This report contains the collective views of an international group of experts and does not necessarily represent the decisions or the stated policy of the United Nations Environment Programme, the International Labour Organization, or the World Health Organization.

Concise International Chemical Assessment Document 51

First draft prepared by Dr Bob Benson, US Environmental Protection Agency, Denver, Colorado, USA

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.

World Health Organization

Geneva, 2003

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.

The Inter-Organization Programme for the Sound Management of Chemicals (IOMC) was established in 1995 by UNEP, ILO, the Food and Agriculture Organization of the United Nations, WHO, the United Nations Industrial Development Organization, the United Nations Institute for Training and Research, and the Organisation for Economic Co-operation and Development (Participating Organizations), following recommendations made by the 1992 UN Conference on Environment and Development to strengthen cooperation and increase coordination in the field of chemical safety. The purpose of the IOMC is to promote coordination of the policies and activities pursued by the Participating Organizations, jointly or separately, to achieve the sound management of chemicals in relation to human health and the environment.

WHO Library Cataloguing-in-Publication Data

1,1-Dichloroethene (vinylidene chloride).

(Concise international chemical assessment document ; 51)

1.Dichloroethylenes - adverse effects 2.Dichloroethylenes - toxicity 3.Risk assessment

4.Environmental exposure I.International Programme on Chemical Safety II.Series

ISBN 92 4 153051 0           (NLM Classification: QV 633)

ISSN 1020-6167

©World Health Organization 2003

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FOREWORD

Concise International Chemical Assessment Documents (CICADs) are the latest in a family of publications from 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 join the Environmental Health Criteria documents (EHCs) 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 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 on page 2 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.

The Steering Group will also advise IPCS on the appropriate form of the document (i.e., EHC or 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 based on an existing national, regional, or international review. 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.

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.

Board members, authors, reviewers, consultants, and advisers who participate in the preparation of a CICAD are required to declare any real or potential conflict of interest in relation to the subjects under discussion at any stage of the process. Representatives of nongovernmental organizations may be invited to observe the proceedings of the Final Review Board. Observers may participate in Board discussions only at the invitation of the Chairperson, and they may not participate in the final decision-making process.

This CICAD on 1,1-dichloroethene (vinylidene chloride) was prepared by the US Environmental Protection Agency (EPA). The US EPA considered data identified as of April 2001 in the national assessment document (US EPA, 2002d). Information on the nature of the peer review and availability of this national assessment document is presented in Appendix 1. The literature search for this CICAD was updated in August 2002. Information on the peer review of this CICAD is presented in Appendix 2. This CICAD was approved as an international assessment at a meeting of the Final Review Board, held in Monks Wood, United Kingdom, on 16–19 September 2002. Participants at the Final Review Board meeting are presented in Appendix 3. The International Chemical Safety Card for 1,1-dichloroethene (ICSC 0083), produced by the International Programme on Chemical Safety (IPCS, 2000), has also been reproduced in this document.

1,1-Dichloroethene (CAS No. 75-35-4), or 1,1-DCE, does not occur naturally. It is produced commercially by the dehydrochlorination of 1,1,2-trichloroethane in the presence of excess base or by thermal decomposition of methyl chloroform (1,1,1-trichloroethane). 1,1-DCE is used as a captive intermediate in the production of hydrochlorofluorocarbons (HCFC-141b and HCFC-142b), in the production of chloroacetyl chloride, and in the production of homo-, co-, and terpolymers (latex and resin). The polymers are used in a variety of consumer products, including food packaging, textiles, and outdoor furniture.

1,1-DCE can be found in the environment from release during its manufacture and use, from the breakdown of polyvinylidene (PVDC) products, and from the biotic or abiotic breakdown of 1,1,1-trichloroethane, tetrachloroethene, 1,1,2-trichloroethene, and 1,1-dichloroethane. The principal sources of environmental exposure for humans are ambient air and contaminated drinking-water.

In groundwater, biotransformation of 1,1-DCE can form vinyl chloride through reductive dechlorination.

The high vapour pressure and low water solubility of 1,1-DCE favour relatively high concentrations in the atmosphere compared with other environmental compartments. Atmospheric hydroxyl radicals play a major role in the degradation of 1,1-DCE. The atmospheric half-life is estimated at 16 h. Volatilization is the major transport process from water, soil, and sediment. Bioaccumulation is expected to be low based on the chemical’s octanol/water partition coefficient and water solubility.

1,1-DCE is rapidly absorbed following inhalation and oral exposure. Because of its low relative molecular mass and hydrophobic nature, dermal absorption is also likely, but there are no relevant published data. Although 1,1-DCE is rapidly distributed to all tissues, most of the free 1,1-DCE, its metabolites, and covalently bound derivatives are found in the liver and kidney. 1,1-DCE is rapidly oxidized by cytochrome P450-dependent monooxygenase 2E1 (CYP2E1) to 1,1-dichloroethene oxide (DCE-epoxide), 2-chloroacetyl chloride, and 2,2-dichloroacetaldehyde. The major metabolites, DCE-epoxide and 2-chloroacetyl chloride, can react with glutathione (GSH), water, or tissue macromolecules. It is not known if the metabolism of 1,1-DCE is the same in humans, although in vitro microsomal preparations from human liver and lung form the same initial products.

The only existing epidemiological study is inadequate to assess the cancer or non-cancer health effects of 1,1-DCE.

Following high-dose exposure by the oral or inhalation route, the target organs in experimental animals are the liver, the kidney, and the Clara cells of the lung. Following low-dose, long-term exposure, the liver is the major target organ in rats following oral or inhalation exposure, but the kidney is the major target organ in mice following inhalation exposure.

Bioassays for cancer by the oral route of exposure have been conducted in rats, mice, and trout. Although these bioassays have protocol limitations, none provides any significant evidence that 1,1-DCE is a carcinogen by the oral route of exposure. Bioassays for cancer by the inhalation route of exposure have been conducted in rats, mice, and hamsters. Most of these bioassays also have protocol limitations. One bioassay in male mice showed an increase in the incidence of kidney adenocarcinomas at one exposure level. There is evidence that the induction of kidney adenocarcinomas is a sex- and species-specific response related to the expression of CYP2E1 in the kidney of male mice. The results of the one bioassay showing an increase in tumours in one sex and at one exposure level in a single rodent species are not sufficient to justify an exposure–response assessment.

1,1-DCE causes gene mutations in microorganisms in the presence of an exogenous activation system. Most tests with mammalian cells in vitro or in vivo show no evidence of genotoxicity.

There is no evidence that reproductive toxicity or teratogenicity is a critical effect for 1,1-DCE. No reproductive or developmental toxicity was observed at an oral exposure that caused minimal toxicity in the liver of the dams. There is some evidence of developmental variations in the heart after oral exposure, but it is not clear if these effects are directly caused by exposure to 1,1-DCE. There is evidence of fetal toxicity (delayed ossification) following inhalation exposure in the absence of maternal toxicity.

One study shows no evidence that 1,1-DCE causes skin sensitization.

The toxicity of 1,1-DCE is associated with cytochrome P450-catalysed metabolism of 1,1-DCE to reactive intermediates that bind covalently to cellular macromolecules. The extent of binding is inversely related to loss of GSH, so that severities of tissue damage parallel the decline in GSH. Thus, the responses to 1,1-DCE at low doses with little depletion of GSH are expected to be very different from the responses at high doses causing substantial GSH depletion.

The critical effect from oral exposure is minimal hepatocellular mid-zonal fatty change in female Sprague-Dawley rats. Based on a BMDL₁₀ (the lower 95% confidence limit on the benchmark dose [BMD] for a 10% response) of 4.6 mg/kg body weight per day and a total uncertainty factor of 100, the tolerable intake is 0.05 mg/kg body weight per day.

The critical effect from inhalation exposure is minimal hepatocellular mid-zonal fatty change in female Sprague-Dawley rats. Based on a BMCL₁₀ (the lower 95% confidence limit on the benchmark concentration [BMC] for a 10% response) of 6.9 mg/m³ and a total uncertainty factor of 30, the tolerable concentration is 0.2 mg/m³.

Human exposure to 1,1-DCE is likely to be highly variable due to site-specific contamination. However, data suggest that the mean exposure from drinking-water will not exceed 6–9 × 10^–5 mg/kg body weight per day for a 70-kg individual consuming 2 litres per day. The oral exposure from food and soil is most likely negligible. Data suggest that the upper end of the range for the mean concentration of 1,1-DCE in air will not exceed 0.004 mg/m³. Thus, human exposure is expected to be far below the tolerable intake of 0.05 mg/kg body weight per day and the tolerable concentration of 0.2 mg/m³.

There are only limited data on the effects of 1,1-DCE in the aquatic and terrestrial environments. In studies conducted in closed systems, the EC₂₀ for inhibition of the growth of a mixed methanotrophic culture was 0.05 mg/litre; the 72-h EC₅₀ for inhibition of growth of green alga Chlamydomonas reinhardtii was 9.12 mg/litre; and the 96-h LC₅₀ for bluegill (Lepomis macrochirus) was 74 mg/litre. The limited data on occurrence of 1,1-DCE in surface water suggest that concentrations are in the microgram per litre range, indicating that acute toxic risks from 1,1-DCE for the aquatic environment are minimal. There are no long-term toxicity data with which to assess sublethal effects of 1,1-DCE on any organisms. However, because of the rapid volatilization of 1,1-DCE from the aquatic and terrestrial environments, no significant risk is expected.

Synonyms for 1,1-dichloroethene (CAS No. 75-35-4; C₂H₂Cl₂) include 1,1-DCE, 1,1-dichloroethylene, vinylidene chloride, and vinylidene dichloride. Its relative molecular mass is 96.94; its solubility in water is 2.5 g/litre; its octanol/water partition coefficient (log K_ow) is 1.32; its vapour pressure is 67 kPa at 20 °C; and its measured Henry’s law constant is 23.2 kPa·m³/mol at 20 °C. Its chemical structure is shown below:

1,1-Dichloroethene

The chemical and physical properties of 1,1-DCE are summarized in the International Chemical Safety Card (ICSC 0083), which is included in this document.

The conversion factors² for 1,1-dichloroethene in air (at 20 °C and 101.3 kPa) are as follows:

1 ppm = 4.0 mg/m³

1 mg/m³ = 0.25 ppm

Because of its volatility, 1,1-DCE is well suited to determination by gas chromatography (GC) using a variety of detectors, including flame ionization (FID), electron capture, electrolytic conductivity (ECD), and mass spectrometry (MS). The major analytical limitation is interference by other constituents of the media analysed.

Methods are available for quantifying 1,1-DCE in environmental samples (air, water, soil, sediment). Quantification of 1,1-DCE in air with a detection limit in the 1 µg/m³ range typically uses cold or absorbent column trapping followed by GC/FID (Foerst, 1979; Sidhu, 1980). A method for whole air samples using SUMMA canisters with a comparable method of detection is available (US EPA, 1988; Brymer et al., 1996). A method with a working range of 2–20 mg/m³ for a 5-litre air sample is available for monitoring in the workplace (NIOSH, 1994). This method uses GC/FID. Quantification of 1,1-DCE in water with a detection limit in the 0.1–0.5 µg/litre range uses headspace analysis and GC/FID or GC/ECD (Piet et al., 1978; Otson & Williams, 1982). Other methods with a comparable detection limit include purge and trap with GC/ion trap detector (Eichelberger et al., 1990; US EPA, 1995) and purge and trap with GC/MS (US EPA, 1998). Some newer techniques for sample collection include solid-phase microextraction (Arthur et al., 1992; Shirey, 1995) and sample introduction using membrane technology (Bauer & Solyom, 1994; Wong et al., 1995). Quantification of 1,1-DCE in soils and sediments with a detection limit in the 5–10 µg/kg range uses extraction with an organic solvent or purging with an inert gas, trapping, and GC/MS (DeLeon et al., 1980; Speis, 1980; Amaral et al., 1994; US EPA, 1998).

Methods are also available for quantifying 1,1-DCE in biological samples (breath, food, body tissues). Quantification of 1,1-DCE in human breath with a detection limit of 0.16 µg/m³ uses spirometry for sample collection, cryogenic or Tenax trapping, and GC/MS (Wallace et al., 1982, 1984). Quantification of 1,1-DCE in food (Gilbert et al., 1980) and body tissues (Lin et al., 1982) with a detection limit in the 5–10 µg/kg range uses a headspace technique (Gilbert et al., 1980), purge and trap, and GC/ECD (Lin et al., 1982) or GC/MS (Easley et al., 1981; Hiatt, 1983).

1,1-DCE is not known to occur as a natural product. It is produced commercially by the dehydrochlorination of 1,1,2-trichloroethane in the presence of excess base or by thermal decomposition of methyl chloroform (1,1,1-trichloroethane). 1,1-DCE is used as a captive intermediate in the production of hydrochlorofluorocarbons (HCFC-141b and HCFC-142b), in the production of chloroacetyl chloride, and in the production of homo-, co-, and terpolymers (latex and resin) (also known as PVDC polymers) (W. Stott, personal communication, 2002). These polymers are produced as emulsion polymers, as solvent-soluble powders for coating applications, and as resins for extrusion and co-extrusion. PVDC copolymers containing 79–90% 1,1-DCE are used to form moisture and vapour barrier coatings and films, with applications as food packaging products. PVDC copolymers containing 10–70% 1,1-DCE are used to improve flame and ignition resistance properties in the final product. Residual 1,1-DCE in PVDC used for food packaging products typically ranges from 5 to <1 mg/kg, the limit of detection of the method. Other consumer products containing PVDC include PVDC-latex for carpet backing (<2 mg/kg residual 1,1-DCE), PVDC-latex for Foil Scrim Kraft (<3 mg/kg residual 1,1-DCE), PVDC-latex for photographic film coating (<100 mg/kg residual 1,1-DCE), PVDC for flame retardant fibres for clothing and outdoor awnings (<100 mg/kg residual 1,1-DCE), and PVDC-fluorinated copolymers for application on textiles (<100 mg/kg residual 1,1-DCE). Further processing decreases the residual 1,1-DCE in the final consumer product.

1,1-DCE can be found in the environment from release during its manufacture and use; from the breakdown of PVDC products; and from the biotic or abiotic breakdown of 1,1,1-trichloroethane, tetrachloroethene, 1,1,2-trichloroethene, and 1,1-dichloroethane. 1,1-DCE is frequently found at hazardous waste sites.

In the early 1980s, annual world production was estimated at 306 000 tonnes (IPCS, 1990). IPCS (1990) estimated that 1%, or 3000 tonnes, was released to ambient air. The US EPA (2002c) reported that 74 tonnes were released to ambient air and 0.06 tonne was released to surface water in 1999 in the USA. For the period 1988–1999, the average annual release to ambient air was 99 tonnes and the average annual release to surface water was 0.39 tonne in the USA (US EPA, 2002c).

5.1.1 Air

The high vapour pressure and low water solubility of 1,1-DCE favour relatively high concentrations in the atmosphere compared with other environmental compartments. Atmospheric hydroxyl radicals play a major role in the degradation of 1,1-DCE. The rate constant for the oxidation reaction ranges from 0.8 × 10^–11 to 2.6 × 10^–11 cm³/molecule per second, with an atmospheric half-life estimated at 16 h (Grosjean, 1991). The major reaction products are formaldehyde, phosgene, and hydroxyacetyl chloride. Removal of 1,1-DCE from the atmosphere by reactions with chlorine atoms, peroxy radicals, ozone, and nitrate are quantitatively insignificant, as is removal by rain droplets or by adsorption on particulate matter in the atmosphere (IPCS, 1990; Grosjean, 1991).

5.1.2 Water

Consideration of the physical and chemical properties of 1,1-DCE indicates that volatilization is the major transport process from water. The half-life for evaporation of 1,1-DCE (1 g/litre) from a stirred aqueous solution with a depth of 6.5 cm was 27.2 min at 20 °C (Dilling, 1977). The calculated half-life of 1,1-DCE was 6 days in static pond water and 1 day in mobile river water (IPCS, 1990). Photolysis and hydrolysis are not likely to be significant pathways for removal of 1,1-DCE from aqueous solutions.

5.1.3 Soils and sediments

Volatilization is expected to be the major process for the removal of 1,1-DCE from soils and sediments. The chemical’s octanol/water partition coefficient and solubility in water also suggest that leaching from soils to water can occur.

MITI (1992) reported no significant biodegradation in a closed bottle test using activated sludge. This test was conducted according to Organisation for Economic Co-operation and Development (OECD) Guideline 302C. However, Tabak et al. (1981) measured microbial degradation of 78% of 1,1-DCE (5 mg/litre) following 7 days of incubation at 25 °C in a static culture flask, in the dark, with settled domestic wastewater as microbial inoculum. With subsequent incubations (after adaptation), 100% loss of compound occurred. At 10 mg/litre, 45% loss was found in the first 7 days of incubation. Volatilization losses over 7 days at 25 °C were 24% and 15% at 5 and 10 mg/litre, respectively.

The biotransformation of 1,1-DCE has been studied in anaerobic microcosms (Barrio-Lage et al., 1986; Vogel & McCarty, 1987). The biotransformation first produces vinyl chloride by reductive dehalogenation, which is subsequently mineralized to carbon dioxide. Barrio-Lage et al. (1986) obtained sediment and water from two different locations in the Florida Everglades. This material was spiked with 1,1-DCE to 5 mg/litre and incubated anaerobically at 25 °C in the dark for up to 6 months. In this experiment, 1,1-DCE was consumed and vinyl chloride was produced. The first-order rate constant for depletion of 1,1-DCE was 1.67 × 10^–4/h in one microcosm and 3.57 × 10^–4/h in the other, indicating a very slow rate of dehalogenation. Not all of the 1,1-DCE depleted appeared as vinyl chloride, indicating that mechanisms of biotransformation other than reductive dechlorination occur. Glod et al. (1997) presented evidence for an alternative pathway of biotransformation involving cobalamin. Cobalamin is involved in the enzymatic reduction of halogenated ethenes by a variety of anaerobic bacteria. Glod et al. (1997) investigated the mechanism of the reduction of 1,1-DCE in homogeneous solutions containing titanium(III) citrate as the bulk electron donor. Cob(I)alamin reduced 1,1-DCE to ethene and ethane in pH-dependent reactions. At pH 7, acetylene was detected as the major intermediate, and only a minor amount of vinyl chloride was formed. A possible mechanism involves addition of cob(I)alamin to 1,1-DCE. The deprotonated form of the β-dichloroethyl cobalamin then loses a chloride to form a carbene intermediate, which is transformed into a β-chlorovinyl cobalamin by transfer of an α-hydrogen. The β-chlorovinyl cobalamin is converted to acetylene by β-elimination through a cobalt (III)-pi complex. At pH 9, where the overall reaction was much slower than at pH 7, about equal amounts of vinyl chloride and ethene were formed simultaneously. The formation of vinyl chloride at pH 9 possibly involves two different pathways. The first pathway involves β-elimination of a chloride from β-dichloroethyl cobalamin with a cobalt (III)-pi complex as an intermediate. The second pathway involves a dissociative electron transfer yielding a chlorovinyl radical as the intermediate.

Dolan & McCarty (1995) studied the aerobic biotransformation of 1,1-DCE using a mixed methanotrophic culture isolated from a US Superfund site. The initial oxidation of 1,1-DCE is believed to be catalysed by methane monooxygenase, which has a broad substrate specificity. The degradation of 1,1-DCE stopped after the first few hours due to the toxicity of the metabolites of 1,1-DCE, most likely the epoxide or an acyl chloride.

Bioaccumulation is expected to be low based on the octanol/water partition coefficient and water solubility of 1,1-DCE. A bioconcentration factor of 4 and a bioaccumulation factor of 6.9 were reported for fish (Atri, 1985). A bioaccumulation factor of less than 13 was reported for common carp (Cyprinus carpio) (MITI, 1992).

6.1.1 Air

Singh et al. (1981) and Brodzinski & Singh (1983) reported data on 1,1-DCE in ambient air in 30 US locations, and Guicherit & Schulting (1985) measured 1,1-DCE in ambient air at three sites in the Netherlands. All three data sets are consistent and indicate typical mean values of 20–120 µg/m³ and maxima of 40–560 µg/m³. Measurements taken in industrial areas near sources of 1,1-DCE in these studies yielded a much higher mean value of 120 × 10³ µg/m³ and a maximum of 270 × 10³ µg/m³. US EPA (1985) reported mean values of 0.02 µg/m³ in non-industrial areas and 8.7 µg/m³ in industrial areas. US EPA (2002a) compiled ambient air quality data from 1982 to 2001 for 1,1-DCE at locations throughout the USA. The data for 2001 provide the arithmetic mean of the samples collected from 87 individual locations. The range in the arithmetic mean values for these locations was 0.004–4 μg/m³.³

6.1.2 Water

Of the 439 hazardous waste sites in the USA with detections of 1,1-DCE, 186 sites reported quantitative information on the concentration of 1,1-DCE in water (US EPA, 2002b). The mean maximum detected concentration in water was 2 mg/litre.

Hallbourg et al. (1992) investigated the occurrence of 1,1-DCE in groundwater and surface water near three municipal landfills in Florida, USA, using a method with a detection limit of 1 µg/litre. 1,1-DCE was not detected in any surface water or in any well near two of the landfills. At the third landfill, 1,1-DCE was not detected in two wells, but was detected in a third well at 24.4 µg/litre.

Yamamoto et al. (1997) determined the concentration of 1,1-DCE in surface water from 30 sites in Osaka, Japan, using a method with a detection limit of 0.4 µg/litre. 1,1-DCE was detected in only 3 of 136 samples. The maximum concentration found was 1 µg/litre.

Stangroom et al. (1998) reported on the occurrence of 1,1-DCE in 1995 from data collected by the National Centre for Toxic and Persistent Substances in the United Kingdom. 1,1-DCE was detected in 4% of industrial effluents in England and Wales at a mean concentration of 0.67 µg/litre.

6.1.3 Sewage sludge

Wilson et al. (1994) reported on the occurrence of 1,1-DCE in digested sewage sludge from 12 locations (rural, urban, and industrial) in northwest England. The mean concentration of 1,1-DCE was 7.97 mg/kg dry weight (range 1.92–16.6 mg/kg dry weight).

6.1.4 Soil

Of the 439 hazardous waste sites in the USA with detections of 1,1-DCE, 45 reported quantitative information on the concentration of 1,1-DCE in soil (US EPA, 2002b). The mean maximum detected concentration in soil was 90 mg/kg.

For the general population, potential sources of exposure to 1,1-DCE include breathing contaminated air, drinking contaminated water, dermal contact with contaminated water, and eating contaminated food.

Wallace et al. (1986) reported a 5-year US EPA study in urban populations of personal exposures to 1,1-DCE from air. Among the 400 participants, 1,1-DCE was detected only occasionally in personal air samples at levels exceeding 1 µg/m³ based on a 24-h average. No recent data were located.

Storm (1994) summarized the results of monitoring drinking-water sources in California, USA, from 1984 to 1992 using a method with a detection limit of 0.1 µg/litre. Of 11 686 sources sampled, 120 contained 1,1-DCE with a mean concentration of 2.45 µg/litre (based on 1437 individual samples from the 120 sources).⁴

Biziuk et al. (1996) investigated the occurrence of 1,1-DCE in the drinking-water of the Gdansk district of Poland using a method with a detection limit of 0.01 µg/litre. 1,1-DCE was not detected in the 22 samples analysed.

Chung et al. (1997) reported on the occurrence of 1,1-DCE in raw water, treated water, and tap water from 1993 to 1995 in water supplies from six cities in Korea (Seoul, Pusan, Taegu, Taejon, Kwangju, and Inchon) using a method with a detection limit of 0.012 µg/litre. The mean concentration of 1,1-DCE in raw water and treated water at the treatment plants and in tap water from individual households was 0.012, 0.022, and 0.019 µg/litre, respectively.

No reliable data are available to estimate the exposure from food. However, this source is expected to be negligible.

In the workplace, Ott (1976) reported a range of 20–280 mg/m³ for 1,1-DCE in a fibre production facility. These data were collected in 1960 and 1965. US EPA (1985) reported levels of 1,1-DCE in monomer and polymer plants of 90–100 µg/m³ and 25–50 µg/m³, respectively. No recent data were located.

1,1-DCE is rapidly absorbed following inhalation and oral exposures. Because of its low relative molecular mass and hydrophobic nature, dermal absorption is also likely, but there are no relevant published data. In rats treated with 1,1-DCE by gavage in corn oil, complete gastrointestinal absorption occurs at <350 mg/kg body weight (Jones & Hathway, 1978a,b; Putcha et al., 1986). 1,1-DCE is easily transported across the alveolar membrane. At a concentration of 600 mg/m³ and below in inspired air, equilibrium or near steady state is reached in the blood in rats in approximately 45 min (Dallas et al., 1983). Continued uptake in rats reflects to some extent continuing deposition in fatty tissues, but is primarily a result of metabolism of 1,1-DCE.

The major route of excretion for unchanged 1,1-DCE is through the lungs (Jones & Hathway, 1978a). The majority of the 1,1-DCE, however, is rapidly metabolized to non-volatile compounds and covalently bound derivatives (McKenna et al., 1978a,b). In both animal and human tissue, CYP2E1 catalyses the initial oxidation of 1,1-DCE (Dowsley et al., 1996). The covalent binding and cellular damage in kidney, lung, and liver correlate with the high concentration of CYP2E1 in certain cell populations in these tissues (Forkert, 2001). Mice metabolize more 1,1-DCE than rats. For example, when given 50 mg/kg body weight by oral gavage in corn oil, mice excrete 6% and rats excrete 28% of the dose as unchanged 1,1-DCE through the lungs (Jones & Hathway, 1978b). When exposed to 40 mg/m³ by inhalation for a single 6-h period, mice excrete 0.65% and rats excrete 1.63% of the absorbed dose as unchanged 1,1-DCE through the lungs (McKenna et al., 1977). Intraperitoneal administration of [¹⁴C]1,1-DCE at 125 mg/kg body weight to mice resulted in the highest concentrations of covalent binding (based on protein content) in the kidney, lung, and liver (Okine et al., 1985; Okine & Gram, 1986a,b).

Fig. 1: Pathways for 1,1-dichloroethene metabolism and toxicity (see text for details).

The primary metabolites of 1,1-DCE formed in rat hepatic microsomal incubations are DCE-epoxide, 2,2-dichloroacetaldehyde, and 2-chloroacetyl chloride (Costa & Ivanetich, 1982, 1984; Liebler et al., 1985, 1988). These metabolites were also identified from mouse microsomal incubations (Dowsley et al., 1995). All these electrophilic metabolites undergo secondary reactions, including oxidation, conjugation with GSH, and hydrolysis. The major products formed are GSH conjugates, 2-(S-glutathionyl)acetyl glutathione [B] and 2-S-glutathionyl acetate [C], that are believed to be derived from the DCE-epoxide (Fig. 1). S-(2,2-Dichloro-1-hydroxy)ethyl glutathione [A], the GSH conjugate formed from reaction of GSH with 2,2-dichloroacetaldehyde, was not observed in liver microsomal incubations containing GSH (Dowsley et al., 1995). The acetal, together with chloroacetic acid and S-(2-chloroacetyl)-glutathione [D], the hydrolysis and GSH-conjugated products of 2-chloroacetyl chloride, respectively, were detected at levels much lower than those of the DCE-epoxide-derived conjugates [B] and [C]. In human liver and lung microsomal incubations (Dowsley et al., 1999), the DCE-epoxide-derived GSH conjugates [B] and [C] were the major metabolites detected. 2,2-Dichloroacetaldehyde was detected at low levels. Liver microsomes from three out of five human samples metabolized 1,1-DCE to the epoxide-derived GSH conjugates at levels that were 2.5- to 3-fold higher than those in mouse liver microsomes, based on GSH conjugate formed per milligram of microsomal protein. These GSH conjugates were also the major products formed in lung microsomes from eight human samples; only low levels of 2,2-dichloroacetaldehyde were formed. The mean level of GSH conjugates formed by lung microsomes from humans was about 50% of the amount formed by lung microsomes from mice.

The significance of the metabolic pathway in the liver involving 2,2-dichloroacetaldehyde is unclear. Existing evidence, however, suggests that this pathway is of minor toxicological importance. In addition to 2,2-dichloroacetaldehyde and the GSH conjugate, other potential metabolites include the acetal (the hydration product of the aldehyde), dichloroacetic acid, and dichloroethanol. An initial study with rat liver microsomes found a trace level of 2,2-dichloroacetaldehyde but no detectable dichloroacetic acid (Costa & Ivanetich, 1982). A later report using isolated rat hepatocytes detected dichloroacetic acid and trace levels of 2,2-dichloroacetaldehyde, 2,2-dichloroethanol, and chloroacetic acid (Costa & Ivanetich, 1984). Forkert (1999a) and Forkert & Boyd (2001), using intact mice, found no acetal in liver cytosol. However, the acetal was detected in the bile in one study (Forkert, 1999a) but not mentioned as being found in the bile in the other study (Forkert & Boyd, 2001). In early studies on the metabolism of 1,1-DCE, none of the potential metabolites from this pathway was reported as being found in the urine of rodents using techniques that readily identified chloroacetic acid (McKenna et al., 1977, 1978a,b; Jones & Hathway, 1978a,b). A pharmacokinetic analysis showed that any dichloroacetic acid formed in the liver is rapidly metabolized in the liver to two-carbon, non-chlorinated chemicals and carbon dioxide (Merdink et al., 1998).

The oxidative metabolism of 1,1-DCE reaches saturation in rats at an oral exposure of 10–50 mg/kg body weight and an inhalation exposure of 790 mg/m³ (McKenna et al., 1977; Andersen et al., 1979; Dallas et al., 1983; D’Souza & Andersen, 1988).

Because 1,1-DCE is lipophilic and has a blood-to-air partition coefficient of 5 in rats (D’Souza & Andersen, 1988), any 1,1-DCE not metabolized following oral or inhalation exposure is rapidly exhaled unchanged when exposure is terminated. Based on its low octanol/water partition coefficient, 1,1-DCE will not bioaccumulate in tissues to any significant extent.

D’Souza & Andersen (1988) developed physiologically based pharmacokinetic (PBPK) models for 1,1-DCE in the rat for both oral and inhalation exposure. No validated model is available for humans. D’Souza & Andersen (1988) used allometric scaling to estimate comparative amounts of epoxide formed (mg/kg body weight) in rats and humans. Cardiac output and pulmonary ventilation were scaled by (body weight)^0.7, V_max was scaled by (body weight)^0.74, and body fat was estimated at 7% in the 200-g rat and 20% in the 70-kg human. When the oral exposure was less than 5 mg/kg body weight, the estimated amount of epoxide formed was about the same in rats and humans. When the inhalation exposure was less than 400 mg/m³, the estimated amount of epoxide formed was 5-fold lower in humans than in rats.

El-Masri et al. (1996a,b) used a combination of gas uptake experiments in Sprague-Dawley rats and PBPK modelling to assess the potential for interaction between 1,1-DCE and trichloroethene. Both substrates are activated by CYP2E1. Thus, there is a potential for competitive inhibition when simultaneous exposure to both substrates occurs. The results of the gas uptake experiments confirmed a model based on competitive inhibition. There was, however, no evidence of competitive inhibition when exposure to both substrates was 400 mg/m³ or less. As environmental exposures to these chemicals are expected to be below 400 mg/m³, there is little potential for reduced toxicity from 1,1-DCE when individuals are also exposed to trichloroethene.

Mice are more sensitive than rats to the acute toxicity of 1,1-DCE. Toxicity is enhanced by fasting (Jaeger et al., 1974, 1975, 1977a,b; Andersen & Jenkins, 1977; McKenna et al., 1978a,b; Chieco et al., 1981; Moslen et al., 1985), GSH depletion (Jaeger et al., 1974, 1977b; Andersen et al., 1980; Kanz et al., 1988; Moussa and Forkert, 1992), or vehicles that promote rapid oral absorption — for example, aqueous Tween versus mineral or corn oil (Chieco et al., 1981). Toxicity is decreased by agents that decrease metabolism by the cytochrome P450 system (Andersen et al., 1978; Moslen et al., 1989) or hypothyroidism, which increases intracellular GSH (Kanz et al., 1991).

The National Toxicology Program (NTP, 1982) conducted studies to determine lethality in five male and five female F344 rats (9 weeks old) and in five male and five female B6C3F₁ mice (9 weeks old) after a single exposure to 1,1-DCE by gavage in corn oil at 0, 10, 50, 100, 500, or 1000 mg/kg body weight. Animals were observed for 14 days. Mortality was 0/10, 1/10, 0/10, 0/10, 1/10, and 2/10 in rats and 0/10, 0/10, 1/10, 0/10, 8/10, and 10/10 in mice at 0, 10, 50, 100, 500, and 1000 mg/kg body weight, respectively. NTP did not calculate an LD₅₀ for either species or report clinical signs of toxicity.

Representative oral LD₅₀s and inhalation LC₅₀s are summarized in Table 1. Clinical signs of toxicity were not reported in these studies.

Table 1: Summary of oral LD₅₀s and inhalation LC₅₀s for exposure of rats and mice to 1,1-dichloroethene.

The target organs for toxicity after acute oral or inhalation exposure are the liver, the kidney, and the Clara cells of the lung. The effects in the liver include an increase in liver enzymes in the serum (Jenkins et al., 1972; Jaeger, 1977; Jenkins & Andersen, 1978; Reynolds et al., 1980), severe histopathological damage, including disruption of bile canaliculi, cytoplasmic vacuolization, and haemorrhagic necrosis (Reynolds et al., 1984; Kanz & Reynolds, 1986), an increase in covalent binding of 1,1-DCE (Jaeger et al., 1977a,b; Forkert & Moussa, 1991, 1993), and a decrease in GSH (Reichert et al., 1978, 1979; Kanz et al., 1988; Forkert & Moussa, 1991, 1993) mediated by CYP2E1 metabolism of 1,1-DCE to intermediates that react with GSH (Kainz et al., 1993; Lee & Forkert, 1995). Details of these studies are reported in US EPA (2002d).

Toxic effects of 1,1-DCE exposure in the kidney include increased kidney weight, increased blood urea nitrogen and creatinine (Jenkins & Andersen, 1978; Jackson & Conolly, 1985), and histopathological changes, including vacuolization, tubular dilatation, and necrosis of the proximal tubules (Jenkins & Andersen, 1978; Jackson & Conolly, 1985). These changes are correlated with metabolic activation of 1,1-DCE by CYP2E1 in the proximal tubules, decreased GSH concentration, increased covalent binding of 1,1-DCE, and the presence of a relatively high concentration of β-lyase activity in kidney tissue (Dekant et al., 1989; Brittebo et al., 1993; Dekant, 1996). In addition, renal toxicity can be inhibited by pretreatment of animals with aminooxyacetic acid, an inhibitor of renal cysteinyl-β-lyase (Ban et al., 1995; Cavelier et al., 1996).

The effects in the Clara cells of the lung include extensive histopathological changes (Forkert & Reynolds, 1982; Forkert et al., 1985, 1990), repair of damage through cell proliferation (Forkert et al., 1985), depletion of GSH, and covalent binding of 1,1-DCE mediated through the formation of DCE-epoxide by CYP2E1 (Forkert & Moussa, 1991; Moussa & Forkert, 1992; Lee & Forkert, 1995; Dowsley et al., 1996; Forkert, 1999b). Details of these studies are reported in US EPA (2002d).

Warbrick et al. (2001) tested the ability of 1,1-DCE to cause skin sensitization using the local lymph node assay. 1,1-DCE was dissolved in acetone:olive oil (4:1 v/v) to give a 0%, 10%, 25%, or 50% concentration. Groups of mice (n = 4) were exposed topically on the dorsum of both ears to 25 µl of the test solution daily for 3 consecutive days. Five hours after the injection of [³H]methyl thymidine, the mice were sacrificed. The draining auricular lymph nodes were excised and tested for incorporation of [³H]thymidine. 1,1-DCE failed to elicit a positive response at any concentration tested.

8.3.1 Oral

NTP (1982) conducted a 14-day study in male and female F344 rats (five animals of each sex, 9 weeks old) administered 1,1-DCE by gavage in corn oil. Survival was 10/10, 10/10, 10/10, 10/10, 7/10, and 3/10 at 0, 10, 50, 100, 500, and 1000 mg/kg body weight per day, respectively. Mean body weight was significantly depressed at 500 mg/kg body weight per day and above. Haemorrhagic necrosis in the liver was observed in all of the rats that died at 500 and 1000 mg/kg body weight per day. NTP did not report clinical signs of toxicity.

NTP (1982) conducted a 14-day study in male and female B6C3F₁ mice (five of each sex, 9 weeks old) administered 1,1-DCE by gavage in corn oil. Survival was 10/10, 10/10, 10/10, 10/10, 10/10, and 0/10 at 0, 10, 50, 100, 500, and 1000 mg/kg body weight per day, respectively. Haemorrhagic necrosis in the liver was observed in all mice at 1000 mg/kg body weight per day. NTP did not report clinical signs of toxicity.

8.3.2 Inhalation

Prendergast et al. (1967) evaluated the toxicity of 1,1-DCE in Long-Evans and Sprague-Dawley rats, Hartley guinea-pigs, beagle dogs, New Zealand albino rabbits, and squirrel monkeys. The test animals (15 rats per group, 15 guinea-pigs per group, 3 rabbits per group, 2 dogs per group, and 3 monkeys per group) were exposed to 1,1-DCE vapours for 8 h/day, 5 days/week, for a total of 30 exposures at 395 ± 32 mg/m³. The age of the animals was not specified. The exposed animals were evaluated for visible signs of toxicity, mortality, and haematological, pathological, and body weight changes. In this study, there were no deaths, no visible signs of toxicity, and no haematological or histopathological changes attributed to exposure to 1,1-DCE. Rabbits and monkeys lost weight (3.6% and 5.9%, respectively). The no-observed-adverse-effect level (NOAEL) in this study is 395 mg/m³ (the highest exposure tested), equivalent to an adjusted NOAEL based on continuous exposure of 94 mg/m³.

Plummer et al. (1990) exposed black hooded Wistar rats to 1,1-DCE at 1000 mg/m³ for 6 h/day, 5 days/week, for 4 weeks (six males and six females, age not specified) or continuously at 200 mg/m³ for 4 weeks (except for two 1.5-h periods per week) (18 males and 18 females, age not specified). The total exposure (concentration × time) was nearly the same for the two profiles (132 000 mg/m³·h for the continuous exposure and 120 000 mg/m³·h for the intermittent exposure). Animals in the intermittent exposure group showed signs of early coagulative necrosis in the liver (incidence not reported). Animals (11/12) in the continuous exposure group showed less severe injury, including fatty changes in variable numbers of hepatocytes and only very occasional focal liver cell necrosis. The lowest-observed-adverse-effect level (LOAEL) in this study is 200 mg/m³.

8.4.1 Oral

NTP (1982) conducted a study in male and female F344 rats (10 of each sex, 9 weeks old) administered 1,1-DCE by gavage in corn oil at 0, 5, 15, 40, 100, or 250 mg/kg body weight per day. Animals were exposed 5 times per week for 13 weeks. Representative tissues (skin, lung and bronchi, trachea, bone and bone marrow, spleen, lymph nodes, heart, salivary gland, liver, pancreas, stomach, small intestine, large intestine, kidneys, urinary bladder, pituitary, adrenal, thyroid, parathyroid, mammary gland, prostate and seminal vesicles or uterus, testis or ovary, brain, thymus, larynx, and oesophagus) from animals receiving 250 mg/kg body weight per day and from control animals were examined microscopically. Livers from all groups were examined. Three female rats receiving 250 mg/kg body weight per day died during the first week of the study. No other rats died. The mean body weight was depressed 13% for male rats receiving 250 mg/kg body weight per day compared with controls. Mean body weight in other groups was comparable. Only the liver showed effects attributed to 1,1-DCE. At 250 mg/kg body weight per day, the three female rats that died showed severe centrilobular necrosis. Minimal to moderate hepatocytomegaly was seen in the rest of the rats at 250 mg/kg body weight per day. Minimal to mild hepatocytomegaly was seen in 6/10 male rats and 3/10 female rats that received 100 mg/kg body weight per day. No biologically significant changes were observed in rats that received 40 mg/kg body weight per day or below. The NOAEL in this study is 40 mg/kg body weight per day (equivalent to 28.6 mg/kg body weight per day after adjusting for 5 days/week exposure); the LOAEL is 100 mg/kg body weight per day (equivalent to 71.4 mg/kg body weight per day after adjusting for 5 days/week exposure).

NTP (1982) conducted a study in male and female B6C3F₁ mice (10 of each sex, 9 weeks old) administered 1,1-DCE by gavage in corn oil at 0, 5, 15, 40, 100, or 250 mg/kg body weight per day. Animals were exposed 5 times per week for 13 weeks. Representative tissues (see above paragraph) from mice receiving 100 and 250 mg/kg body weight per day and from control animals were examined microscopically. Livers from all groups were also examined. Survival was 10/10, 10/10, 10/10, 9/10, 8/10, and 0/10 in males and 10/10, 9/10, 9/10, 10/10, 7/10, and 1/10 in females at 0, 5, 15, 40, 100, and 250 mg/kg body weight per day, respectively. At 100 mg/kg body weight per day, there was a decrease in mean body weight in males (14%) but not females. No biologically significant change in mean body weight was observed at lower exposures. Only the liver showed effects attributed to 1,1-DCE. Centrilobular necrosis of the liver was observed in 5/10 males and 5/10 females that received 250 mg/kg body weight per day and 2/10 males and 2/10 females that received 100 mg/kg body weight per day. No biologically significant changes in the liver occurred in mice receiving 40 mg/kg body weight per day or below. The NOAEL in this study is 40 mg/kg body weight per day (equivalent to 28.6 mg/kg body weight per day after adjusting for 5 days/week exposure); the LOAEL is 100 mg/kg body weight per day (equivalent to 71.4 mg/kg body weight per day after adjusting for 5 days/week exposure).

Quast et al. (1983) conducted a study in beagle dogs (four per group, 8 months old) administered 1,1-DCE by gavage in peanut oil at 0, 6.25, 12.5, or 25 mg/kg body weight per day for 97 days. There were no significant differences among any groups in appearance and demeanour, mortality, body weight, food consumption, haematology, urinalysis, clinical chemistry determinations, organ weights, and organ-to-body-weight ratios. No exposure-related gross or histopathological changes were present in tissues. There was no depletion of the non-protein sulfhydryl levels in the liver or kidneys. The no-observed-effect level (NOEL) in this study is 25 mg/kg body weight per day (the highest exposure tested).

8.4.2 Inhalation

Prendergast et al. (1967) evaluated the toxicity of 1,1-DCE in Long-Evans or Sprague-Dawley rats, Hartley guinea-pigs, beagle dogs, New Zealand albino rabbits, and squirrel monkeys. The test animals (15 rats per group, 15 guinea-pigs per group, 3 rabbits per group, 2 dogs per group, and 3 or 9 monkeys per group) were exposed continuously for 90 days to 1,1-DCE vapours at 20 ± 2.1, 61 ± 5.7, 101 ± 4.4, or 189 ± 6.2 mg/m³. The concurrent controls included 304 rats, 314 guinea-pigs, 48 rabbits, 34 dogs, and 57 monkeys. The age of the animals was not specified. The exposed animals were evaluated for visible signs of toxicity, mortality, and haematological, biochemical, pathological, and body weight changes. There was apparent exposure-related mortality in guinea-pigs and monkeys. Mortality was 2/314, 2/45, 3/15, 3/15, and 7/15 in guinea-pigs and 1/57, 1/21, 0/9, 2/3, and 3/9 in monkeys in the 0, 20, 61, 101, or 189 mg/m³ exposure groups, respectively. The guinea-pigs died on the 3rd and 4th days of exposure, between the 3rd and 6th days of exposure, and between the 4th and 9th days of exposure in the 61, 101, and 189 mg/m³ groups, respectively. The monkeys died on days 39 and 47 and on days 26, 60, and 64 in the 101 and 189 mg/m³ groups, respectively. There were no visible signs of toxicity in any surviving animals. Because visible signs of toxicity were not observed and only non-life-threatening liver damage is apparent in this study (see below), the mortality data in guinea-pigs and monkeys are given no weight. Varying degrees of growth depression were found in all exposures, but growth depression was significant in all species only at 189 mg/m³. The test animals exhibited no significant haematological alterations, and serum urea nitrogen levels were within control limits in all exposures in which determinations were made. Significant elevations of serum glutamic–pyruvic transaminase and liver alkaline phosphatase activities were found in rats (a 3-fold and 1.75-fold increase, respectively) and guinea-pigs (a 7-fold and 2.4-fold increase, respectively) exposed to 189 mg/m³ (other species not tested), but not at 20 mg/m³ (enzyme levels at intermediate exposures not tested). Histopathological examination of liver from dogs, monkeys, and rats revealed damage at 189 mg/m³ (other species not examined). The effects observed included fatty metamorphosis, focal necrosis, haemosiderosis deposition, lymphocytic infiltration, bile duct proliferation, and fibrosis. The changes were most severe in dogs. Sections of kidney from all rats showed nuclear hypertrophy of the tubular epithelium. No detectable liver or kidney damage was observed in any species exposed to 101 mg/m³ or less. The NOAEL in this study is 101 mg/m³; the LOAEL is 189 mg/m³.

8.5.1 Oral

Bioassays for cancer by the oral route of exposure have been conducted in rats (Ponomarkov & Tomatis, 1980; NTP, 1982; Quast et al., 1983; Maltoni et al., 1985), mice (NTP, 1982), and trout (Hendricks et al., 1995). Some of these bioassays were conducted at an exposure below the maximum tolerated dose. The bioassay conducted by Maltoni et al. (1985) exposed the animals for only 1 year. The bioassay in rats (Quast et al., 1983) and the bioassay in mice (NTP, 1982) were well conducted, and both showed some toxicity in the liver at the highest exposure. Neither of these bioassays provides any significant evidence that 1,1-DCE is a carcinogen by the oral route of exposure.

Quast et al. (1983) summarized the results of a 2-year chronic toxicity and carcinogenicity study of 1,1-DCE conducted in Sprague-Dawley rats (6–7 weeks old). Humiston et al. (1978) reported detailed data from the study. There were 80 rats of each sex in the control group and 48 rats of each sex in each exposed group. The 1,1-DCE was incorporated in the drinking-water of the rats at nominal concentrations of 0, 50, 100, or 200 mg/litre. The time-weighted average exposure over the 2-year period was 7, 10, and 20 mg/kg body weight per day for males and 9, 14, and 30 mg/kg body weight per day for females. There were no significant differences among the groups in appearance and demeanour, mortality, body weight, food consumption, water consumption, haematology, urinalysis, clinical chemistry determinations, organ weights, or organ-to-body-weight ratios. After 1 year on study, there was no depletion of the non-protein sulfhydryl levels in the liver or the kidneys (Rampy et al., 1977). The only treatment-related effect observed in rats was minimal hepatocellular mid-zonal fatty change and hepatocellular swelling. In the male rats at the termination of the study, there was an increased incidence of minimal hepatocellular fatty change (control, 14/80; 50 mg/litre, 5/48; 100 mg/litre, 13/48; 200 mg/litre, 19/47) and minimal hepatocellular swelling (control, 0/80; 50 mg/litre, 1/48; 100 mg/litre, 2/48; 200 mg/litre, 3/47). The changes were statistically significant (P < 0.05) only in the 200 mg/litre group. In female rats at the termination of the study, there was an increased incidence of minimal hepatocellular fatty change (control, 10/80; 50 mg/litre, 12/48; 100 mg/litre, 14/48; 200 mg/litre, 22/48; statistically significant [P < 0.05] at 100 and 200 mg/litre) and minimal hepatocellular swelling (control, 3/80; 50 mg/litre, 7/48; 100 mg/litre, 11/48; 200 mg/litre, 20/48; statistically significant [P < 0.05] in all groups). No hepatocellular necrosis was evident at any exposure. In addition, there was no change in liver weight, no change in clinical chemistry measurements diagnostic for liver damage, and no other indication of abnormal liver function. Based on the minimal nature of the hepatocellular swelling reported by the authors, this effect is not considered biologically significant and is not an adverse effect in this study. The statistically significant hepatocellular mid-zonal fatty change, however, is considered a minimal adverse effect in this study. Accordingly, the NOAEL in male rats is 10 mg/kg body weight per day and the LOAEL is 20 mg/kg body weight per day; the NOAEL in female rats is 9 mg/kg body weight per day and the LOAEL is 14 mg/kg body weight per day. The US EPA conducted a BMD analysis for the results in female rats (Appendix 4). In female rats, the BMD₁₀ (BMD for a 10% response) is 6.6 mg/kg body weight per day, and the BMDL₁₀ is 4.6 mg/kg body weight per day.

8.5.2 Inhalation

Bioassays for cancer by the inhalation route of exposure have been conducted in rats (Lee et al., 1977, 1978; Viola & Caputo, 1977; Hong et al., 1981; Maltoni et al., 1985; Quast et al., 1986; Cotti et al., 1988), mice (Lee et al., 1977, 1978; Hong et al., 1981; Maltoni et al., 1985), and hamsters (Maltoni et al., 1985). None of these bioassays was conducted by a protocol that meets current standards. The major defects in most of these bioassays include exposure of the animals for 1 year and exposure at less than the maximum tolerated dose. The only bioassay showing some evidence of carcinogenicity was the study in Swiss-Webster mice (Maltoni et al., 1985). This study was conducted at or near the maximum tolerated dose, as animals exposed at 200 mg/m³ died after a few exposures.

Quast et al. (1986) and Rampy et al. (1977) reported results from studies that exposed male and female Sprague-Dawley rats (Spartan substrain, 86 animals per sex per group) to 1,1-DCE by inhalation 6 h/day, 5 days/week, for up to 18 months. Interim sacrifices

(4–5 animals per sex per group) occurred at 1, 6, and 12 months. Rats were exposed to 1,1-DCE concentrations of 40 or 160 mg/m³ for the first 5 weeks of the study. Based upon the absence of observable treatment-related effects among rats sacrificed after 1 month of exposure, the concentrations were increased to 100 and 300 mg/m³. Exposures were continued at these concentrations through the 18th month of the study. The surviving animals were then held without exposure to 1,1-DCE until 24 months. Cytogenetic evaluations were performed on a separate group of animals (four per sex) exposed to 0, 100, or 300 mg/m³ for 6 months. There were no exposure-related changes in mortality, appearance and demeanour, body weight, clinical chemistry determinations, haematological evaluations, urinalysis, or cytogenetic evaluation of bone marrow preparations. Minimal hepatocellular fatty change in the mid-zonal region of the hepatic lobule was observed in both male and female rats in the 100 and 300 mg/m³ groups at the 6-month interim sacrifice (male: control, 0/5; 100 mg/m³, 1/5; 300 mg/m³, 4/5; female: control, 0/5; 100 mg/m³, 2/5; 300 mg/m³, 4/5). The fatty change was also observed at the 12-month sacrifice, but there was no indication of progression of severity (male: control, 0/5; 100 mg/m³, 3/5; 300 mg/m³, 5/5; female: control, 0/5; 100 mg/m³, 5/5; 300 mg/m³, 5/5). At the 18-month sacrifice, the incidence of this change was no longer increased in male rats (control, 0/27; 100 mg/m³, 0/25; 300 mg/m³, 1/27). However, the change persisted in female rats (control, 0/16; 100 mg/m³, 6/29; 300 mg/m³, 7/20). In female rats, the fatty change was statistically significant (P < 0.05) only at the higher exposures. During the last 6 months of the study, after exposure had been discontinued, this effect was no longer discernible (male: control, 0/46; 100 mg/m³, 1/47; 300 mg/m³, 0/51; female: control, 0/49; 100 mg/m³, 0/46; 300 mg/m³, 1/48). Although the minimal hepatocellular mid-zonal fatty change is reversible and did not result in altered organ weight, clinical chemistry changes diagnostic for liver damage, or any obvious decrement in liver function, the fatty change in liver is considered a minimal adverse effect. Accordingly, the NOAEL in male rats in this study is 300 mg/m³ (the highest exposure tested). The NOAEL in female rats in this study is 100 mg/m³; the LOAEL is 300 mg/m³. The US EPA conducted a BMC analysis (Appendix 4). In female rats, using a conversion factor of 3.97, the BMC₁₀ (the BMC for a 10% response) is 59.9 mg/m³ and the BMCL₁₀ is 38.9 mg/m³, equivalent to 6.9 mg/m³ adjusted for continuous exposure (38.9 mg/m³ × 6/24 × 5/7).

Maltoni et al. (1985) conducted a carcinogenicity and toxicity study of 1,1-DCE in Swiss mice. Animals (9 or 16 weeks old) were exposed by inhalation to 0, 40, or 100 mg/m³ for 4 h/day, 4–5 days/week, for 52 weeks. Groups of animals exposed to 200 mg/m³ and higher in this study showed extreme toxicity after only a few exposures, causing termination of this portion of the bioassay. There were two control groups, one with 180 animals (90 of each sex) and the other with 200 animals (100 of each sex). The 40 mg/m³ group had 60 animals (30 of each sex). There were two groups exposed to 100 mg/m³, one with 60 animals (30 of each sex) and the other with 240 animals (120 of each sex). Following the 52-week exposure, animals were observed until spontaneous death (total duration 126 weeks). Body weight was measured every 2 weeks during the 52-week exposure and every 8 weeks thereafter. Full necropsy and histopathological examination were performed. There were no biologically significant changes in body weight. The exposed animals had a somewhat higher survival than controls. There was a statistically signifi-cant increase (P < 0.01) compared with controls in kidney adenocarcinomas in male mice at 100 mg/m³, but not in male mice at 40 mg/m³ or in female mice at either exposure. As reported in US EPA (1985), the incidence was 0/126 (0%), 0/25 (0%), and 28/119 (23.5%) in male mice in the combined control, 40 mg/m³, and combined 100 mg/m³ groups, respectively. Because of the study design, it is impossible to assess tumour development in relation to histopathological changes associated with typical age-related changes in the kidney. There was a statistically significant increase (P < 0.01) compared with control in mammary carcinomas in female mice at both exposures, but there was no clear exposure–response relationship. As reported in US EPA (1985), the incidence was 3/185 (1.6%), 6/30 (20%), and 16/148 (11%) in females in the combined control, 40 mg/m³, and combined 100 mg/m³ groups, respectively. There was also a statistically significant increase (P < 0.01) compared with control in pulmonary adenomas in both exposed groups, but there was no clear exposure–response relationship. As reported in US EPA (1985), the incidence was 6/153 (3.9%), 11/28 (39.3%), and 23/141 (16.3%) in males and 6/178 (3.4%), 3/30 (10%), and 18/147 (12.2%) in females in the combined controls, 40 mg/m³, and combined 100 mg/m³ groups, respectively. There were no pulmonary carcinomas in any mice. The incidence data are reported as the number of tumour-bearing animals compared with the number of animals alive when the first tumour was observed in that organ (kidney adenocarcinoma, 55 weeks; mammary tumour, 27 weeks; pulmonary adenoma, 36 weeks). The researchers discounted the significance of the mammary and pulmonary tumours.

8.5.3 Dermal

Van Duuren et al. (1979) evaluated the carcinogenicity of 1,1-DCE in male and female non-inbred Ha:ICR Swiss mice. Carcinogenicity was assessed in three types of tests: a dermal initiation-promotion assay; a repeated dermal application assay; and a subcutaneous injection assay. Vehicle, no-treatment, and positive control groups were included in the tests. In the initiation-promotion assay, 1,1-DCE was tested as a tumour-initiating agent with phorbol myristate acetate as the promoter. Thirty female mice were treated with 121 mg 1,1-DCE each. A significant increase (P < 0.005) was observed in skin papillomas (nine in eight mice). In the repeated dermal application assay, exposures of 40 and 121 mg per mouse were used. 1,1-DCE was applied to the back of the shaved animals (30 females per dose). No sarcomas were observed at the treatment site. Although 19 mice in the high-dose group and 12 in the low-dose group had lung tumours and 2 mice in the high-dose group had stomach tumours, the tumour incidence at both sites was not significantly different from controls (30 lung tumours and 5 stomach tumours). In the subcutaneous injection assay, the test animals were given weekly injections of 2 mg of 1,1-DCE. After 548 days on test, none of the animals injected with 1,1-DCE developed sarcomas at the injection site. 1,1-DCE showed initiating activity in the two-stage carcinogenesis experiments but was inactive as a whole-mouse dermal carcinogen and after subcutaneous injection.

There is a fairly extensive database on the genotoxicity of 1,1-DCE. A complete summary of the results of genetic toxicity testing is provided in Table 2. 1,1-DCE induced mutations in Salmonella typhimurium and Escherichia coli in the presence of an exogenous metabolic activation system. 1,1-DCE was also weakly mutagenic in Salmonella typhimurium TA 100 in the absence of a metabolic activation system. In Saccharomyces cerevisiae, 1,1-DCE induced reverse mutation and mitotic gene conversion in vitro and in a host-mediated assay in mice. In a single study in Saccharomyces cerevisiae, it induced aneuploidy in the presence and absence of metabolic activation. In vitro, gene mutations were increased in mouse lymphoma cells but not in Chinese hamster lung cells with or without an exogenous metabolic system. In a single study, 1,1-DCE induced sister chromatid exchanges in Chinese hamster lung cells in the presence of an exogenous metabolic activation system, but not in its absence. In single studies in vivo, 1,1-DCE did not induce micronuclei or chromosomal aberrations in bone marrow or in fetal erythrocytes of mice or in bone marrow of rats, and it did not induce dominant lethal mutations in mice or rats. A test for chromosomal damage in the mouse lymphoma system has not been conducted.

Table 2: Genetic and related effects of 1,1-dichloroethene.

Reitz et al. (1980) investigated the ability of 1,1-DCE to cause DNA alkylation, DNA repair, and DNA replication in liver and kidney of rats and mice. Male Sprague-Dawley rats (body weight 200–250 g) and male CD-1 mice (body weight 18–20 g) were exposed by inhalation for 6 h to 40 or 200 mg/m³. There was only a minimal increase in DNA alkylation in both rats and mice at 200 mg/m³. Similarly, DNA repair in kidney of mice was only minimally increased at 200 mg/m³. However, tissue damage (kidney nephrosis at 200 mg/m³, minimal effect at 40 mg/m³), an increase in DNA replication (7-fold increase in [³H]thymidine incorporation at 40 mg/m³, 25-fold increase at 200 mg/m³), and an increase in mitotic figures occurred. There was no observed histopathological damage or increased DNA replication in the liver of mice at 40 or 200 mg/m³. In rats, there was a small increase in DNA replication (2-fold increase in [³H]thymidine incorporation) in the kidney but no increase in liver at 40 mg/m³. This assay was not conducted in rats exposed to 200 mg/m³.

8.7.1 Effects on fertility

Nitschke et al. (1983) evaluated the reproductive and developmental toxicity of 1,1-DCE in Sprague-Dawley rats. Three generations of test animals were exposed to drinking-water containing nominal 1,1-DCE concentrations of 0 (initially 15 males and 30 females), 50, 100, or 200 mg/litre (initially 10 males and 20 females at each exposure). The authors provided no information on water consumption. This study was a companion study to Quast et al. (1983), using the same concentrations of 1,1-DCE in drinking-water. In Quast et al. (1983), the average exposure for females was 9, 14, and 30 mg/kg body weight per day. After 100 days of exposure, the rats were mated. In this three-generation study, there were no biologically significant changes in fertility index, average number of pups per litter, average body weight of pups, or pup survival at any exposure. Neonatal survival was decreased from concurrent control values in the F₂ and F_3a litters of dams ingesting 1,1-DCE from drinking-water. The survival indices, however, were within the range of control values for this strain of rats in this laboratory. The authors attributed the decreased survival index in F₂ to increased litter size at birth in dams exposed to 1,1-DCE. The apparent effect seen in the F_3a litters was not repeated in subsequent matings of the same adults to produce either the F_3b or the F_3c litters. The authors attributed the decreased survival in the F_3a litters as being due to chance. Histopathological examination of tissues of rats exposed to 1,1-DCE in the drinking-water in utero, during lactation, and post-weaning revealed slight hepatocellular fatty change and an accentuated hepatic lobular pattern of a reversible nature in the adult rats (data not reported, but the observation is consistent with that reported by Quast et al. [1983] in a chronic bioassay). These effects were observed in the 100 and 200 mg/litre groups in the F₁ generation and in all groups of the F₂ generation. The authors did not present incidence data or report statistical analysis. Exposure to 1,1-DCE in drinking-water at concentrations causing mild, dose-related changes in the liver did not affect the reproductive capacity of rats through three generations, which produced six sets of litters. The NOAEL for reproductive and developmental toxicity in this study is 200 mg/litre for exposure to 1,1-DCE in drinking-water (the highest exposure tested and about 30 mg/kg body weight per day).

8.7.2 Developmental toxicity

8.7.2.1 Oral

8.7.2.2 Inhalation

Siletchnik & Carlson (1974) investigated the effects of epinephrine on the cardiac sensitization of 1,1-DCE in male albino rats. The test animals (body weight 250–400 g) were exposed to 1,1-DCE at 0 or 100 000 ± 2400 mg/m³, and the dose of epinephrine was titrated to determine the minimum concentration needed to produce arrhythmias. A dose of 4 µg epinephrine/kg body weight failed to induce cardiac arrhythmias in air-exposed animals. However, the dose necessary to produce life-threatening arrhythmias was 2.0 µg/kg body weight following 58–61 min of exposure to 1,1-DCE, 1.0 µg/kg body weight following 64 min of exposure to 1,1-DCE, and 0.5 µg/kg body weight following 67–80 min of exposure to 1,1-DCE. The cardiac sensitization was found to be completely reversible upon discontinuation of the exposure to 1,1-DCE.

The various observations on the toxicity and metabolism of 1,1-DCE indicate that cytotoxicity is associated with cytochrome P450-catalysed metabolism of 1,1-DCE to reactive intermediates that bind covalently to cellular macromolecules. The extent of binding is inversely related to loss of GSH, so that severities of tissue damage parallel the decline in GSH (Forkert & Moussa, 1991; Moussa & Forkert, 1992). Hepatotoxicity is also exacerbated by treatments that diminish GSH (Jaeger et al., 1974, 1977b; McKenna et al., 1978b; Andersen et al., 1980). Thus, the responses to 1,1-DCE at low doses with little depletion of GSH are expected to be very different from the responses at high doses causing substantial GSH depletion. The targets of toxicity are centrilobular hepatocytes and bronchiolar Clara cells (Forkert et al., 1985), cell types that are rich in CYP2E1 (Forkert et al., 1991; Forkert, 1995). Immunohistochemical studies showed formation of DCE-epoxide–cysteine protein adducts within the centrilobular hepatocytes and Clara cells (Forkert, 1999a,b). Following short-term exposure of mice to high concentrations of 1,1-DCE, the degree of cellular damage in Clara cells and hepatocytes in various strains of mice correlates with the extent of formation of DCE-epoxide and the level of CYP2E1 in the tissue (Forkert, 2001; Forkert & Boyd, 2001; Forkert et al., 2001). In combination, these findings indicate that DCE-induced toxicity is associated with the formation and reactivity of the DCE-epoxide within the target centrilobular hepatocytes and Clara cells (Forkert, 2001).

Speerschneider & Dekant (1995) investigated the metabolic basis for the species- and sex-specific nephrotoxicity and tumorigenicity of 1,1-DCE. In kidney microsomes from Swiss-Webster male mice, the rate of oxidation of 1,1-DCE depended on the hormonal status of the animals. Oxidation of 1,1-DCE was decreased by castration and restored when the castrate was supplemented with exogenous testosterone. In kidney microsomes from naive female mice, the rate of oxidation of 1,1-DCE was significantly lower than in males, but could be increased by administration of exogenous testosterone. Using an antibody to rat liver CYP2E1, the researchers showed expression of a cross-reacting protein in male mouse kidney microsomes that was regulated by testosterone and correlated with the ability to oxidize 1,1-DCE and other substrates for CYP2E1 (e.g., p-nitrophenol and chlorozoxazone). The researchers also showed that different strains of mice express different levels of CYP2E1. The strains most sensitive to the effects of 1,1-DCE express greater levels of CYP2E1. Nephrotoxicity in Swiss-Webster mice after inhalation of 1,1-DCE was observed in males and in females treated with exogenous testosterone, but not in naive females. In kidney microsomes obtained from both sexes of rats and in six samples of human kidney from male donors, no p-nitrophenol oxidase activity was detected. Other research groups have also reported the absence of detectable CYP2E1 in human kidney tissue (Amet et al., 1997; Cummings et al., 2000).

Ott et al. (1976) investigated the health records of 138 employees occupationally exposed to 1,1-DCE in processes not involving vinyl chloride. The individuals included in the study had worked in experimental or pilot plant polymerization operations, in a monomer production process as tankcar loaders, or in a production plant manufacturing a monofilament fibre. Overall, there were no significant differences in haematology, clinical chemistry, or mortality between the exposed cohort and the controls. This study is too limited to derive useful information on the toxicity of 1,1-DCE to humans, because there was an inadequate number of subjects and only a limited number of end-points examined.

There are only limited data on the effects of 1,1-DCE on organisms in aquatic and terrestrial environments.

The data on the effects of 1,1-DCE on aquatic species are summarized in Table 3. A major limitation in most of these studies is the lack of prevention of the volatilization of 1,1-DCE. In addition, increase in biomass, not growth rate, was measured in the studies with aquatic plants.

Table 3: Aquatic toxicity of 1,1-dichloroethene.