Information on the peer review of this CICAD is presented in Appendix 3. This CICAD was first discussed as an international assessment at a meeting of the Final Review Board, held in Hanoi, Viet Nam, on 27 September – 1 October 2004. Participants of the Final Review Board meeting are listed in Appendix 4. Due to conflicting views on interpretation of data on critical end-points, the draft CICAD was referred to a WHO Consultative Group, which met at the United Kingdom Centre for Ecology and Hydrology in Monkswood, Cambridgeshire, on 25–27 April 2005. Participants of the Consultative Group meeting are listed in Appendix 5. The CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Nagpur, India, on 31 October – 3 November 2005. Participants of the Final Review Board meeting are listed in Appendix 6. The International Chemical Safety Card (ICSC 0076) for tetrachloroethene, produced by IPCS (2000), has also been reproduced in this document.

Tetrachloroethene (CAS No. 127-18-4) is a clear, colourless, volatile liquid with an etheric odour.

The most recent figures for annual production of tetrachloroethene in the EU and the USA are 164 000 and 160 000 tonnes, respectively, and apply to 1994 (EU) and 1998 (USA). Production in the EU and USA has approximately halved over the last 10–20 years. The major uses of tetrachloroethene are in the dry cleaning of textiles and as a chemical intermediate. Tetrachloroethene is also used in metal degreasing. Tetrachloroethene is released to the atmosphere during use, and the major portion of atmospheric releases is attributed to evaporative losses during dry cleaning.

Tetrachloroethene volatilizes readily from soil and surface water and undergoes degradation in air to produce phosgene, trichloroacetyl chloride, hydrogen chloride, carbon monoxide, and carbon dioxide. Its half-life in air is approximately 3–5 months. In water, it is resistant to abiotic and aerobic degradation, but it is biodegraded under anaerobic conditions to yield trichloroethene, dichloroethene, vinyl chloride, ethane, and ethene. It does not bioaccumulate to any significant extent in aquatic organisms. Tetrachloroethene is detected in outdoor air, usually at concentrations below 1–2 µg/m³. In Dutch homes, the median indoor air concentration was 4 µg/m³, with maxima of about 50–200 µg/m³. Concentrations can be much higher in buildings where dry cleaning operations are carried out. In drinking-water, tetrachloroethene concentrations are generally below 1–10 µg/l. Higher concentrations can occur in groundwater near polluted sites. Mean intake from food, drinking-water, and air is approximately 0.5–3 µg/kg body weight per day.

Tetrachloroethene is well absorbed by mammals following inhalation or oral exposure and is subsequently distributed mainly to adipose tissue, with smaller amounts found in the liver, brain, kidneys, and lungs. Dermal absorption may also occur. Humans and laboratory animals excrete most of the absorbed tetrachloroethene unchanged in expired air, with minor amounts excreted as urinary metabolites. Metabolism is more extensive in mice than in rats and humans. The major metabolite is trichloroacetic acid; minor metabolites include oxalic acid, dichloroacetic acid, ethylene glycol, trichloroacetyl amide, thioethers, and carbon dioxide. Oxidative metabolism (mediated by cytochrome P450) in the liver is the major pathway, leading to the formation of trichloroacetic acid. At higher exposures, this pathway becomes saturated, and a second pathway involving glutathione conjugation increases in significance. This pathway, which is more important in rats than in humans and mice, leads to the formation of S-(1,2,2-trichlorovinyl)-L-cysteine, which can be cleaved in the kidneys to yield cytotoxic and genotoxic metabolites. Reactive intermediates of both pathways can bind covalently to proteins and nucleic acids.

Neat tetrachloroethene was irritating to human and rabbit skin. The liquid caused only minimal irritation to the rabbit eye, and the vapour was irritating to the eyes and respiratory tract of exposed volunteers. In laboratory animals, acute inhalation and oral toxicities were low. In humans, acute accidental inhalation of unmeasured (but presumably high) concentrations of tetrachloroethene has induced CNS depression, dizziness, fatigue, loss of coordination, coma, reversible liver damage, and some deaths. Similar effects were observed in humans following acute ingestion at doses of about 70–90 mg/kg body weight.

Most of the available occupational studies involve people repeatedly exposed, predominantly to tetrachloroethene, but possibly also to other solvents, in the dry cleaning and electronics industries and during metal degreasing operations. Although information on the levels of individual exposures is lacking, average measured tetrachloroethene exposures were typically about 100 mg/m³. In these studies, there was some evidence of toxicity to the CNS and kidney. The neurotoxicity studies showed a common theme of disrupted visual spatial function and CNS cognitive processing of visual information. Although all of the occupational studies of neurotoxicity have limitations, the most informative study found deficits in behavioural tests at a mean exposure level of 83 mg/m³. In the most informative study on kidney effects, there were indications of injury to both tubular and glomerular regions of the kidney at a mean exposure level of 100 mg/m³. There was no clear evidence of liver toxicity in these studies.

On repeated exposure, liver, kidney, and the CNS are the major target organs in laboratory animals. Mice were more sensitive than rats to the liver toxicity of tetrachloroethene.

There is limited evidence that tetrachloroethene is a carcinogen in humans exposed occupationally. Available studies generally lack good information on exposure levels and on exposure to other solvents. The widespread use of tetrachloroethene in the dry cleaning industry did not begin until the 1960s; excess tumour incidence, if occupationally related, could be attributable in part to exposure conditions prior to the widespread use of tetrachloroethene. Where cancer mortality was examined among workers in dry cleaning establishments, elevated mortality was seen in relation to cancer of the oesophagus and cervix. There was some suggestion of an excess in kidney cancers. Three studies reported an excess of non-Hodgkin’s lymphoma, which was not statistically significant; in addition, there may have been multiple solvent exposure. General population and case–control studies gave no convincing evidence for any increased risk of total or specific cancers arising from exposure to tetrachloroethene in drinking-water.

Tetrachloroethene was clearly carcinogenic in laboratory animals. On repeated inhalation, it induced leukaemia in both sexes of F344 rats (in two studies) and malignant kidney tumours in male F344 rats in one study (of two). In inhalation studies, it induced malignant liver tumours in both sexes of B6C3F1 and BDF1 mice and benign Harderian gland tumours in BDF1 male mice. On repeated oral administration, tetrachloroethene induced malignant liver tumours in both sexes of B6C3F1 mice.

Tetrachloroethene has been fairly extensively examined for genotoxicity potential. In vivo, it did not cause chromosomal aberrations in the bone marrow of rats or mice or micronuclei in mouse bone marrow. Sperm abnormalities were not induced in rats or hamsters, but a low-purity grade increased the percentage of abnormal sperm in mice. Tetrachloroethene did not induce dominant lethal mutations in rats. In other assays, it did not damage DNA in the kidneys of rats or lungs of mice; however, transient DNA damage was reported in the liver and kidney of exposed mice. Tetrachloroethene did not induce sex-linked recessive lethal mutations in fruit flies. When tested in vitro, tetrachloroethene did not cause mutations in Ames bacterial assays, chromosome damage, or sister chromatid exchanges in hamster cells, mutations in mouse cells, or UDS in human, rat, or mouse cells. Although a few assays have produced positive results, a weight-of-evidence approach suggests that tetrachloroethene itself does not have significant in vivo genotoxic potential. Mammalian metabolites of tetrachloroethene have induced mutations in Ames assays.

Currently, no mechanisms have been proposed for the leukaemias and benign Harderian gland tumours induced in rats and male mice, respectively. Non-genotoxic mechanisms have been recognized for the formation of kidney tumours in male rats and liver tumours in mice for some chemicals. The available data on mode of action for tetrachloroethene are limited, and the dose–response data related to these recognized mechanisms are not consistent with the dose–response relationships for cancer induction by tetrachloroethene. In the absence of suitable supporting evidence to the contrary, it is concluded that the cancers produced by tetrachloroethene in rodents are of potential relevance to humans.

Some epidemiological studies of women occupationally exposed to tetrachloroethene have shown increased risks of spontaneous abortion; there is insufficient information to draw conclusions in respect of other adverse reproductive outcomes, such as decreased fertility and fetal malformations. Reproductive and developmental studies in rats, mice, and rabbits suggest that tetrachloroethene is fetotoxic at doses that also cause maternal toxicity. Several studies exposing pregnant rats and rabbits found no evidence of structural malformations in the offspring, but one such study in mice reported unspecified soft tissue malformations in the young (at a maternally toxic dose). Limited evidence is suggestive of slight changes in neurochemistry and CNS function in young rats and mice following exposure of the dams during pregnancy.

In occupationally exposed cohorts, the most consistent adverse finding was neurotoxicity; therefore, the most informative study on neurotoxic effects in exposed workers was used to derive a TC. The mean exposure level (83 mg/m³) was taken as a LOAEC. This was converted to an equivalent concentration for continuous exposure (20 mg/m³), and two uncertainty factors of 10 were applied (one to account for interindividual differences, the other because the selected concentration was a LOAEC rather than a NOAEC), to derive a TC of 0.2 mg/m³. For comparative purposes, a similar approach was used for studies reporting nephrotoxicity. The most informative study yielded a mean occupational exposure of 100 mg/m³, which generated a TC of 0.24 mg/m³, a value in good agreement with the TC protective against neurotoxic effects. Available data indicate that liver toxicity would occur only at exposures higher than those that affect the CNS and kidney. A TC for spontaneous abortions was not derived. However, the TC of 0.2 mg/m³ is more than 3 orders of magnitude lower than the exposure concentration that induced mild adverse effects in laboratory animals, and so it was considered to be protective against reproductive toxicity in humans.

The available information on oral exposure was inadequate for derivation of a TDI by the oral route. However, as tetrachloroethene is well absorbed after inhalation or ingestion and there is little evidence of first-pass metabolism, a PBPK model was used to derive a TDI. The model predicted that tetrachloroethene consumed in drinking-water at a dose level of 0.047 mg/kg body weight per day would yield an AUC in plasma similar to that from continuous exposure to tetrachloroethene at 0.2 mg/m³ in inhaled air. This oral figure was rounded to give a TDI of 50 µg/kg body weight.

Tetrachloroethene has induced several types of tumour in rats and mice. Currently, there is no convincing evidence that these tumours arise via modes of action that operate only in rodents, and hence their relevance to humans cannot be dismissed. Therefore, a BMC approach was used, and a BMC and its lower confidence limit (BMCL) were calculated for each animal tumour. Of the tumours observed in experimental animals, hepatocellular adenomas and carcinomas in male mice yield highest predicted risks. The TC derived above, 0.2 mg/m³, corresponds to a cumulative lifetime risk of 0.4 × 10⁻³ when a linear extrapolation is applied to the BMC₁₀ as the point of departure.

Concentrations of tetrachloroethene in the atmosphere or indoor air in Europe and the USA are generally more than an order of magnitude lower than the TC, even in urban areas. In the vicinity of point sources, the observed concentrations also fall below the TC. In buildings where tetrachloroethene is used (notably dry cleaning facilities), concentrations clearly exceeding the TC have been measured. Drinking-water concentrations of tetrachloroethene in different countries in Europe are usually below 10 µg/l, leading to tetrachloroethene doses of below about 0.3 µg/kg body weight per day. This can be compared with the TDI of 50 µg/kg body weight. It should be noted that groundwater concentrations at polluted sites may exceed 1 mg/l.

For terrestrial organisms, the lowest PNEC was 10 µg/kg wet weight in soil. As this was higher than PECs, which ranged from 0.06 to 3.9 µg/kg, it was concluded that tetrachloroethene is unlikely to pose any significant risks to terrestrial organisms. For aquatic organisms, the lowest PNEC was 51 µg/l. PECs ranged from 0.002 to 9.1 µg/l, and so current tetrachloroethene exposures were considered to be a low risk to aquatic organisms. A similar conclusion was reached for sediment-dwelling organisms, where the lowest PNEC was calculated to be 277 mg/kg sediment, compared with a highest calculated PEC of 57 µg/kg. Tetrachlorethene was also considered unlikely to be a risk to microorganisms in sewage treatment processes, with lowest PNEC and highest PEC values of 11.2 mg/l and 16–23 µg/l, respectively. An additional risk assessment was carried out for plants exposed to atmospheric tetrachloroethene. The lowest PNEC was 8.2 µg/m³ air. PECs were generally below this value, but a higher value (36 µg/m³) was measured near a site where tetrachloroethene was produced and processed, leading to the conclusion that there is a need to limit risks of harm to plants from air emissions at such sites.

Tetrachloroethene (CAS No. 127-18-4) is also known as perchloroethylene, tetrachloroethylene, and 1,1,2,2-tetrachloroethene and is often abbreviated to PER or PERC. Its molecular formula is C₂Cl₄, and its relative molar mass is 165.8. Tetrachloroethene’s chemical structure is shown in Figure 1.

Figure 1: Chemical structure of tetrachloroethene.

At room temperature, tetrachloroethene is a clear, colourless liquid with an etheric odour. Selected physical/chemical properties are presented in Table 1.

Table 1: Physical and chemical properties of tetrachloroethene.

^a Data listed in source documents (EC, 2001; de Raat, 2003).

Additional properties are given in the International Chemical Safety Card (ICSC 0076) reproduced in this document.

The conversion factors⁴ for tetrachloroethene in air (at 20 °C and 101.3 kPa) are as follows:

1 ppm = 6.89 mg/m³

1 mg/m³ = 0.145 ppm

To prevent the slow decomposition of tetrachloroethene to trichloroacetyl chloride and phosgene by oxidation, low concentrations of stabilizers (including amines, epoxides, and phenols) are added. Suppliers have reported these as 2,3-epoxypropyl isopropylether (3 g/kg), 2,6-bis(1,1-dimethylethyl)-4-methylphenol (<0.1 g/kg), 2,4-di-tert-butylphenol (<0.05 g/kg), 4-methylmorpholine (<0.1 g/kg), diisopropylamine (<0.5 g/kg), tert-butyl glycidyl ether (<5 g/kg), and tert-amylphenol (<20 mg/kg) (EC, 2001).

Reported impurities include 1,1,1-trichloroethane (<100 mg/kg), carbon tetrachloride (<50 mg/kg), dichloromethane (<2 mg/kg), other chlorinated solvents (<50 mg/kg), trichloroethene (<50 mg/kg), and water (<50 mg/kg) (EC, 2001).

The source documents present several methods for determining tetrachloroethene concentrations in air but give no methods for other environmental media. The compound is always collected by adsorption. Sorbents used are activated charcoal or Tenax, the former being desorbed by elution with organic solvents (e.g. carbon disulfide), the latter by elution of the heated sorbent with an inert gas, followed by condensation. The desorbed material is fractionated by GC. Detection and quantification are based on FID or MS, while the identification of the compound is based on retention time and mass spectra. Several methods are outlined below. The most sensitive are NEN method 2948/2965 and IARC method 12.

ISO method 9486: (E) — A known volume of air is passed through a glass or metal tube packed with activated charcoal. The organic vapours are adsorbed onto the charcoal. The collected vapours are desorbed using a suitable solvent and analysed with a GC equipped with an FID or another suitable detector. This method can be used for the measurement of concentrations of airborne vapours of tetrachloroethene between approximately 1 and 1000 mg/m³ (about 0.2–200 ml/m³) when 10 litres of air are sampled. Organic components that have the same or nearly the same retention time as tetrachloroethene in the GC analysis will interfere. Proper selection of GC columns and programme conditions will minimize interference (ISO, 1991).

NEN method 2947/2964 — Air is drawn through a tube with two sections, both containing activated coconut charcoal to adsorb gaseous tetrachloroethene. The compound is subsequently desorbed with carbon disulfide (containing an internal standard) and is determined by GC, using FID. The method has been validated over a range of 2.5–1600 mg/m³ and has a detection limit of 238 µg/m³ (Dutch Normalisation Institute, 1999a, 2000a).

NEN method 2948/2965 — The sample is collected by adsorption on Tenax (200 mg) and analysed by thermal desorption of volatile components into a GC, using FID. The method has been validated over the range of 0.02–400 mg/m³ and has a detection limit of 0.1 µg/m³ (Dutch Normalisation Institute, 1999b, 2000b).

NEN method 2950 — The sample is collected on an indicator tube and analysed by reading the colour change. The method has been validated over a range of 140–1150 mg/m³. The coefficient of variation was 25% (Dutch Normalisation Institute, 1999c).

NIOSH methods 1003 and 3704 — Method 3704 is specific for tetrachloroethene in exhaled breath and air. Sampling is by gas bag or direct injection, with measurement using a portable GC with a PID. The method has an LOD of about 0.07 mg/m³ and is applicable over the 0.7–700 mg/m³ range (NIOSH, 1998). Method 1003 can be used for various halogenated hydrocarbons. The sampler is a solid sorbent tube containing coconut shell charcoal, and GC with FID is used to measure tetrachloroethene. The working range is 60–13 000 mg/m³, the LOD is 14 mg/m³, and the LOQ is 49 mg/m³ (NIOSH, 2003).

IARC method 5 — Air is drawn through a tube with two sections, both containing activated coconut charcoal to adsorb the gaseous compound. The compound is subsequently desorbed with carbon disulfide (containing an internal standard), followed by GC, using FID. A calibration curve is employed, and a correction curve is applied for desorption efficiency. This method has been validated over a range of 136–4060 mg/m³ using a 3-litre sample. The breakthrough volume is 21 litres at 2750 mg/m³. The detection limit depends on the analyte and lies normally in the useful range (MacKenzie Peers, 1985).

IARC method 12 — Air is drawn through a cartridge containing 1–2 g of Tenax. The cartridge is placed in a heated chamber and purged with an inert gas, which transfers the volatile compound from the cartridge onto a cold trap and subsequently onto a high-resolution (capillary) GC column, which is held at low temperature (e.g. −70 °C). The column temperature is then increased, and the component eluting from the column is identified and quantified by MS. Component identification is normally accomplished by a library search routine, using GC retention times and mass spectral characteristics. The limit of detection is generally in the order of 0.1–1.0 µg/m³ (Riggin, 1985).

BIA method 8690 — The Berufsgenossenschaftliches Institut für Arbeitssicherheit has published a method using Dräger active coal tubes, type B, and GC using FID. The limit of detection is 1.2 mg/m³ for an air volume of 40 litres (Schutz & Coenen, 1989).

The USEPA has also published useful methods for quantifying tetrachloroethene in air, specifically TO-1, TO-3, and TO-14A (USEPA, 1999). OSHA has published a validated method for determination in workplace atmospheres (OSHA, 1999).

For biological monitoring purposes, the concentrations of tetrachloroethene are determined in expired air or blood. Concentrations in expired air can be determined in the same manner as those in ambient air. Tetrachloroethene is removed from blood or tissues by evaporation or by extraction with organic solvents. Evaporated tetrachloroethene can be concentrated with Tenax before analysis with GC/MS or GC with ECD; analysis can also be performed without prior concentration (headspace analysis). The solvent extracts are also analysed by GC/MS or GC with ECD. Several methods are available, including the following.

IARC method 24 — This method can be used for the determination of tetrachloroethene in expired air. The breath sample is dried over calcium sulfate and led through a Tenax GC cartridge. The adsorbed tetrachloroethene is subsequently thermally desorbed and led into a GC/MS. The detection limit of the method is 0.33 µg/m³, and the linear range for the analysis depends mainly on the adsorption breakthrough volume and on the sensitivity of the MS (Pellizari et al., 1985b).

NIOSH method 3704 — Tetrachloroethene in exhaled air can be measured by this method. Sampling is by gas bag or direct injection, with measurement using a portable GC with a PID. The method has an LOD of about 0.07 mg/m³ and is applicable over the 0.7–700 mg/m³ range (NIOSH, 1998).

IARC method 25 — This method is suitable for the determination of tetrachloroethene in blood and tissues. The volatile tetrachloroethene is recovered from a blood sample by warming the sample and passing an inert gas over it. Tissues are first macerated in water, then treated in the same manner as blood. Tetrachloroethene is trapped on a Tenax GC cartridge, then recovered by thermal desorption and analysed by GC/MS. For a 10-ml blood sample, the limit of detection is about 3 ng/ml. Detection limits of about 6 ng/g are typical for 5-g tissue samples. Upper limits are approximately 100 times the lower limits (Pellizari et al., 1985a).

IARC method 27 — Tetrachloroethene concentrations in blood can be determined with this method. The specimen is extracted with n-hexane, and the concentration of tetrachloroethene in the organic phase is determined by GC, using ECD. The limit of detection is 5 µg/l (Pekari & Aitio, 1985).

DFG method 1 — For the determination of tetrachloroethene in blood, an organic matrix is prepared from the sample. The volatile compound is removed from the matrix by increasing the temperature. The headspace of the matrix is then analysed with GC, using ECD. The detection limit is 1.2 µg/l (Angerer & Schaller, 1991).

According to source documents (IARC, 1995; EC, 2001), natural production of tetrachloroethene by temperate, subtropical, and tropical algae and by one red microalga has been reported (Abrahamsson et al., 1994).

Atmospheric releases of tetrachloroethene can occur due to evaporative losses during dry cleaning. Other atmospheric emissions may result during manufacture, from use in metal degreasing, in production of fluorocarbons and other chemicals, in the textile industry, and in miscellaneous solvent-associated applications (ATSDR, 1997). Tetrachloroethene may also be disposed of to land and surface water (TRI, 2004).

Mainly as a result of industrial spillage, tetrachloroethene has been found in air, soil, surface water, seawater, sediments, drinking-water, aquatic organisms, and terrestrial organisms.

Annual production of tetrachloroethene in the USA was estimated to be about 350 000 tonnes in 1981 but had fallen to 169 000 tonnes by the mid-1990s (IARC, 1995). Demand (domestic production plus imports minus exports) for tetrachloroethene in the USA was 126 000 tonnes in 1996 and 143 000 tonnes in 1999 (NTP, 2002). In 1998, the USA produced 160 000 tonnes (of which 18 100 tonnes were exported) and imported 13 600 tonnes, giving a total demand of 155 500 tonnes (HSIA, 1999). The projected demand for 2003 in the USA was 153 000 tonnes (NTP, 2002). In 2004, total demand in the USA was estimated to be about 161 000 tonnes, of which about 16 300 tonnes were imported. An additional 18 600 tonnes were exported (HSIA, 2005). In the EU, total production capacity is 100 000–200 000 tonnes per annum, with an actual reported value in 1994 of 164 000 tonnes (European Chlorinated Solvent Association, personal communication, 1995, cited in EC, 2001). A graph on the Euro Chlor web site indicates that consumption fell from about 230 000 tonnes in 1990 to about 80 000 tonnes in 2004 (Euro Chlor, 2005). In 1994, 56 000 tonnes were exported (European Chlorinated Solvent Association, personal communication, 1996, cited in EC, 2001). For 1979, annual production was estimated to be 50 000–100 000 tonnes in Eastern Europe and about 55 000 tonnes in Japan. Germany, France, Italy, and the United Kingdom are major European producing countries, with Austria, Scandinavia, Spain, Switzerland, and Benelux producing lower amounts (IARC, 1995; EC, 2001; de Raat, 2003).

Currently, tetrachloroethene is produced mainly by oxychlorination, chlorination, and/or dehydrochlorination reactions of hydrocarbons or chlorinated hydrocarbons, most commonly the chlorination of propylene and the oxychlorination of 1,2-dichloroethane (Brooke et al., 1993).

Tetrachloroethene is used mainly as a solvent for dry cleaning and as a chemical intermediate, with additional use for vapour degreasing in metal cleaning. It is also used for processing and finishing in the textile industry, as an extraction solvent, as an anthelmintic, as a heat exchange fluid, in grain fumigation, and in the manufacture of fluorocarbons (EC, 2001; de Raat, 2003). In 1994, use in dry cleaning accounted for 38% of EU production volume, and tetrachloroethene accounted for about 90% of the total solvent used by the European dry cleaning industry (EC, 2001). In the USA in 1998–2000, 50% of tetrachloroethene was used for chemical intermediates, primarily as a basic intermediate in the production of HFC-134a, a popular alternative to CFC refrigerants. It is also used for the synthesis of HCFC-123 and HCFC-124, as well as HFC-125. Dry cleaning accounted for 21–25% of use in the USA, and automotive aerosols (brake cleaners) and metal degreasing each accounted for a further 10%. Tetrachloroethene is also used as an insulating fluid and cooling gas in electrical transformers, paint removers, printing inks, adhesive formulations, paper coatings, and aerosol formulations such as water repellents (HSIA, 1999; NTP, 2001; HSDB, 2003).

In 1990, about 53% of world demand for tetrachloroethene was for dry cleaning (cleaning fluid used by about 75% of all dry cleaners). Approximately 23% was used as a chemical intermediate, principally for the production of Freons, 13% for metal cleaning, and 11% for other uses (Linak et al., 1992; IARC, 1995).

It has been estimated that 80–85% of the tetrachloroethene used annually in the USA is released into the atmosphere (ATSDR, 1997), although the percentage is likely to be lower in the USA in present times, due to improved technology and regulatory restrictions. A major portion of the atmospheric releases is attributed to evaporative losses during dry cleaning. Other atmospheric emissions result from metal degreasing, production of fluorocarbons and other chemicals, use in the textile industry, and miscellaneous solvent-associated applications (ATSDR, 1997). Tetrachloroethene has to a large degree been used in small, geographically scattered, and possibly poorly controlled workplace settings and consequently has, in the USA, become a common contaminant at Superfund waste sites and has been a surface water and groundwater pollutant (NTP, 2001; Aschengrau et al., 2003). In the USA in 2002, the amount of tetrachloroethene estimated to have been released by manufacturing facilities was about 1300 tonnes. This was mainly on site as point source (625 tonnes) or fugitive (400 tonnes) air emissions. Land disposal accounted for about 45 tonnes. Disposal to surface waters was estimated to be about 0.36 tonnes (TRI, 2004).

The authors of the draft EU Risk Assessment Report (the final version of which was released after finalization of this CICAD) carried out a detailed consideration of environmental releases of tetrachloroethene resulting from production, use as a chemical intermediate, use in dry cleaning, use in metal cleaning operations, other uses, and disposal. Three separate scenarios were considered. In a local scenario, emissions to water and air were calculated for point sources and concentrations estimated close to the sources. This local assessment included production, use as an intermediate, dry cleaning, and metal cleaning. In the regional scenario, the EU Technical Guidance Document (ECB, 2003) was followed, with the assumption that the regional environment contained one large-scale production plant and one site using tetrachloroethene as an intermediate, together with 10% of the total EU activity for other uses and 10% of the estimated EU fugitive emissions due to landfill site disposal. The continental region scenario included the remaining releases from production sites, uses, and fugitive emissions from landfill sites. The amounts of tetrachloroethene released to the environment from various sources were summarized in a table, reproduced here as Table 2. These figures were used to calculate PECs (see section 11.2) (EC, 2001).

Table 2: Summary of environmental releases of tetrachloroethene.^a

^a From EC (2001).
^b Regional and continental releases adjusted to 365 days per year.

Tetrachloroethene is distributed between environmental compartments by volatilization, precipitation, and adsorption. It is predominantly released and transported to the atmosphere. Based upon its environmental chemistry, computer models predict that the atmosphere will be the major sink for tetrachloroethene (EC, 2001).

Tests indicate that tetrachloroethene can be adsorbed onto soils of varying organic carbon contents, but amounts adsorbed are negligible, and tetrachloroethene is relatively mobile in groundwater in the absence of any removal processes. Tetrachloroethene can leach rapidly through sandy soil (0.0087% organic matter) into groundwater. In a bank filtration system, tetrachloroethene was rapidly transported to groundwater. It was estimated that only 0.01% was adsorbed to particulate matter (Zoeteman et al., 1980; Wilson et al., 1981; Schwarzenbach et al., 1983). Sorption of non-ionic compounds such as tetrachloroethene depends on the organic carbon content of the soils and sediments and on the nature of the organic matter. When measuring the adsorption coefficient of tetrachloroethene for different soil types, adsorption was highest with anthracite (organic carbon 80.1%) and lowest with lignite (organic carbon 18.5%) (Grathwohl, 1990).

The equilibrium constants were measured for four types of granular media: sandy loam soil, organic top soil, peat moss, and granular activated carbon. Adsorption increased as the carbon content increased, being least with the sandy loam soil (1% organic carbon) and highest with the granular activated carbon (74% organic carbon) (Biswas et al., 1992). Others found similar results upon measuring the adsorption and desorption of tetrachloroethene on a range of soils and clays. Sorption was found to be rapid in all cases and was highest in soils with a high organic carbon content (Doust & Huang, 1992).

Soil organic carbon/water adsorption partition coefficients (K_oc) of 6.5 (over 24 h) and 7.3 (over 72 h) have been reported for tetrachloroethene in fine sand loam soil with aqueous solutions of 4.18–68.2 µg/l (Pignatello, 1990). Bentonite clay adsorbed 22% of tetrachloroethene from a solution containing 1 mg/l after 30 min. No further sorption was noted after this time. Peat moss adsorbed 40% of tetrachloroethene from a 1 mg/l solution in 10 min (Dilling et al., 1975). Reported log K_oc values at 20 ºC vary from 1.6 to 2.7 (Kenaga, 1980; Mabey et al., 1982; Giger et al., 1983; Friesel et al., 1984; Seip et al., 1986; Abdul et al., 1987; Lee et al., 1989; Zytner et al., 1989); a log K_oc of 2.40 (251 l/kg) was taken as representative for tetrachloroethene in the EU risk assessment (EC, 2001). Using this log K_oc value, the following partition coefficients were calculated for tetrachloroethene using the EU Technical Guidance Document (ECB, 2003) method:

Tetrachloroethene released to surface waters rapidly volatilizes to the atmosphere, at a rate dependent upon the degree of mixing in the water system. Removal is more rapid from water systems with a high degree of mixing, which depends upon water movements and wind speed. The evaporation half-life of tetrachloroethene from field measurements and theoretical considerations is in the order of 1–10 days in rivers and 10–30 days in lakes and ponds (ECETOC, 1999; EC, 2001). Using representative oxygen aeration rates for various bodies of water, half-lives for the evaporation of tetrachloroethene have been calculated as 5–12 days (pond), 3 h – 7 days (river), and 3.6–14 days (lake) (Lyman et al., 1981). In a study in which the evaporation of tetrachloroethene from a 1 mg/l solution was measured at ambient temperatures, the half-life was between 24 and 28 min with constant stirring at 200 rpm and was about 90 min with stirring for 15 s every 5 min (Dilling et al., 1975). In later experiments using the same technique, the measured half-life for evaporation was 20–27 min (Dilling, 1977).

Volatilization from water was reported as 0.18 µg/cm² per hour (Wilson et al., 1981). The half-life for evaporation from water was reported as 3.2 min under stirring conditions (Chiou et al., 1980). Volatilization of tetrachloroethene was measured in a model mesocosm (a tank containing 13 m³ of seawater plus associated planktonic and microbial communities) that was mixed for 2 h, 4 times a day. The measured volatilization half-lives were 11 days in winter, 25 days in spring, and 14 days in summer (Wakeham et al., 1983).

Volatilization of tetrachloroethene from dry soil is likely to be rapid due to its high vapour pressure and low adsorption to soil. Volatilization from a sandy soil was reported as 0.103 µg/cm² per hour (Wilson et al., 1981).

Tetrachloroethene has been detected in rainwater and has the potential to dissolve in atmospheric water droplets and be deposited by rainout. Trichloroacetic acid formed by the photodegradation of tetrachloroethene may be rained out, together with the hydrogen chloride formed. Trichloroacetic acid has been found in rainwater samples, soil samples, and spruce needles (EC, 2001).

A model (FUGMOD [OECD workshop] Mackay Level I) has been used to calculate the distribution of tetrachloroethene in the environment as follows: air (99.69%), water (0.23%), soil (0.07%), sediment (<0.01%), and biota (<0.01%). Using another model (FUGMOD [OECD workshop] Mackay Level III), the distribution of tetrachloroethene in the environment was calculated as follows: air (76.39%), water (23.32%), soil (0.06%), and sediment (0.23%). A release rate of 1000 kg/h was used in the Level III model. It was assumed that 90% of releases were to air and 10% to water (EC, 2001). The fate of tetrachloroethene in a wastewater treatment plant, as estimated with EUSES,⁵ is 91.2% to air, 6.54% to water, and 2.2% to sludge, with zero degradation (EC, 2001).

BCFs of approximately 40–50 have been reported for aquatic species with tetrachloroethene. For bluegill (Lepomis macrochirus) exposed to 3.43 µg/l for 21 days at 16 ºC, a BCF of 49 was reported (Barrows et al., 1980). For rainbow trout (Oncorhynchus mykiss), a BCF of 40 was reported (Neely et al., 1974). Based on these data, no significant bioaccumulation of tetrachloroethene in fish is expected. BCF figures of 312 and 101 have been calculated for the marine microalgae Heterosigma akashiwo (dinoflagellate) and Skeletonema costatum (diatom), respectively (Wang et al., 1996).

The octanol–water partition coefficient (log K_ow) value for tetrachloroethene is below 3, indicating a low potential for bioaccumulation. The BCF for fish is calculated as 28.2 by the EU Technical Guidance Document method (the current version of which is ECB, 2003), and this value was used in the EU risk assessment (EC, 2001).

5.4.1 Atmospheric degradation

Tetrachloroethene will react in the atmosphere with a number of photochemically produced species. The major removal process for tetrachloroethene from the atmosphere results from reaction with hydroxyl radicals. Atkinson (1985) reviewed the available data for this process and recommended the following value for the second-order reaction rate constant:

k_OH = 9.64 × 10⁻¹² exp(−1209/T) cm³/s per molecule

This gives a value for k_OH of 1.23 × 10⁻¹³ cm³/s per molecule at T = 277 K (or 4 °C).

The atmospheric lifetime of tetrachloroethene due to reaction with hydroxyl radicals has been estimated to be around 0.43 years (WMO, 1991). The EU Technical Guidance Document (ECB, 2003) recommends a value for the atmospheric hydroxyl radical concentration of 5 × 10⁵ molecules/cm³. Using this concentration, a half-life of about 3.2 months (lifetime = 4.6 months) is estimated for the reaction. The estimated half-life is long enough to allow transport of tetrachloroethene from the point of emission (ECETOC, 1999; EC, 2001).

The reaction with atmospheric chlorine atoms is thought to be the next most important atmospheric degradation mechanism for tetrachloroethene. The second-order rate constant for the reaction has been quoted (Nicovich et al., 1996) as:

k_Cl = 4.0 × 10⁻¹¹ cm³/s per molecule

The actual concentration of chlorine radicals in the atmosphere is unknown. Concentrations of about 1000 molecules/cm³ have been suggested, but a study by Sidebottom & Franklin (1996) suggests that the actual concentration in the troposphere is generally close to zero and is at most 500 molecules/cm³. The half-life for the reaction between chlorine radicals and tetrachloroethene is estimated (for [Cl·] = 1000 or 500 molecules/ cm³) as 6–12 months (lifetime 9–17 months) (ECETOC, 1999; EC, 2001).

The overall lifetime for these two major processes combined is thought to be around 3 months (ECETOC, 1999), although the exact contribution of the reaction with chlorine atoms to the overall degradation of tetrachloroethene is uncertain.

Class & Ballschmiter (1987) measured the concentration of tetrachloroethene in the atmosphere at sites remote from any anthropogenic sources in the Northern and Southern hemispheres. The lifetime in the Northern Hemisphere was estimated as 0.46 years (5–6 months) and in the Southern Hemisphere as 0.18 years (2 months). These lifetimes were calculated based on estimated release rates and measured levels.

The lifetime for removal of tetrachloroethene by gas-phase photolysis has been calculated to be about 3 years in the troposphere. Direct photolysis is therefore thought to be of negligible importance compared with other tropospheric removal mechanisms (ECETOC, 1999; EC, 2001). Reactions of tetrachloroethene with other atmospheric species, such as ozone (k <3 × 10⁻²⁰ cm³/s per molecule [Atkinson & Carter, 1984]; k <2 × 10⁻²³ cm³/s per molecule [Franklin, 1994]), oxygen atoms (k(O³P) = 1.6 × 10⁻¹⁴ cm³/s per molecule, k(O¹D) <5 × 10⁻¹⁰ cm³/s per molecule [Franklin, 1994]), nitrate radicals (k <1 × 10⁻¹⁶ cm³/s per molecule [Atkinson et al., 1992]; k <5.2 × 10⁻¹⁷ cm³/s per molecule [Franklin, 1994]), and hydroperoxy radicals (k <1 × 10⁻¹⁷ cm³/s per molecule [Franklin, 1994]), have been reported but are thought to be insignificant atmospheric degradation processes (estimated atmospheric lifetimes for these processes range from >5 to >1500 years) for tetrachloroethene (Franklin, 1994).

For the EU risk assessment, the reaction rates with hydroxyl and with chlorine radicals were combined to give an overall half-life of 96 days. This was used in the EUSES calculations (entered as the reaction rate with hydroxyl radicals that would give a half-life of 96 days) (EC, 2001).

In laboratory studies on the photochemical degradation of tetrachloroethene in air, the main products identified were phosgene, trichloroacetyl chloride, hydrogen chloride, carbon dioxide, and carbon monoxide, but other products, such as carbon tetrachloride, dichloroacetyl chloride, and chloroform, have also been detected (ECETOC, 1999). When tetrachloroethene (30 mg/m³) was irradiated with air containing nitrogen dioxide in a smog chamber for 140 min, around 7% of the tetrachloroethene reacted, forming carbon monoxide (0.31 mg/m³), ozone (0.27 mg/m³), hydrogen chloride (0.64 mg/m³), and phosgene (0.49 mg/m³). Trichloroacetyl chloride was also identified (Gay et al., 1976). After 7 days’ illumination with simulated sunlight, the product yields were around 70–85% phosgene and 8% carbon tetrachloride. The concentration of carbon tetrachloride continued to increase long after the tetrachloroethene had disappeared, indicating that it was formed from an intermediate substance, probably trichloroacetyl chloride (Singh et al., 1975). Others analysed the products formed during the reaction of tetrachloroethene with hydroxyl radicals. During the 2-h experiment, only about 10% of the tetrachloroethene reacted. The main products formed were phosgene (product yield 47–52%) and trichloroacetyl chloride (product yield 39–41%). When the experiment was repeated in the presence of a chlorine atom scavenger (ethane), there was a marked decrease in the amount of trichloroacetyl chloride formed (product yield <15%), indicating that trichloroacetyl chloride is formed by chlorine atom attack on the tetrachloroethene (Tuazon et al., 1988).

In a detailed study of the products from hydroxyl radical–initiated reactions of tetrachloroethene, the compound (181 mg/m³) was irradiated for 12 h in air (20% relative humidity) using hydrogen peroxide as the source of hydroxyl radicals. All of the tetrachloroethene reacted, producing trichloroacetyl chloride (46 mg/m³; 23.2% yield), carbon dioxide (9.7 mg/m³; 20% yield), carbon monoxide (5.9 mg/m³; 18% yield), phosgene (7.8 mg/m³; 7% yield), and carbon tetrachloride (126 µg/m³; 0.07% yield). The shapes of the degradation curves were consistent with two competing reactions — i.e. addition of hydroxyl radical to form phosgene and addition of chlorine atoms to form trichloroacetyl chloride, the latter becoming more prevalent with time as the chlorine atom concentration in the reaction chamber increased. Further experiments showed that carbon tetrachloride was formed by photolysis of trichloroacetyl chloride, and not directly from tetrachloroethene. The conversion of trichloroacetyl chloride to carbon tetrachloride was estimated to be about 0.1%, based on 24-h illumination (Itoh et al., 1994).

It is clear that two competing reactions occur in closed laboratory studies. The main products formed of relevance to the environment are phosgene, trichloroacetyl chloride, and carbon tetrachloride. Phosgene is derived from hydroxyl radical addition to tetrachloroethene. Trichloroacetyl chloride is derived from chlorine atom addition to tetrachloroethene, and carbon tetrachloride is formed as a result of further degradation of trichloroacetyl chloride. The main reaction pathways for the chlorine atom addition are shown below (EC, 2001; based on Sidebottom & Franklin, 1996 and Itoh et al., 1994):

Chlorine atom addition:

Cl₂C=CCl₂ + Cl· --> CCl₃CCl₂·

CCl₃CCl₂· + O₂ --> CCl₃CCl₂O₂·

CCl₃CCl₂O₂· + NO --> CCl₃CCl₂O·

CCl₃CCl₂O· (85%) --> CCl₃COCl + Cl·

CCl₃CCl₂O· (15%) --> COCl₂ + CCl₃· (--> COCl₂)

CCl₃COCl + hv --> Cl· + CCl₃CO· --> CO + CCl₃· --> --> COCl₂

CCl₃COCl (0.1%) + hv --> CCl₄ + CO

It has been suggested that trichloroacetyl chloride is a major atmospheric degradation product of tetrachloroethene. Trichloroacetyl chloride can hydrolyse to form trichloroacetic acid, which can be washed out of the atmosphere (Reimann et al., 1996). Chloroacetic acids are toxic to many plants, and some have been used as herbicides. In laboratory experiments, seemingly large yields of trichloroacetyl chloride and carbon tetrachloride are formed; in the environment, however, this behaviour is likely to be modified. In laboratory studies, chlorine atoms formed during hydroxyl radical–initiated degradation of tetrachloroethene can build up in the test system, and so the chlorine addition pathway can effectively compete with hydroxyl radical addition, resulting in high yields of trichloroacetyl chloride (and subsequently carbon tetrachloride). In the environment, however, there are many other chemical species (e.g. hydrocarbons) that are capable of scavenging the reactive chlorine atoms, and so the proportion of tetrachloroethene reacting via this pathway will be much diminished in the environment. This has been demonstrated in laboratory studies in which a chlorine atom scavenger was added (EC, 2001).

The lifetime of tetrachloroethene in the troposphere is such that the fraction of emitted tetrachloroethene that enters the stratosphere is low (about 1% of atmospheric emissions). In the stratosphere, tetrachloroethene will be degraded by reaction with hydroxyl radicals. It may also undergo photolysis (ECETOC, 1999).

5.4.2 Aquatic degradation and biodegradation

5.4.2.1 Abiotic degradation

Degradation of tetrachloroethene in water by hydrolysis is very slow, with reported half-lives in the order of years (ECETOC, 1999). Tetrachloroethene may be removed from aquatic systems by photochemical reactions involving free radicals or electronically excited molecular species. These reactions are likely to compete with volatilization only in still, sunlit waters, where volatilization is limited by the available surface area for evaporation (ECETOC, 1999). Reductive pathways involving transition metals or their organic complexes may be significant in the presence of soils or sediments (ECETOC, 1999). No further information on these processes has been found (EC, 2001).

5.4.2.2 Aerobic biodegradation

No degradation of tetrachloroethene was detected in a 190-h incubation with a mixed culture of methane-utilizing bacteria (Fogel et al., 1986) or when using a culture of the ammonium-oxidizing bacterium Nitrosomonas europaea (Vannelli et al., 1990). No degradation was observed in a modified shake flask, a closed bottle biodegradation test, or a river die-away study. A 21-day acclimation period was used in the closed bottle test and river die-away study (Mudder, 1982). No degradation of tetrachloroethene (at 9–74 µg/l initial concentrations) was observed in cultures containing a bacterial inoculum obtained from primary sewage effluent and incubated in the dark for 25 weeks at 20 ºC (Bouwer et al., 1981). No degradation was observed when a sterile salt solution containing tetrachloroethene at 10–30 µg/l was applied continuously to an up-flow glass column containing inert material, seeded with primary sewage and operated at 22–23 ºC under aerobic conditions for 2 years (Bouwer & McCarty, 1982). No degradation of tetrachloroethene was observed during the infiltration of river water to groundwater. Samples were taken from groundwater sites near a contaminated river over 1 year, and conditions were predominately aerobic (Schwarzenbach et al., 1983).

Although tetrachloroethene is persistent under aerobic conditions, 60–90% decreases in concentrations have been reported in studies using aerobic soil columns (Phelps et al., 1991; Enzien et al., 1994); this was possibly due to the presence of anaerobic niches within the column beds, although no specific evidence of anaerobic biodegradation was found (EC, 2001). One study, using a static culture, BOD-based flask method with an inoculum of domestic sewage sludge, indicated that tetrachloroethene may undergo primary degradation, the rate increasing with adaptation of the microorganisms. Losses of tetrachloroethene (initial concentration 5 mg/l) in four consecutive 7-day periods were 45%, 54%, 69%, and 87%, respectively. Volatilization accounted for the remainder (Tabak et al., 1981).

5.4.2.3 Anaerobic biodegradation

Tetrachloroethene undergoes anaerobic degradation by reductive dechlorination. Reported degradation products are trichloroethene, dichloroethene, vinyl chloride, ethene, and ethane and vary with the experimental conditions used. The inocula used in the majority of experiments were adapted, and degradation of tetrachloroethene was usually observed at elevated temperatures and in the presence of nutrients. Several methanogenic organisms were found to be capable of dechlorinating tetrachloroethene. The redox potential is important in determining the level of dechlorination. Anaerobic dechlorination takes place under methane- and sulfate-reducing conditions. For dechlorination to occur, an electron donor is usually required — for example, acetate or lactate (EC, 2001).

Reductive dechlorination of tetrachloroethene was studied in a fixed-bed column containing anaerobic river sediment and anaerobic granular sludge and continuously percolated with an anaerobic mineral medium. Reducing conditions were maintained in the column by the presence of sodium sulfide (10 mg/l), and lactate was used as an electron donor (1 mmol/l). The flow rate through the column was 15 ml/h at 20 ºC in the dark. After adaptation, tetrachloroethene was dechlorinated stepwise via trichloroethene, cis-1,2-dichloroethene, and vinyl chloride to ethene. Ethene was then reduced to ethane within 24 h. The conversion of tetrachloroethene to ethane was 95–98% during a 24-h period at an initial concentration of 1.5 mg/l. Lowering the column temperature to 10 ºC caused an initial decrease in the conversion to ethane; after 2 weeks at the lower temperature, however, only ethane and ethene were detected in the effluent (De Bruin et al., 1992).

Reductive dechlorination of tetrachloroethene was also studied in a methanogenic culture allowed to adapt for 115 days. Analysis of samples taken every 2 days between days 115 and 135 showed that freshly added tetrachloroethene was degraded to vinyl chloride (about two thirds) and ethane (about one third), with traces of trichloroethene and dichloroethene. After 170 days, tetrachloroethene was degraded to ethene (80%) and vinyl chloride (20%) within 2 days, even though methanogenic activity (as measured by methane production) had declined. When longer periods were allowed between additions of tetrachloroethene, it was found that the amount of vinyl chloride remaining after 4 days was <1% of the total products (DiStefano et al., 1991). In methanol- and hydrogen-fed mixed anaerobic cultures, tetrachloroethene was biodegraded to ethene (80%) and vinyl chloride (20%), with traces of trichloroethene and dichloroethene (DiStefano et al., 1992).

Testing of several anaerobic bacteria, including four strains of acetate-utilizing methanogens (Methanosarcria sp., Methanosarcria mazei, Methanosarcria acetivorans, and Methanothrix sp.), Desulfovibrio desulfuricans, Clostridium pasteurianium, Clostridium butyricum, and a pure-culture dehalogenator (Desulfomonile tiedjei; DCB-1), found that Methanosarcria sp., Methanosarcria mazei cultures, and DCB-1 could degrade tetrachloroethene to trichloroethene. The process by which methanogens dechlorinate tetrachloroethene is a co-metabolic process and appears to be dependent on the formation of methane from the carbon source (Fathepure et al., 1987; Fathepure & Boyd, 1988).

Fathepure & Tiedje (1994) studied the reductive dechlorination of tetrachloroethene at 35 ºC in a continuously fed, up-flow biofilm reactor, inoculated with an enriched culture containing anaerobic Desulfomonile tiedjei bacteria. After steady-state conditions had being achieved (4 months), tetrachloroethene was added, and the column was left to acclimatize for 3–4 weeks. Degradation rates of between 78% and 86% were measured for tetrachloroethene concentrations of 0.26–1.0 mg/l, and trichloroethene and dichloroethene were found as degradation products.

Freedman & Gosset (1989) studied the anaerobic degradation of tetrachloroethene with microorganisms from a wastewater treatment plant. The organisms were adapted prior to use by anaerobic incubation at 35 ºC with aqueous tetrachloroethene. When the added tetrachloroethene had been degraded, a sample of the culture was removed from each bottle and replaced by fresh medium and tetrachloroethene. Operating in this semi-continuous way, sixth-generation cultures were obtained. Redox conditions were maintained by the presence of Fe²⁺ ions. Ethene was reported as the main degradation product formed, with traces of trichloroethene and dichloroethene. Reductive dechlorination was found to occur only when methanol was used as co-metabolite. The addition of hydrogen to dechlorinating microcosms increased the dechlorination rate by about 500 times after 200 days.

Holliger et al. (1993) isolated a bacterium capable of growing on tetrachloroethene from an inoculum derived from anaerobic sediment and anaerobic granular sludge. Hydrogen or formate was necessary as electron donor for growth. Tetrachloroethene was degraded in the anaerobic packed-bed column from which the bacterium was derived. The main degradation product was ethane, and traces of cis-1,2-dichloroethene, trichloroethene, vinyl chloride, and ethene were detected.

Kästner (1991) found that reductive dechlorination of tetrachloroethene to cis-1,2-dichloroethene occurred upon the transfer from aerobic to anaerobic conditions. The cultures used contained aerobic isolates in a co-culture with a Bacillus sp. and a Desulfotomaculum sp. under conditions of limited oxygen supply. Degradation of tetrachloroethene occurred only in cultures that were initially aerobic but became anaerobic after a few days’ incubation. No degradation was observed in cultures that were anaerobic from the start. Transformation of tetrachloroethene required a decrease in redox potential of the system caused by sulfide formation from degradation of sulfur compounds present in the system.

Liang & Grbi-Gali (1993) studied the degradation of tetrachloroethene under methanogenic conditions using aquifer material obtained from contaminated sites. They incubated microcosms containing inocula and anaerobic mineral medium at 35 ºC. Reducing conditions were maintained by the presence of Fe²⁺ ions. Trichloroethene, trans-1,2-dichloroethene, and vinyl chloride were detected as the degradation products.

Trichloroethene and cis-1,2-dichloroethene were detected as degradation products under both sulfate-reducing and methanogenic conditions in microcosms packed with soil containing tetrachloroethene at 11.5 mg/kg (Pavlostathis & Zhuang, 1993). In further studies, it was shown that reductive dechlorination of tetrachloroethene (1 mg/l) in a methanogenic culture derived from contaminated soil was influenced by temperature (peaked at 35 ºC, then decreased above 45 ºC) and acidity (maximal at pH 7) (Zhuang & Pavlostathis, 1995).

When microcosms prepared from aquifer solids and distilled water spiked with tetrachloroethene at 3 µmol/l were incubated in the dark, anaerobic biotransformation to trichloroethene and cis-1,2-dichloroethene was complete within 7 days (Ninomiya et al., 1994). In a continuous-flow fixed-film methanogenic column with a 2-day retention time, the tetrachloroethene concentration was anaerobically reduced from 20.5 mg/l to 4.4 µg/l (99.98% reduction). Trichloroethene, dichloroethene, and (mainly) vinyl chloride were identified as the degradation products. Mineralization to carbon dioxide was detected by tracer studies (Vogel & McCarty, 1985). In a deoxygenated anaerobic medium seeded with a methanogenic culture, tetrachloroethene (0.2 mg/l) was completely degraded (mainly to trichloroethene) within 8 weeks (Bouwer & McCarty, 1983). When studied using two continuous columns (in series) seeded by a primary sewage effluent and a methanogenic bacterial inoculum, respectively, steady-state removal of tetrachloroethene was 86% after a 10-week acclimation period. The columns were operated in the dark at 22–23 ºC for 19 months (Bouwer & McCarty, 1983).

Suflita et al. (1988) compared biodegradation under methanogenic and sulfate-reducing conditions. In both instances, tetrachloroethene was dechlorinated by sequential reduction reactions to form mainly trichloroethene, dichloroethene, and vinyl chloride. Degradation was found to occur faster in the methanogenic cultures. Gibson & Sewell (1992) showed that dehalogenation occurs in the presence of a suitable electron donor, such as lactate and acetate. Trichloroethene and dichloroethene were detected as degradation products.

Reductive anaerobic dechlorination of tetrachloroethene can lead to the formation of vinyl chloride in groundwater, and the vinyl chloride detected in groundwater may be the result of tetrachloroethene degradation. However, several studies show that the process of dechlorination can continue under anaerobic conditions, with the end-products being ethene and ethane. Several other chlorinated solvents such as trichloroethene are also likely to break down in the environment to vinyl chloride. Therefore, while vinyl chloride may be formed from tetrachloroethene in groundwater, it is not possible to quantify the extent to which this process contributes to the levels of vinyl chloride found. Possible risks from this were not assessed in the EU risk assessment (EC, 2001). Some concerns have been expressed over the possible production of trichloroacetic acid in surface waters from the breakdown of tetrachloroethene. From the studies presented here, tetrachloroethene appears to be relatively stable to degradation in surface waters, with volatilization being the main removal process. In anaerobic environments, reductive dechlorination appears to be the most likely degradation pathway, producing trichloroethene, vinyl chloride, and ultimately ethene and ethane. No evidence for the formation of trichloroacetic acid from tetrachloroethene in water was found in the literature (EC, 2001).

The reactivity of tetrachloroethene in the troposphere has been reported as being low enough so as not to contribute significantly to tropospheric ozone formation and the related "photochemical smog". The photochemical ozone creation potential of tetrachloroethene in the troposphere is estimated as 1, expressed relative to 100 for ethene (a substance thought to be important in photochemical ozone production) (Derwent & Jenkin, 1990).

The small fraction of tetrachloroethene that enters the stratosphere (about 1% of atmospheric emissions) from the troposphere will be degraded by reaction with hydroxyl radicals, but some might undergo photolysis to yield products that may lead to ozone depletion. The actual impact is likely to be negligible compared with that of other ozone-depleting chemicals, such as CFCs and methyl chloroform (1,1,1-trichloroethane). The estimated stratospheric chlorine loading potential of tetrachloroethene is less than 0.01. Some of the degradation products of tetrachloroethene formed in the troposphere may enter the stratosphere and contribute to ozone depletion. The contribution that these products might make to ozone depletion has not been quantitatively assessed (ECETOC, 1999).

A source document (EC, 2001) has concluded that the reactivity of tetrachloroethene in the troposphere (the half-life is around 3–5 months) is such that it is not thought to contribute significantly to tropospheric ozone formation. Gas-phase photolysis and rainout are thought to be of negligible importance in the removal of tetrachloroethene from the troposphere. The lifetime of tetrachloroethene in the troposphere is such that the amount entering the stratosphere is low. Studies into stratospheric ozone depletion mention that tetrachloroethene is a possible ozone depleter, although its potential is significantly lower than that of other ozone-depleting chemicals. Degradation products of tetrachloroethene in the troposphere may enter the stratosphere; of these, carbon tetrachloride is a known ozone depleter. The amounts of carbon tetrachloride entering the stratosphere due to tetrachloroethene degradation are thought to be negligible when compared with other sources of carbon tetrachloride emissions. No data were found quantifying the contribution that tetrachloroethene makes to ozone depletion, either directly or indirectly via its degradation products. An expert working group on ozone depletion (WMO, 1991) considered that tetrachloroethene makes a negligible contribution to ozone depletion relative to other ozone-depleting chemicals, such as CFCs, HCFCs, carbon tetrachloride, and 1,1,1-trichloroethane. Tetrachloroethene is not expected to contribute significantly to global warming (EC, 2001).

6.1.1 Ambient air

The draft EU Risk Assessment Report concluded that the majority of measured levels are below 10 µg/m³ and, indeed, mostly below 1 µg/m³ (EC, 2001). A review of measured tetrachloroethene concentrations in Germany in 1988 found that concentrations in rural areas ranged between 0.5 and 2 µg/m³, and concentrations in urban areas ranged between 2 and 15 µg/m³ (BUA, 1993). The ATSDR in the USA concluded that background levels lie generally in the lower ppt range (1 ppt = 6.89 ng/m³) in rural and remote areas; values in the higher ppt and lower ppb range (1 ppb = 6.89 µg/m³) are found in urban and industrial areas and areas near point sources of pollution (ATSDR, 1997). The most recent modelling of exposure information for the USA by the USEPA is based on 1996 information and is published as part of the National Air Toxics Assessment. The results show that, for 95% of counties in the USA, the atmospheric level was 0.29 µg/m³ or below. The highest county level was 1.39 µg/m³ (USEPA, 2002). Earlier surveys of the air in nine cities in the USA showed concentrations between 0.2 and 52 µg/m³, with averages ranging from 2 to 4 µg/m³ (IPCS, 1984).

Global background concentrations measured in 1989 were 0.09 µg/m³ for the Northern Hemisphere and 0.02 µg/m³ for the Southern Hemisphere (Koppmann et al., 1993).

A median outdoor level of 2 µg/m³ was measured in a study in Holland (Lebret et al., 1986). In a study performed between October 1987 and September 1988 in an "urban canyon" in the centre of Turin, Italy (120 samples taken during 10 consecutive days, 24 h each, during approximately 1 year; 31 measurements during winter and 28 during summer), it was found that contamination of air was higher in winter than in summer, the mean atmospheric concentrations being either 8.70 µg/m³ and 4.75 µg/m³, respectively (Gilli et al., 1990a), or 13.6 µg/m³ and 4.75 µg/m³, respectively (Gilli et al., 1990b). It was found that the indoor/outdoor concentration ratio was higher in winter that in summer, median concentration ratios being 2.15 and 1.38, respectively (Gilli et al., 1990a,b).

Monitoring in the Rhine valley in 1996 showed a maximum value from one site (Freiburg) of 2.9 µg/m³; the highest 98th-percentile value for a sampling site was 1.8 µg/m³ at Karlsruhe (UMEG, 1997). Data from 1998 for the Nordrhein–Westfalen region, including the industrialized area between Dortmund and Cologne, showed a peak value of 2.4 µg/m³, with average values generally below 0.5 µg/m³ (LUA, 1999). Bruckmann et al. (1989) sampled air from 12 sites in Hamburg between April 1986 and April 1987 and characterized activities near the sites. Significantly higher concentrations were found at three sites: near a chemical laundry (dry cleaning), a rubber factory, and an industrial area with metalworking industry and small chemical factories. The highest yearly average concentration was 71 µg/m³, while the overall average concentration was 3.5 µg/m³ (Bruckmann et al., 1989). Surveys of the air in 14 cities in Germany showed average concentrations between 1.7 and 6.1 µg/m³ (IPCS, 1984).

Monitoring air concentrations near a tetrachloroethene production and processing facility over a 4-week period when both areas of activity were operating (24-h samples taken for 28 days) found an average daily mean concentration of 36 µg/m³ (EC, 2001).

In Germany, annual mean levels of 41 µg/m³ and 69 µg/m³ were detected downwind of a chemical laundry and a rubber factory, respectively (IPCS, 1984).

6.1.2 Indoor air

The median level of tetrachloroethene in about 400 Dutch homes was 4 µg/m³, while maximum levels varied between 49 and 205 µg/m³ (Lebret et al., 1986).

Levels can be much higher in buildings housing dry cleaning facilities. For example, sampling (over 100 samples) of air in six residential apartments in two buildings where dry cleaning was carried out on the ground floor revealed tetrachloroethene concentrations ranging from 50 to 6100 µg/m³, with means ranging from 358 to 2408 µg/m³. One month after the dry cleaning facilities ceased operating, atmospheric concentrations in the apartment air had declined substantially but still ranged from 10 to 800 µg/m³ (Schreiber et al., 2002).

6.1.3 Drinking-water

Tetrachloroethene was reported in drinking-water in Germany at 1.3 µg/l (Lahl et al., 1981). More recently, figures for samples of drinking-water in Germany were given as <0.001 µg/l (51% of samples), 0.001–0.5 µg/l (40% of samples), and >0.5 µg/l (9% of samples) (Bauer, 1991). Levels in Finnish samples ranged up to 0.05 µg/l (Reunanen & Kroneld, 1982; Kroneld, 1986). Tetrachloroethene was detected in 454 of 2682 samples taken by 29 water companies in the United Kingdom. The detection limit ranged from 0.1 to 1 µg/l, and the maximum concentration detected was 12.2 µg/l. The range and mean were not given (personal communication to United Kingdom Environment Agency from United Kingdom Drinking Water Inspectorate, 1995, cited in EC, 2001).

Industrial disposal is the typical likely source of drinking-water contamination with tetrachloroethene. However, in the late 1960s through the early 1980s, tetrachloroethene leached into the drinking-water supplies of Cape Cod, Massachusetts, in the USA from an inner vinyl liner that was present in certain cement pipes (in which a slurry of a vinyl plastic and tetrachloroethene was used to coat the inside of the pipe before shipment), affecting 1000 km of the pipes. Typical levels in the water of affected towns in the region ranged from 1.60 to 7.75 mg/l in low-flow locations and from 1.5 to 80 µg/l in medium- and high-flow locations (Demond, 1982; Aschengrau et al., 2003).

6.1.4 Surface water

Tetrachloroethene has been measured in surface (river) waters in Germany, Finland, the Netherlands, Italy, France, Switzerland, the United Kingdom, and the USA. Concentrations ranged from 0.01 to 168 µg/l, with levels typically below 5 µg/l (Reunanen & Kroneld, 1982; Ahel et al., 1984; Hellmann, 1984; Staples et al., 1985; Aggazzotti & Predieri, 1986; Kroneld, 1986; Van de Meent et al., 1986; Marchand et al., 1988; Van der Graff, 1988; Bohlen et al., 1989; Malle, 1990; RIVM, 1993; EC, 2001).

Analysis of coastal and estuarine waters of Germany, the United Kingdom, Sweden, France, Greece, and the Mediterranean revealed that tetrachloroethene concentrations were below 3 µg/l (Hellmann, 1984; Fytianos et al., 1985; Marchand et al., 1986, 1988; Van de Meent et al., 1986; Abrahamsson et al., 1989; Hurford et al., 1989; Dawes & Waldock, 1994; England and Wales National Rivers Authority, personal communication, 1995, cited in EC, 2001).

Summarizing several other reviews (e.g. IPCS, 1984; IARC, 1995), de Raat (2003) presented average and maximum concentrations in seawater of 0.012 and 2.6 µg/l, respectively. Surface water from the Atlantic Ocean was said to contain 0.0002–0.0008 µg/l. The highest concentration found in the surface waters of Lake St. Clair (Canada/Michigan, USA) was 0.47 µg/l (de Raat, 2003).

Rainwater samples in Germany, the Netherlands, Switzerland, the United Kingdom, and the USA contained tetrachloroethene at <0.005–0.15 µg/l; the highest figure was found in an industrial area (Kawamura & Kaplan, 1983; Atri, 1985; Van de Meent et al., 1986; Czuczwa et al., 1988; Kubin et al., 1989; Renner et al., 1990).

6.1.5 Groundwater

The EU Risk Assessment Report concluded that groundwater concentrations of tetrachloroethene vary widely. Although groundwater concentrations are generally higher than concentrations in surface water, this could reflect the fact that groundwater measurements tend to be taken where a problem (e.g. a spill) is thought to exist (EC, 2001). Groundwater levels are usually below 10 µg/l (Fielding et al., 1981; Trowborst, 1981; Fahmi, 1984; Aggazzotti & Predieri, 1986; Goodenkauf & Atkinson, 1986; Sagunski et al., 1987; Heil et al., 1989; Bauer, 1991; England and Wales National Rivers Authority, personal communication, 1995, cited in EC, 2001), but concentrations as high as 1300 µg/l have been reported for a contaminated site (Leschber et al., 1990).

One source document reported that tetrachloroethene has been measured at 0.01–46 µg/l groundwater in western Europe, and the maximum concentration reported in groundwater in the Netherlands was 22 µg/l (de Raat, 2003).

6.1.6 Sediment and soil

Tetrachloroethene has been measured in sediment samples at 1–50 µg/kg wet weight in Germany (Alberti, 1989) and at <5 µg/kg wet weight in the USA (Staples et al., 1985). One source document reports a maximum concentration of 4.8 mg/m³ in sediments (de Raat, 2003).

Samples of soil air taken in Germany contained tetrachloroethene at 2.1–4.5 µg/m³ (Frank et al., 1989).

6.1.7 Sewage and municipal wastewater

The EU risk assessment reports tetrachloroethene concentrations in municipal wastewaters in Germany, the United Kingdom, France, Switzerland, and the USA. In Germany, concentrations in effluents ranged from 0.01 to 5.9 µg/l (Bohlen et al., 1989). In the United Kingdom, concentrations in effluent were typically up to 2 µg/l, with a maximum of 144 µg/l (Brown, 1978). The median concentration in a study in the USA was 5 µg/l (Staples et al., 1985). Swiss effluent samples contained 0.03–6.4 µg/l (means 0.16–1.0 µg/l) (Fahmi, 1984). Effluent concentrations are lower than influent concentrations. Influent and effluent concentrations were 15 µg/l and 1 µg/l, respectively, in a Swiss study (Fahmi, 1984). In samples taken in various regions of France, influent concentrations ranged from 1.05 to 23 µg/l, while concentrations in effluents ranged from not detectable to 8.5 µg/l; the detection limits were not disclosed (Marchand et al., 1988, 1989). One source document reports that the influent of a sewage treatment plant contained tetrachloroethene at 6.2 µg/l, while the effluent contained 3.9 and 4.2 µg/l before and after chlorination, respectively (de Raat, 2003).

Tetrachloroethene was measured at 2.8–10 µg/l in effluents from various industrial activities (ceiling coating material manufacture, metalworking, chemical product packaging, treatment of industrial effluents, surface treatment, wood preservation, and paint manufacture), at 7–29 µg/l in effluent from a car equipment manufacturing facility, and at 508 µg/l in a chemical industry effluent (DRIRE Franche Comté, 1996). Dry cleaning industry effluent in Finland contained 2.5–580 000 µg/l (geometric mean 2.5 µg/l, arithmetic mean 88 µg/l) (Finnish Environment Agency, personal communication, 1996, cited in EC, 2001).

6.1.8 Food

De Raat (2003) presented a table (reproduced here as Table 3) of published data on tetrachloroethene concentrations in food. Being an adaptation of information from three review sources (IPCS, 1984; IARC, 1995; ATSDR, 1997), these figures might not reflect current concentrations.

Table 3: Concentration of tetrachloroethene in food products.^a

^a As presented in source document (de Raat, 2003).

The most important routes of exposure to tetrachloroethene for members of the general population appear to be inhalation of the compound in ambient air and ingestion via drinking-water. Available data indicate that dermal exposure is not important for most people (de Raat, 2003).

An EU risk assessment for human health is being drafted. The current draft contains estimated figures for daily human intake, based on typical human consumption and inhalation rates. The total human dose of tetrachloroethene based on "background" exposure is estimated (using "reasonable worst-case assumptions") to be 0.43 µg/kg body weight per day, or 30 µg/day for a person weighing 70 kg. Equivalent estimated figures are higher for persons living near a manufacturer (19 µg/kg body weight per day) or above a dry cleaning establishment (1.67 mg/kg body weight per day) (EC, 2004).

In the USA, the average daily intake by the inhalation route, assuming ambient concentrations of 2.1–17.3 µg/m³ and inhalation of 20 m³ of air per day, is estimated to be 41–204 µg. The average daily intake from water, assuming concentrations of 0.3–3 µg/l and ingestion of 2 litres of water per day, is estimated to be 0.6–6 µg (ATSDR, 1997).

In Switzerland and Germany, total daily intakes via food were calculated to be 160 µg and 87 µg, respectively (IPCS, 1984).

The breath of residents living above 12 dry cleaning shops in the Netherlands was found to contain a mean tetrachloroethene concentration of 5 mg/m³, while the breath of residents living adjacent to the shops contained 1 mg/m³ (IPCS, 1984). Tetrachloroethene and trichloroacetic acid concentrations in blood and trichloroacetic acid concentrations in urine were determined primarily over the course of a week for 29 persons living in the vicinity of dry cleaning shops in Germany. The concentrations of tetrachloroethene in blood depended on the floor and the construction type of the building where the people resided, but not on the type of system used in the dry cleaning shops (Popp et al., 1992).

In Turin, Italy, blood samples of 30 volunteers (15 females, 15 males) contained a mean tetrachloroethene concentration of 1.33 µg/l during winter and 0.46 µg/l during summer (Gilli et al., 1990a) (see also section 6.1.1).

A study in Modena, Italy, reported tetrachloroethene levels in the ambient air of 30 homes of dry cleaners (located well away from the dry cleaning premises), alveolar air contemporaneously from the (36) dry cleaners, and samples of end-exhaled air (alveolar air) from 34 subjects who were not themselves occupationally exposed, but who were members of the households of dry cleaners. These were compared with samples from 41 members of the general population (located in the same district near the dry cleaners’ homes). Tetrachloroethene levels in dry cleaners’ homes were significantly higher than in control houses (geometric means: 265 vs 2 µg/m³, P < 0.001). Tetrachloroethene levels in the alveolar air exhaled by dry cleaners, their family members, and control subjects were statistically significantly different (geometric means: 5140, 225, and 3 µg/m³, respectively; P < 0.001) (Aggazzotti et al., 1994).

Due to the age of the reports, data in this section might not reflect current experience.

Exposure levels for organic solvents at Dutch workplaces were measured by the Dutch Ministry of Social Affairs and Employment. During cleaning activities in dry cleaning establishments, metal industries (cleaning machinery parts and degreasing activities), and offset-printing offices, breathing-zone air levels of up to 350 mg/m³, 270 mg/m³, and 110 mg/m³ were observed, respectively (Doorgeest et al., 1986).

A NIOSH (USA) survey (1977–1979) of 44 dry cleaning facilities showed exposures for machine operators to range from 30 to 1030 mg/m³. Geometric mean exposures for machine operators, pressers, and seamstresses and in front counter areas were 150, 23, 21, and 21 mg/m³, respectively. A study of the dry cleaning industry in the United Kingdom indicated exposure levels similar to those observed in American studies (ATSDR, 1997).

An 8-h TWA concentration of up to 4000 mg/m³ can occur in dry cleaning establishments. In the United Kingdom, over 90% of 493 8-h measurements in 131 dry cleaning establishments revealed concentrations below 680 mg/m³, and over 50% of these samples revealed concentrations below 200 mg/m³. Similar results were obtained in a survey of 46 dry cleaning establishments in Germany (IPCS, 1984).

At a railway works where tetrachloroethene was used as a cleaning agent, 94% of 104 8-h measurements exceeded 680 mg/m³, with peaks up to 1290 mg/m³ (IPCS, 1984).

A number of investigations have shown that tetrachloroethene is well absorbed by humans exposed by inhalation (Stewart et al., 1961b, 1970; Fernandez et al., 1976; Hake & Stewart, 1977; Monster et al., 1979, 1983; Benoit et al., 1985; Opdam & Smolders, 1986, 1987; Pezzagno et al., 1988), as would be predicted from the blood/air partition coefficient (Table 4). On exposure at 500 or 990 mg/m³, respiratory absorption was in excess of 90% at the start of exposure and fell to about 50% after 8 h of exposure (Monster et al., 1979). Absorption was 78–93% in volunteers exposed at 340–630 mg/m³ (Benoit et al., 1985).

Table 4: Blood/air, tissue/air, and tissue/blood human partition coefficients for tetrachloroethene.^a

^a From Gearhart et al. (1993).

An absorption figure of 50% was reported in rats exposed at 340 mg/m³ for 3 h (Dallas et al., 1994a). The amounts collected post-exposure (in exhaled air, urine, and carcass) in rats and mice exposed by inhalation suggest that absorption from the respiratory tract is considerable (Pegg et al., 1979; Schumann et al., 1980, 1982). Following a 4-h whole body exposure of rats at 3320 mg/m³, tetrachloroethene was found in the blood at 26 mg/l (Frantik et al., 1998).

Data on dermal absorption in humans are limited. A 3-min exposure of the forearm (27 cm² area) of six volunteers to tetrachloroethene liquid resulted in an average dermal absorption rate of 0.68 mg/cm² per minute (Kezic et al., 2001). It has been estimated that immersion of the hands and forearms in tetrachloroethene for 1 h would correspond to an inhalation exposure of 850 mg/m³ for 8 h. When five subjects immersed one thumb in tetrachloroethene for 40 min (inhalation was prevented), the mean tetrachloroethene concentration in exhaled air peaked at 2.1 mg/m³ within 10 min of the end of exposure and slowly decreased to 1.4 mg/m³ after 5 h (Stewart & Dodd, 1964). When two subjects immersed one hand in tetrachloroethene for 5 min (with prevention of inhalation), blood concentrations immediately after exposure were much higher in the immersed arm than in the non-immersed arm, and concentrations did not become similar until 2 h had passed, suggesting considerable local tissue absorption (Aitio et al., 1984). Upon exposure to the vapour, respiratory absorption seems to be much more important than dermal absorption. When three volunteers were exposed for 3.5 h at 4000 mg/m³ via the skin only, 48 mg was exhaled in the following 50 h. It was estimated that 4.2 g would have been exhaled following combined skin and respiratory exposure at 4000 mg/m³ for 3.5 h, suggesting that the inhalation route would account for about 99% of the total absorption in subjects exposed by both routes (Riihimäki & Pfäffli, 1978).

When 0.5 ml of tetrachloroethene was applied to the skin (2.9 cm²) of intact mice for 15 min, the total tetrachloroethene absorbed was estimated to be 177 µg, of which 173 µg was found in exhaled air. This amount equates to an absorption rate of 24.4 nmol/cm² per minute (Tsuruta, 1975). Absorption in vitro appeared to be much slower, a figure of 0.067 nmol/cm² per minute being calculated when tetrachloroethene was applied to isolated rat skin (Tsuruta, 1977).

Quantitative data on absorption in humans following ingestion are not available. However, the systemic toxicity and the presence of tetrachloroethene and its metabolites in the blood and urine of humans who accidentally ingested tetrachloroethene suggest that it is readily absorbed through the human gastrointestinal tract (Koppel et al., 1985; ATSDR, 1997).

Laboratory animal studies indicate rapid and extensive absorption of tetrachloroethene following oral administration (Daniel, 1963; Pegg et al., 1979; Frantz & Watanabe, 1983; Mitoma et al., 1985; Clement International Corporation, 1990; ATSDR, 1997).

In humans repeatedly exposed by inhalatio