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Screening Assessment Report
Chemical Abstracts Service Registry Number
(PDF Format - 365 KB)
Table of Contents
- Identity of Substances
- Physical and Chemical Properties
- Releases to the Environment
- Environmental Fate
- Persistence and Bioaccumulation Potential
- Potential to Cause Ecological Harm
- Potential to Cause Harm to Human Health
Pursuant to paragraphs 68(b) and 68(c) of the Canadian Environmental Protection Act, 1999 (CEPA 1999), the Ministers of the Environment and of Health have conducted a screening assessment of Ethene, 1,1-dichloro- (1,1-dichloroethene), Chemical Abstracts Service Registry Number 75-35-4. 1,1-Dichloroethene was identified as a high priority for assessment of human health risk because it has been classified by other agencies on the basis of carcinogenicity.
1,1-Dichloroethene is a chlorinated organic compound that was used in solvents and as an intermediate in a variety of chemical processes. Based on a survey issued under section 71 of CEPA 1999, between 10 and 100 tonnes of 1,1-dichloroethene were manufactured and imported into Canada in 2000. However, it is no longer produced or imported into Canada. Small amounts of 1,1-dichloroethene are created unintentionally in several industrial processes; most of this material is reformed into other substances within the facilities.
Globally, 1,1-dichloroethene is used primarily as an intermediate in the manufacture of polyvinylidene chloride polymers and copolymers, which may in turn be used in a variety of end products such as food plastic wrap, carpet latex backing, fire- and ignition-resistant clothing, vapour barriers for insulation, paper and board coatings, and photographic film. 1,1-Dichloroethene may persist as an unintended manufacturing residue in some of these items that may be present in Canadian commerce. 1,1-Dichloroethene may also be used in the production of hydrochloro-fluorocarbons, chloroacetyl chloride, and latex and resins, as an aid in ore flotation, as a solvent in paint and varnish remover, and as a vapour degreaser and industrial cleaning agent.
1,1-Dichloroethene is a reportable substance to the National Pollutant Release Inventory (NPRI); reported releases have steadily declined from 87 kg in 2000 to 1 kg in 2003. Since 2003, no companies have reported releases of 1,1-dichloroethene to the NPRI.
1,1-Dichloroethene can be released during the breakdown of polyvinylidene chloride products and during the abiotic and biotic decomposition of the drycleaning and degreasing solvents 1,1,1-trichloroethane, 1,1,2,2-tetrachloroethene (tetrachloroethene or perchloroethylene), 1,1,2-trichloroethene and 1,2-dichloroethane. Many of these solvents are no longer used in Canada and disposal practices have improved, such that significant new sources of 1,1-dichloroethene in groundwater and soil are unlikely. Recent monitoring data show that 1,1-dichloroethene is present in urban air at very low concentrations, often just above analytical detection limits.
Based on experimental and modelled data for 1,1-dichloroethene, the substance is expected to degrade readily in air, soil and water. Based on its physical and chemical properties and predictions from bioaccumulation models, the substance is not expected to bioaccumulate in aquatic organisms. Therefore, 1,1-dichloroethene does not meet criteria for persistence and bioaccumulation potential as set out in the Persistence and Bioaccumulation Regulations. In addition, available empirical ecotoxicity data (for mammals, aquatic and terrestrial plants, invertebrates and vertebrates) indicate that 1,1-dichloroethene is not highly hazardous to non-human organisms.
Based on the low concentrations measured or expected in the Canadian environment and low hazards of the substance there is very low likelihood of ecological harm from 1,1-dichloroethene in the Canadian environment, especially given the low persistence and potential for bioaccumulation of 1,1-dichloroethene and decreasing uses and releases.
Based on the information available, it is concluded that 1,1-dichloroethene is not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment on which life depends.
The general population exposure to 1,1-dichloroethene is mainly from indoor air and possibly food. A comparison of the lowest critical inhalation effect level for non-cancer effects with the highest median concentration of 1,1-dichloroethene in indoor air from recent studies in Canada, and a comparison of the critical oral effect level for non-cancer effects and the upper-bounding estimate of daily intake, result in margins of exposure which are considered adequate to address uncertainties in the health effects and exposure databases for chronic non-cancer effects. Additionally, for non-cancer effects based on a less than chronic period, a comparison of the lowest critical inhalation effect level for such periods with the highest 95th percentile concentration of 1,1-DCE in air from recent studies in Canada results in a margin of exposure which is considered adequate to address uncertainties in the health effects and exposure databases.
A critical effect for characterization of risk to 1,1-dichloroethene is carcinogenicity. Following lifetime inhalation of 1,1-dichloroethene at high concentrations, mice developed renal tumours. A comparison of the critical effect level for cancer and the upper-bounding estimate of daily intakes results in margins of exposure which are considered adequate to address uncertainties in the health effects and exposure databases for cancer effects. Additionally, available information suggests that the mode of tumour induction in experimental animals may not be relevant to humans.
Based on the information available, it is concluded that 1,1-dichloroethene is not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.
Based on available information for environmental and human health considerations, it is concluded that 1,1-dichloroethene does not meet the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999.
This screening assessment was conducted pursuant to section 68 of the Canadian Environmental Protection Act, 1999 (CEPA 1999) (Canada 1999). This section of the Act authorizes the Ministers of the Environment and of Health to conduct assessments of substances to determine whether they meet or may meet the criteria set out in section 64 of the Act.
Screening assessments focus on information critical to determining whether a substance presents, or may present, a risk to the environment or to human health, according to the criteria set out in section 64 of CEPA 1999. Screening assessments examine scientific information and develop conclusions by incorporating a weight-of-evidence approach and precaution.  Ethene, 1,1-dichloro- (1,1-DCE) (Chemical Abstracts Service Registry Number 75-35-4) was identified as a priority for assessment of human health risk because it has been classified by other agencies on the basis of carcinogenicity.
The 2005 State of the Science Report for a Screening Health Assessment of 1,1-DCE was posted on the Health Canada website on November 3, 2005. The State of the Science Report for a Screening Health Assessment was externally reviewed by staff of Toxicology Advice and Consulting Limited and Toxicology Excellence in Risk Assessment, V.C. Armstrong (consultant) and P. Price (The Lifeline Group Inc.) for adequacy of data coverage and defensibility of the conclusions. The external comments were taken into consideration in drafting the State of the Science Report. The health screening assessment included here is an update of the State of the Science Report and, since limited new information was available, has not been subsequently peer reviewed.
This screening assessment includes consideration of information on chemical properties, hazards, uses and exposure. Data relevant to the screening assessment of this substance were identified in original literature, review and assessment documents, stakeholder research reports and from recent literature searches, up to December 2009 for ecological sections of the document and September 2009 for human health sections of the document. Also, data initially reported from draft reports of two Health Canada studies were updated in this assessment based on finalized reports published in 2010. In addition, an industry survey was conducted in 2000 through a Canada Gazette notice issued under authority of section 71 of CEPA 1999. This survey collected data on the Canadian manufacture and import of the DSL pilot project substances (Environment Canada 2001a). Key studies were critically evaluated; modelling results may have been used to reach conclusions.
Evaluation of risk to human health involves consideration of data relevant to estimation of exposure (non-occupational) of the general population, as well as information on health hazards. Decisions for human health are based on the nature of the critical effect and/or margins between conservative effect levels and estimates of exposure, taking into account confidence in the completeness of the identified databases on both exposure and effects, within a screening context. The screening assessment does not represent an exhaustive or critical review of all available data. Rather, it presents a summary of the critical information upon which the conclusion is based.
This final screening assessment was prepared by officials in the Existing Substances Programs at Health Canada and Environment Canada. As mentioned above, the State of the Science Report for a Screening Health Assessment was previously externally reviewed. The ecological component of this assessment has undergone external written scientific peer review/consultation and comments received were considered in the production of this report. Comments on the technical portions relevant to human health were received from BIBRAToxicology Advice and Consulting, The LifeLine Group, and two independent consultants (Dr. Marla Shapiro and Dr. V.C. Armstrong). Additionally, the draft of this screening assessment was published on December 16, 2011 and subject to a 60-day public comment period. Although external comments were taken into consideration, the final content and outcome of the screening assessment remain the responsibility of Health Canada and Environment Canada.
The critical information and considerations upon which this assessment is based are summarized below.
Ethene, 1,1,-dichloro-, also known as 1,1-dichloroethene, will be referred to in this assessment by its acronym, 1,1-DCE. Information on its identity is presented in Table 1.
|DSL name||Ethene, 1,1-dichloro-|
|NCI names||Ethene, 1,1-dichloro- (TSCA, DSL, AICS, SWISS, PICCS, ASIA-PAC, NZIoC)|
1,1-Dichloroethene (ENCS, ECL)
Vinylidene chloride (ENCS, PICCS)
|Other names||Sconatex; Diofan A 565S; Ethene, 1,1-dichloro; Ethylene, 1,1-dichloro-; F 1130a; HCC 1130a; Iso-dichloroethylene; R 1130a; UN 1303; UN 1303 (DOT); VDC; Vinylidene dichloride|
|Major chemical class or use||Alkenes|
|Major chemical sub-class||Halogenated alkene|
|Molecular mass||96.94 g/mol|
 National Chemical Inventories (NCI). 2006: AICS (Australian Inventory of Chemical Substances);ASIA-PAC (Asia-Pacific Substances Lists); DSL (Domestic Substances List); ECL (Korean Existing Chemicals List); EINECS (European Inventory of Existing Commercial Chemical Substances); ENCS (Japanese Existing and New Chemical Substances); NZIoC (New Zealand Inventory of Chemicals); PICCS (Philippine Inventory of Chemicals and Chemical Substances); SWISS (Swiss Giftliste 1 and Inventory of Notified New Substances); and TSCA (Toxic Substances Control Act Chemical Substance Inventory).
 Simplified Molecular Input Line Entry Specification.
Physical and Chemical Properties
Physical and chemical properties of 1,1-DCE are summarized in Table 2 below.
1,1-DCE is a volatile substance that will exist as a liquid at most environmental temperatures. It will evaporate from most substrates including water, although it will also exist in a dissolved state in water. The relatively low Kow and Koc values indicate that it will not bind tightly to organic matter in the environment.
|Melting point (°C)||Experimental||-123||–||PhysProp 2009|
|Boiling point (°C)||Experimental||31.6||–||PhysProp 2009|
|Vapour pressure (Pa)||Experimental||8 × 104||25°C||PhysProp 2009|
|Henry’s Law constant (Pa·m3/mol)||Experimental||2644||24°C||PhysProp 2009|
|Octanol–water partition coefficient (log Kow)|
|Organic carbon-water partition coefficient (log Koc)|
|Estimated||1.85 (Kow method)||–||KOCWIN 2008|
|Water solubility (mg/L)||Experimental||2420||25°C||PhysProp 2009|
|Rate constant for gas-phase reaction with hydroxyl radical (kOH) cm3/ molecule per second||Experimental||1.1 × 10−11||25°C||PhysProp 2009|
1,1-DCE is an anthropogenic substance that has not been identified to occur naturally (BUA 1998; WHO 2003a). Commercial production involves the dehydrochlorination of 1,1,2-trichloroethane in the presence of excess base or by thermal decomposition of methyl chloroform (1,1,1-trichloroethane) (Grosjean 1991; WHO 2003a). These production processes consist primarily of closed system operations in industrial settings (Williams et al. 2006). 1,1-DCE is also a byproduct in a process used to manufacture hydrogen chloride (Environment Canada 2001b).
An industry survey was conducted for the 2000 calendar year under section 71 of CEPA 1999 (Environment Canada 2001a). The Notice with Respect to Certain Substances on the Domestic Substances List (DSL) applied to any person who, during the 2000 calendar year, manufactured or imported 1,1-DCE, whether alone or in a mixture or in a product, in a total quantity greater than 10 000 kg. Based on the survey, between 10 and 100 tonnes of 1,1-DCE were respectively reported to be manufactured and imported into Canada in 2000 (Environment Canada 2001b). During that year, 1,1-DCE was used in a sealer solvent and, as a by-product produced during the manufacture of 1,2-dichloroethane, to manufacture hydrochloric acid (Environment Canada 2001b). Follow-up with an importing company indicated that use of 1,1-DCE in a sealer solvent is no longer ongoing (2004 email from the importing company to Existing Substances Branch, Environment Canada; unreferenced). In addition, use of 1,1-DCE in the manufacture of hydrochloric acid, and its production as a by-product during the manufacture of 1,2-dichloroethane, is no longer ongoing (Dow 2006a). The total quantity in commerce in 2000 of 10 to 100 tonnes was similar to the total quantity in commerce during the period of DSL compilation (1984–1986) of 31 tonnes (Environment Canada 2000).
1,1-DCE is also a product of incomplete combustion of some chlorinated solvents, and thus incineration of hazardous waste presents a possible environmental source of 1,1-DCE (Fuerst et al. 1989). One study indicated that 1,1-DCE and methyl chloride were by far the principal incomplete combustion products for 1,1,1-trichloroethane, with a concentration of 1,1-DCE in incineration exhaust reported as > 200 ppb (Fuerst et al. 1989). However, 1,1,1-trichloroethane production was phased out by 2005 under the Montreal Protocol (Environment Canada 2003).
1,1-DCE was detected in sewage sludge in 3% of 436 samples from the US, at a concentration range of 1–14 000 µg/L in one study (Burns and Roe 1982). In another study, 1,1-DCE was detected at a mean concentration of 7.97 mg/kg dry weight (214 µg/L wet volume) in twelve digested sewage sludge samples (Wilson et al. 1994). Sewage sludge may represent a possible source of environmental 1,1-DCE in Canadian soil, though no recent US data or data on presence in sludge in Canada are available.
Based on responses to a survey issued under section 71 of CEPA 1999 (Environment Canada 2001a), in the 2000 calendar year over 10 tonnes were used in a sealer solvent and over 10 tonnes were used, as a by-product produced during the manufacture of 1,2-dichloroethane, to manufacture hydrochloric acid (Environment Canada 2001b). However, as of 2004, there is no record of 1,1-DCE being used in sealants in Canada (2004 email from the importing company to Existing Substances Branch, Environment Canada; unreferenced). In addition, the chlor-alkali and direct chlorination ethylene dichloride plants that manufactured 1,2-dichloroethane and hydrochloric acid in Fort Saskatchewan, Alberta, were shut down in October 2006 due to economic considerations (Dow 2006a).
1,1-DCE is not expected to be present in cosmetic products in Canada, as it is not listed as an ingredient in the Cosmetic Notification System database (CNS 2009). 1,1-DCE is currently regulated in cosmetic products in Canada as it is within the category of dichloroethylenes, or acetylene chlorides, as specified on the Health Canada cosmetic ingredient hotlist (Health Canada 2009a). There are no registered pesticides that contain 1,1-DCE as an active ingredient or formulant in Canada (PMRA 2007), and 1,1-DCE is not a permitted food additive under Division 16 of the Food and Drug Regulations (Canada 1978).
1,1-DCE is not listed in the Drug Products Database, the Therapeutic Products Directorate’s internal Non-Medicinal Ingredients Database, the Natural Health Products Ingredients Database or the Licensed Natural Health Products Database as a medicinal or non-medicinal ingredient in pharmaceutical drugs, natural health products or veterinary drugs (DPD 2010; NHPID 2010; LNHPD 2010; TPD NMID 2010). The International Conference on Harmonization Guideline Q3C(R4) (ICH 2009)--which is adopted by the Therapeutic Products Directorate (Health Canada 1999) and the Natural Health Products Directorate (Health Canada 2007)--and the International Cooperation on Harmonisation Guideline 18 (VICH 2000)--which is adopted by the Veterinary Drugs Directorate (Health Canada 2001)--list 1,1-DCE as a Class 1 residual solvent (i.e. solvent that should be avoided). As such, it should not be used in the manufacture of drug substances; however, if use is unavoidable in the manufacture of drug or veterinary medicinal products with significant therapeutic advances, the concentration limit for residual 1,1-DCE is 8 ppm.
Globally, 1,1-DCE is used primarily as an intermediate in the manufacture of polyvinylidene chloride (PVDC) polymers and copolymers. Consumer products where PVDC is used include flexible plastic films in the food industry, vapour barriers for insulation, carpets, awnings, piping, coatings for steel pipes, adhesives and photographic film (US EPA 2002a, 2003; WHO 2003a). PVDC and its copolymers are used as flame retardants in paper and board coatings (IPCS 1990). In these end products, 1,1-DCE is only present at low levels. For example, monomer residual levels are reported to be < 5 ppm in carpet latex, photographic film coating, flame retardant and ignition-resistant fibres for industrial clothing, food packaging, and as a component of a compound used in the production of automotive interior foam (US EPA 2002a). However, monomer levels are reported to be higher in an assessment by the World Health Organization released in 2003. 1,1-DCE was reported at a level of < 100 mg 1,1-DCE per kg article in photographic film coating, flame-retardant fibres for clothing and outdoor awnings, and PVDC-fluorinated copolymers for application on textiles (WHO 2003a). The lower levels of 1,1-DCE reported in the review published by the United States Environmental Protection Agency likely resulted from the effects of further processing of articles on reducing the concentration of monomer in the final consumer product (WHO 2003a). The degree to which these end products are present in the Canadian marketplace is unknown.
Other global uses of 1,1-DCE identified in literature searches were as a captive intermediate in the production of hydrochlorofluorocarbons (HCFC-141b and HCFC-142b) and hydrofluorocarbons (HFC 236fa), for use as refrigerants and in fire extinguishers; in chloroacetyl chloride and homo-, co- and terpolymers (latex and resin) (Connor et al. 1998; WHO 2003a), as an aid in ore flotation, as a solvent in paint and varnish remover, as a degreaser, and as an industrial cleaning agent (WHO 2003b). In addition, 1,1-DCE was formerly used as an anaesthetic (WHO 2003b).
Releases to the Environment
Based on responses to a survey issued under section 71 of CEPA 1999, total releases to the environment reported by one notifier were between 10 and 100 tonnes in the 2000 calendar year (Environment Canada 2001b). The section 71 survey did not require specification of the medium of release or of the form of substance released (e.g. pure substance, in a mixture or product). However, the notifier did indicate that the releases occurred as point sources at facilities where the 1,1-DCE was being used as a manufacturing intermediate or being processed for disposal, and that at the time of consumer use, releases to environmental media would no longer occur. As both notifiers that used 1,1-DCE in 2000 are not current users, point releases of 1,1-DCE are assumed to have substantially decreased.
Although 1,1-DCE is a reportable substance to the Canadian National Pollutant Release Inventory (NPRI), no releases were reported after the 2003 calendar year (NPRI 2009). According to the NPRI, releases in 2000 were 87 kg, in 2001 they were 25 kg, in 2002 they were 4 kg, and in 2003 they were 1 kg. Releases were typically from a single company, but usually not the same one.
Historically, there have been releases of 1,1-DCE from industrial sources to surface waters in Canada. The Ontario Ministry of the Environment (OME) reported 1,1-DCE emissions of 0.376 kg/day to the St. Clair River from industrial effluents in 1986–1987 (OME 1991a). The average concentration of 1,1-DCE in process effluents at seven Ontario petroleum refineries was 0.180 μg/L for 43 analyses (detection limit not stated) over a six-month period in 1989 (OME 1991b). In 1989–1990, Ontario chemical manufacturers discharged to Ontario rivers an average of 1.75 kg/day of 1,1-DCE (OME 1992). Over six months in 1990, 1,1-DCE was released at an average concentration of 4.47 μg/L (six analyses) at a company in Thorold, Ontario (OME 1991c). However, 1,1-DCE was not routinely monitored under Ontario’s Municipal/Industrial Strategy for Abatement (MISA) program. The average total discharge of 1,1-DCE from 48 industrial sites along the St. Lawrence River in Quebec in 1992 was 0.136 kg/day (MENVIQ 1993).
In addition to being released into the environment during its manufacture and use, 1,1-DCE can also be released during the breakdown of PVDC products and during the abiotic and biotic decomposition of the drycleaning and vapour degreasing solvents 1,1,1-trichloroethane, tetrachloroethene (perchlorethylene), 1,1,2-trichloroethene and 1,2-dichloroethane due to poor disposal practices (OME 2001; Klier et al. 1999; IPCS 1990; ATSDR 1994; US EPA 1995). The formation of 1,1-DCE from these types of sources is highly variable, often depending on oxidative conditions in contaminated groundwater and landfills. Movement of these groundwater plumes below residences is a potential source of 1,1-DCE vapour intrusion to indoor air (Williams et al. 2006). For example, 1,1-DCE has been detected in groundwater in five studies (Appendix 1, Table A3). However, many of these solvents are no longer used in Canada and disposal practices have improved, such that significant new sources of 1,1-DCE in groundwater and soil are unlikely.
The results of Level III fugacity modelling (EQC 2003; Table 3) indicate that, if the chemical were released solely to air, the majority would remain in air. If released to water, the majority would remain in water with most of the rest partitioning to air. If released to soil, less than half would remain in soil while most of the rest would partition to air.
|Substance released to:||Percentage of substance partitioning into each compartment (%)|
Due to the very high vapour pressure of 1,1-DCE (66 000 Pa at 20°C), a high proportion will ultimately partition to the atmosphere (WHO 2003a) despite its relatively high water solubility (2390 mg/L). The water-air partition coefficient of 0.16 reported by Pearson and McConnell (1975) also implies that most 1,1-DCE will partition to air and only a small amount will remain in water.
Persistence and Bioaccumulation Potential
Empirical and modelled data for the degradation of 1,1-DCE in different media are presented in Tables 4a and 4b, respectively.
Gas-phase oxidation with photochemically produced hydroxyl radicals (OH•) is expected to be the most significant process for atmospheric removal, as this reaction is much faster than other atmospheric reactions (such as reactions with NO3-, ozone, and peroxy radicals), and the OH• is ubiquitous in air. Therefore, an atmospheric half-life was calculated for 1,1-DCE based on the following equation from Leifer (1993):
Half-life = 0.693/[(reaction rate)*(OH• concentration)*(43 200 seconds/12-hr day)]
Using a reaction rate constant of 1.09 × 10-11 (a unit-weighted average of the reaction rates reported by Atkinson ) and a default OH• concentration of 1.5 × 106, this calculation produces an estimated tropospheric half-life of 0.98 days for 1,1-DCE. The primary reaction products for this reaction include formaldehyde (CH2O), phosgene (COCl2) and hydroxyacetyl chloride (CH2ClCOOH) (Grosjean 1991). Other estimates of half-lives in air resulting from oxidation reactions with hydroxyl radicals were generally under 2 days (Table 4a).
1,1-DCE is a known precursor of formaldehyde in air following its degradation via oxidative reactions with hydroxyl radicals. Formaldehyde appears on the List of Toxic Substances in Schedule 1 of CEPA 1999.
Hydrolysis is not a significant degradation pathway for 1,1-DCE (Table 4a). Modelled half-lives for biodegradation reactions in water are estimated to be 28 to 180 days (Table 4b). The primary and ultimate biodegradation results are both ≤ 182 days in water. Due to its very high vapour pressure, the most important process for removal of 1,1-DCE from water is considered to be volatilization. Half-life values for volatilization from surface water bodies calculated by Mabey et al. (1981), and estimated in this assessment using the HENRYWIN (2008) model, ranged from 2.9 hours to 6 days. Measured biodegradation values in soil were 10 days (Ryan et al. 1988). Estimates of the half-life of 1,1-DCE in soil range from 28 days to 180 days and the estimated half-life in sediment is 150 days (BIOWIN 2009).
|Medium||Fate process||Degradation value||Degradation endpoint/units||Reference|
|Air||Photo-oxidation||0.46 days||Half-life||INERIS 2003|
|0.67 days||Grosjean 1991|
|0.98 day||Atkinson 1989|
|2 days||Brown et al. 1975|
|Reactions with NO3- at night||19 days||Grosjean 1990|
|Photolysis||56 days||Pearson and McConnell 1975|
|Ozone reaction||10 years||Grosjean 1990|
|Reaction with peroxy radicals||22 years||Brown et al. 1975|
|Water||Hydrolysis||6–9 months||Half-life||Cline and Delfino 1987|
|Hydrolysis (neutral to slightly basic pH)||1.2 × 108 years||Half-life||Jeffers et al. 1989|
|Soil||Biodegradation||< 10 days||Half-life||Ryan et al. 1988|
|Media||Fate process||Model result and prediction||Extrapolated half-life (days)||Reference|
|Air||Photo-oxidation||4.7 days||≥ 2||AOPWIN 2008|
|0.41–4.1 days||Howard et al. 1991|
|Ozone reaction||219 days||≥ 2||US EPA 1985|
|320 days||≥ 2||AOPWIN 2008|
|Water||Biodegradation MITI Linear||0.48 (not readily degradable)||≥ 182||BIOWIN 2009|
|Primary biodegradation||Days–weeks||≤ 182||BIOWIN 2009|
|Ultimate biodegradation||Weeks–months||≤ 182||BIOWIN 2009|
|Biodegradation||28–180 days||≤ 182||Howard et al. 1991|
|Volatilization||0.12–6 days||≤ 182||Mabey et al. 1981;|
|Groundwater||Anaerobic biodegradation||56–132 days||≤ 182||Howard et al. 1991|
|Anaerobic biodegradation||0.66 (degrades fast)||≤ 182||BIOWIN 2009|
|Anaerobic biodegradation (simulated groundwater environment)||5–6 months||≤ 182||Barrio-Lage et al. 1986|
|Soil||Biodegradation half-life||37.5 days||≤ 182||BIOWIN 2009|
|Biodegradation||28–180 days||≤ 182||Howard et al. 1991|
|Sediment||Biodegradation half-life||150 days||≤ 365||BIOWIN 2009|
Fugacity modelling with the model TaPL3 (Beyer et al. 2000; TaPL3 2000) has been used to estimate a characteristic travel distance for 1,1-DCE of 524 km using the predicted half-life in air of 0.98 days. This is less than the 700 km criteria for long-range transport in air; 1,1-DCE is therefore considered to have a low potential for long-range transport.
Based on the information available, 1,1-DCE does not meet the criterion for persistence as set out in the Persistence and Bioaccumulation Regulations (Canada 2000).
Potential for Bioaccumulation
Bioaccumulation is expected to be low based on the octanol/water partition coefficient of 2.1 (Hansch et al. 1995) 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).
A bioaccumulation factor of 0.96 was estimated for 1,1-DCE with the BCFBAF (2008) model. This model includes the Arnot-Gobas bioaccumulation model for mid-trophic level bioaccumulation in fish with metabolism considered.
Based on information available, 1,1-DCE does not meet the bioaccumulation criteria (BAF or BCF greater than 5000) set out in the Persistence and Bioaccumulation Regulations (Canada 2000).
Potential to Cause Ecological Harm
Ecological Exposure Assessment
In a recent study of ambient air levels conducted as part of the ongoing National Air Pollution Surveillance (NAPS) network, 1,1-DCE was not detected in 1896 samples, at a detection limit of 0.026 μg/m3, among 43 Canada-wide sites during the collection period of January to December 2008 (NAPS 2008). Another recent study, the Regina Indoor Air Quality Study (2007), detected 1,1-DCE at a maximum concentration of 0.014 μg/m3 in outdoor air of residential backyards during the summer of 2007 in Regina, Saskatchewan, with a detection limit of 0.012 μg/m3(Health Canada 2010a). In a similar study conducted in Windsor, Ontario, the Windsor Exposure Assessment Study (2005-2006), 1,1-DCE was not detected in any sample of outdoor air above detection limits of 0.152 µg/m3 and 0.046 µg/m3 for 2005 and 2006 respectively (Health Canada 2010b).
Annual mean air concentrations of 1,1-DCE were obtained for 33 sites across Canada for 2004 as well (NAPS 2008). Montréal had the highest annual mean concentration, 0.016 µg/m3, while Windsor had the lowest, 0.011 µg/m3. The only rural air monitoring site was located at Simcoe, Ontario, which had an annual mean concentration of 0.012 µg/m3, which was the same as the annual mean urban concentration. There were no statistically significant differences in the annual mean air concentrations between the various monitoring locations.
A separate four-week study was conducted in 2005 by the Air Quality Research Branch (AQRB) and the Canadian Meteorological Centre of Environment Canada in central Alberta. Data were collected from existing air quality stations and aircraft-mounted instruments. In the summer of 2005, there were only two samples with measured concentrations of 1,1-DCE above the detection limit of 0.011 µg/m3, and both concentrations were below 0.05 µg/m3 (2005 email from an Environment Canada Meteorological Service program manager to Ecological Assessment Division, Environment Canada; unreferenced).
In 2004 and 2005, Environment Canada and the Fort Air Partnership conducted an air monitoring program that included measuring 1,1-DCE near an industrial area in Fort Saskatchewan, Alberta, as well as at surrounding locations. A similar monitoring program was conducted by Environment Canada near an industrial area in North Vancouver, British Columbia. Air samples were taken from ten sites once every six days between September 2004 and July 2005. Mean monthly concentrations and an overall average concentration were calculated for each site. All monthly means in Fort Saskatchewan and North Vancouver were below the detection limit of 0.011 µg/m3. The highest monthly mean concentration, 0.014 µg/m3, was observed at a site 5 km north of Fort Saskatchewan (2005 email from an Environment Canada risk manager to Ecological Assessment Division, Environment Canada; unreferenced).
Monitoring by the Ontario Ministry of Environment and Energy in the late 1980s to early 1990s found the maximum concentration of 1,1-DCE measured over 30 minutes to be 0.81 μg/m3 at a hazardous waste facility, 0.25 to 0.65 μg/m3 in industrial areas, and 0.68 to 5.7 μg/m3 at landfills in the greater Toronto area (OME 1991d; OMEE 1997).
The atmospheric concentrations of 1,1-DCE in the vicinity of a hypothetical industrial facility were estimated using the air dispersion model SCREEN3 (US EPA 2006). It is a single-source Gaussian plume model that provides maximum 1-hour concentrations for point, area, flare and volume sources at a receptor height. Using a reasonable worst-case scenario for an industrial release of 10 tonnes over a year, the highest estimated 1-hour concentration of 1,1-DCE was 458.5 µg/m3 at a distance of 52 m from the centre of the model facility. A concentration of 3.1 µg/m3 of 1,1-DCE was estimated at a distance of 5 km from the facility.
The detection limit of 1,1-DCE in the recent NAPS survey, 0.026 μg/m3, was used as the predicted environmental concentration (PEC) in deriving risk quotients for air. These are the most recent Canadian data, the concentration is the most conservative of the recent studies, the study is expansive across Canada and the sample size is large (n = 1896).
The highest atmospheric concentration of 458.5 µg/m3(0.459 mg/m3) derived from the SCREEN3 model was selected as the PEC for use in deriving a risk quotient for a reasonable worst-case industrial scenario.
Concentrations of 1,1-DCE have been detected in groundwater samples associated with landfills. In the 1980s, levels ranging from 0.09 to 60 µg/L were detected in 43% of groundwater samples beneath the Gloucester landfill near Ottawa (Lesage et al. 1990). 1,1-DCE was not known to have been disposed of at the site but it is a known degradation product of tetrachloroethylene and 1,1,1-trichloroethane, which were disposed of at the landfill from 1969 to 1980 (Lesage et al. 1990). Carter et al. (2008) analyzed the results from U.S. groundwater quality surveys and found that 1,1-DCE was not a major contaminant, being detected in only 0.66% of sites (11 of 1686). Ellis and Rivett (2007) conducted analyses for volatile organic compounds (VOCs) in groundwater potentially entering the River Tame running through Birmingham, U.K. 1,1-DCE was found above detection limits in 20% of samples, with a maximum concentration of 20 μg/L. They estimated the mean daily flux from groundwater through the riverbed to be 0.1 mg/m2/day or approximately 3 kg/year over 7 km of riverbed.
Given conditions favourable to oxidative growth (nutrients and an oxygen source), many groundwater bacteria are capable of degrading 1,1-DCE to vinyl chloride and ethene. Groundwater can be considered to be a pathway from contaminant sources to sediment and surface waters, if it can be shown that contaminated groundwater is recharging surface waters. This has not been the case with 1,1-DCE. An exposure scenario was not developed for groundwater.
There are few reports of surface water concentrations of 1,1-DCE in Canada above the detection limit of 0.08 μg/L, although 1,1-DCE has at times been detected in both raw and drinking water. In the early 1980s, 1,1-DCE was detected at 12 of 95 water quality monitoring stations in Lake Ontario; the highest concentration was 3.5 μg/L near Scarborough, possibly close to a sewer discharge for the City of Toronto (Kaiser et al. 1983). Nine of 303 stations sampled in the St. Lawrence River in the mid-1980s had “trace” concentrations of 1,1-DCE just above the detection limit (Comba 1985; Comba et al. 1986), yet the maximum concentration reported was 100 μg/L at an industrial outfall near Prescott, Ontario (Comba et al. 1986). Studies across Canada on raw and potable drinking water supplies have rarely found any 1,1-DCE (Otson et al. 1982a; Otson et al. 1982b; Otson 1987; OME 1988; OME 1989; Toronto Water 2004; Health Canada 1994a), although there have been occasional reports of measurable concentrations. For example, a 1,1-DCE concentration of 20 μg/L was measured in one sample of treated drinking water in 1979 (Otson et al. 1982a), and a study of 29 Alberta municipal drinking water supplies by Health Canada between the years of 1978 and 1985 found a maximum concentration of 1.4 µg/L at one location (Health Canada 1994a).
The 1,1-DCE concentration of 100 µg/L (Comba et al. 1986) was selected as the PEC to be used for the risk quotient calculation for an aquatic scenario, as it was judged to represent a reasonable worst-case scenario for an historical industrial discharge.
Sediments and Soil
No data were found for measured concentrations of 1,1-DCE in sediments. Only one study was found, from Ontario in 1993, in which concentrations of 1,1-DCE in soil were measured. Three regions of rural and urban Ontario parkland were sampled, and the highest 98th percentile concentration of 1,1-DCE from all regions was 0.097 μg/kg (OMEE 1993). The maximum concentration was not reported. This result may not be representative of soil concentrations in areas where potential sources of contamination occur, such as industrial areas. As no data were found on 1,1-DCE in sediments, or on the potential toxicity of 1,1-DCE to soil or sediment organisms, scenarios could not be developed for soil or sediment exposures. However, as noted above, 1,1-DCE is not expected to partition to soil or sediment, and exposure could be expected to be negligible.
Ecological Effects Assessment
The toxicity of 1,1-DCE to aquatic organisms has been investigated in a number of studies (Table 5).
Dill et al. (1980) conducted a flow-through bioassay on the effects of 1,1-DCE to fathead minnows (Pimephales promelas), taking into account the volatilization of the chemical from the water. Little variation was found between the median lethal concentration (LC50) values determined for tests of 48 to 96 hours duration. The 96-hr LC50 was determined to be 108 mg/L, and the 96-hr median effect concentration (EC50) was determined to be 75 mg/L. The fish showed signs of distress (loss of equilibrium when swimming and disorientation) during the first 24 hours of exposure and did not recover.
The aquatic organism most sensitive to 1,1-DCE reported in the literature is the alga Chlamydomonas reinhardtii. Brack and Rottler (1994) reported a 72-hr EC10 for growth inhibition of 3.94 mg/L in closed, measured conditions. This value was chosen as the critical toxicity value (CTV) for use in this assessment to predict risks to aquatic organisms.
|Test organism||Endpoint||Value (mg/L)||Reference|
|EC50 growth, 24 to 96-hr|
|> 560||US EPA 1978|
|EC50 growth, 96-hr|
|410||Geyer et al. 1985|
|72-hr EC10 growth|
72-hr EC50 growth
|Brack and Rottler 1994|
|EC50 photosyn., 96-hr|
|712||US EPA 1978|
|LC50 24-hr, static||98||LeBlanc 1980|
|LC50 48-hr, static||76|
|NOEC 48-hr, static||< 2.4|
|LC50 24 to 48-hr, static||11.6||Dill et al. 1980|
|LC50 96-hr||224||US EPA 1978|
|Daphnia magna||48-hr EC50 immobilization||16||CHRIP c2008|
|LC50 24 to 96-hr static||250||Heitmuller et al. 1981|
|NOEC 96-hr static||80|
|LC50 96-hr static||74||Buccafusco et al. 1981|
|LC50 96-hr static||220||Dawson et al. 1977|
|LC50 96-hr static||250||Dawson et al. 1977|
|LC50 24-hr static|
LC50 24-hr flowthrough
|Dill et al. 1980|
|LC50 48-hr static|
LC50 48-hr flowthrough
|LC50 96-hr static|
LC50 96-hr flowthrough
|LC50 10- to 13-day flowthrough||29|
|LC50 5-day flowthrough|
LC50 6-day flowthrough
LC50 7-day flowthrough
LC50 8-day flowthrough
LC50 9-day flowthrough
|LC50 96-hr||45||CHRIP c2008|
 Brack and Rottler (1994) eliminated evaporative losses from the test system.
 EC50 - The concentration of a substance that is estimated to cause some effect on 50% of the test organisms. LC50 – The concentration of a substance that is estimated to be lethal to 50% of the test organisms. NOEC – The no-observed-effect concentration is the highest concentration in a toxicity test not causing a statistically significant effect in comparison to the controls.
[*] Critical toxicity value (CTV).
Toxicity of 1,1-DCE to terrestrial organisms has been investigated in a number of studies, including in plants (Pestemer and Auspurg 1986), microorganisms (Greim et al. 1975; Bronzetti et al. 1983), invertebrates (Viswanathan 1984), and vertebrates (Prendergast et al. 1967; Murray et al. 1979; Quast et al. 1986; Van Duuren et al. 1979; Jones and Hathway 1978a).
No effects were noted in a 14-day test on the growth of wheat (Triticum aestivum), oats (Avena sativa), garden cress (Lepidium sativum), lettuce (Lactuca sativa), white mustard (Sinapis alba), Pak-choi (Brassica chinensis), colza (Brassica napus), turnips (Brassica rapa), perennial ryegrass (Lolium perenne), radish (Raphanus sativus), garden vetch (Vicia sativa), mung bean (Vigna radiata), red clover (Trifolium pretense), grain sorghum (Sorghum bicolor) at up to 1000 mg/kg in soil (Pestemer and Auspurg 1986). The only test found on soil organisms (earthworms) was found to be inadequate by the World Health Organization (WHO 2003a).
The mode of action of 1,1-DCE is non-polar narcosis (US EPA 1999). It is rapidly absorbed during oral and inhalation exposure, with most of the free 1,1-DCE and its metabolites found in the liver and kidney. The target organs during acute exposure by oral or inhalation routes are the liver, the kidney, and the Clara cells of the lungs. During chronic exposure the critical effect is a minor fatty change that occurs in the liver of many organisms. The metabolites of 1,1-DCE, including an epoxide, are responsible for the toxic effect within the target cells (US EPA 2002b).
Gallegos et al. (2007) evaluated data on inhalation toxicity to small mammals for many organic substances, including 1,1-DCE. They estimated a toxicity reference value for small mammals of 4.93 mg/kg-bw (kilograms of body weight) per day based on 17 data points ranging from 0.39 to 130 mg/kg-bw per day. All effect levels used were from chronic exposures resulting in no adverse effects; in essence they are no-observed-effect levels (NOELs).
A median lethal concentration (LC50) in rats after 4 hours of inhalation was 6350 ppm (25 400 mg/m3) (Kirk-Othmer 2007). In a study by Speerschneider and Dekant (1995), 1,1-DCE was shown to have a toxic effect in the kidneys of male mice via inhalation of 188 mg/m3 (47 ppm) over 4 hours. This effect was species- and sex-specific due to the presence of cytochrome P450 2E1 found only in male mice.
A study by Prendergast et al. (1967) showed that continuous inhalation of a concentration of 1,1-DCE of 189 mg/m3(47.3 ppm) over a period of 90 days caused widespread morphological alterations in the livers and kidneys of rats, guinea pigs, dogs and monkeys. Continuous inhalation of 101 mg/m3 (25.3 ppm) had no effect on the test animals. Exposure for 8 hours/day 5 days/week had no effect at 395 mg/m3.
The value by Prendergast et al. (1967) is a no-observed-adverse-effect level (NOAEL) for continuous inhalation effects to a variety of mammals at 101 mg/m3 over 90 days. This value was selected for use as the CTV.
Characterization of Ecological Risk
The approach taken in the ecological risk characterization was to examine various supporting information and develop conclusions based on a weight-of-evidence approach and applying precaution as required under section 76.1 of CEPA 1999. Particular consideration has been given to risk quotient analyses, and to the environmental realism of the exposure scenarios used to derive predicted no-effect concentrations (PNECs) and occurrence in the environment. Endpoint organisms have been selected based on analysis of exposure pathways. For each endpoint organism, a conservative PEC and PNEC were determined. The PNEC is the lowest CTV for the organism of interest divided by an appropriate application factor. A risk quotient (PEC/PNEC) was calculated for each of the endpoint organisms in order to help determine whether there is a potential for ecological risk in Canada.
Application factors were derived using a multiplicative approach, in which 10-fold factors are used to account for various sources of uncertainty associated with making extrapolations and inferences related to intra- and interspecies variability, laboratory-to-field extrapolation, and acute-to-chronic toxicity values.
Selected PECs for this assessment, based on measured concentrations of 1,1-DCE in air (2.6 × 10-5mg/m3 at an urban site in Montréal and 5.7 × 10-3 mg/m3 near a landfill site in Toronto), on modelled concentrations in air (0.46 mg/m3 for an industrial release scenario), and measured in surface water (0.1 mg/L in Lake Ontario) have been previously discussed and are presented in Table 6.
The CTV of 101 mg/m3 over 90 days to a variety of mammals was chosen to represent the concentration of 1,1-DCE resulting in no effects to small mammals continuously inhaling 1,1-DCE (Prendergast et al. 1967). The air concentration at a Toronto landfill (5.7 × 10-3 mg/m3) was used as the PEC. An application factor of 10 was chosen to represent the dilution of 1,1-DCE from below ground to the air above a Toronto area landfill.
The aquatic species most sensitive to 1,1-DCE was the green alga Chlamydomonas reinhardtii (Brack and Rottler 1994). A CTV of 3.94 mg/L, the lowest concentration resulting in decreased growth (EC10), was therefore chosen to represent the concentration of 1,1-DCE resulting in a non-significant level of effects in aquatic organisms. An application factor of 10 was applied to account for lab-to-field extrapolation, resulting in a PNEC for aquatic organisms of 0.394 mg/L.
Risk quotients for 1,1-DCE, obtained by dividing the PEC by the PNEC, are summarized in Table 6.
|Medium||Organism||PEC||CTV||Application factor||PNEC||Risk quotient|
|Surface water (freshwater)||Green algae||0.10 mg/L||3.94 mg/L||10||0.394 mg/L||0.25|
|Soil at a landfill||Burrowing mammals||5.7 × 10−3 mg/m3||101 mg/m3||10||10.1 mg/m3||6 × 10−4|
|Urban air||Mammals||2.6 × 10-5 mg/m3||101 mg/m3||10||10.1 mg/m3||2.5 × 10-6|
|Air at an industrial site||Mammals||0.46 mg/m3||101 mg/m3||10||10.1 mg/m3||0.05|
The risk quotients calculated for water and air are significantly less than 1 (Table 6), indicating that there is very low likelihood of ecological harm from the concentrations of 1,1-DCE found in the Canadian environment.
1,1-DCE is a known precursor of formaldehyde in air following its degradation via oxidative reactions with hydroxyl radicals. Formaldehyde appears on the List of Toxic Substances in Schedule 1 of CEPA 1999.
All of the studies on the biotic effects of 1,1-DCE that were reviewed indicate that relatively high concentrations of 1,1-DCE are required to induce adverse effects, and these high concentrations are not found in the environment in Canada, either because sufficient volumes are not released, or due to environmental fate processes. Thus, based on consideration of 1,1-DCE’s low persistence and low potential for accumulation in organisms, the lack of evidence for any recent, ongoing or anticipated increases in releases to the environment or in concentrations in the ambient environment, and based on indications that current concentrations are below levels that would be anticipated to cause ecological harm, it is proposed that 1,1-DCE is not causing ecological harm in Canada.
Uncertainty in the Evaluation of Ecological Risk
The major uncertainties in this assessment relate to exposure characterization. Monitoring studies during the 1980s and 1990s reported industrial releases of 1,1-DCE to water. However, it appears that there is presently low commercial usage of 1,1-DCE. The only industrial use of 1,1-DCE that produced significant releases of 1,1-DCE (as reported in the section 71 survey or to the NPRI) was in 2000 by one company which reported the use of a large volume of a commercial sealer-solvent, which may have resulted in releases to air. Other uses of 1,1-DCE reportedly produced very little or no emissions, and consequently there may currently be very little exposure of 1,1-DCE to the environment from industrial releases. These uncertainties were dealt with by developing a series of exposure scenarios using ambient monitoring data for a variety of locations and by using a modelled approach for a hypothetical industrial release.
Potential to Cause Harm to Human Health
Environmental Media and Diet
Empirical data were identified for 1,1-DCE environmental concentrations in ambient air, indoor air, raw and treated drinking water, soil and food and beverages in Canada. Empirical data were also identified for environmental media in other locations. All studies identified containing empirical data for each environmental medium are summarized in Appendix 1, Tables A1 to A5.
In a recent study of ambient air levels conducted as part of the ongoing NAPS network, 1,1-DCE was not detected in 1896 samples, at a detection limit of 0.026 μg/m3, among 43 Canada-wide sites during the collection period of January to December 2008 (NAPS 2008). The substance was detected near the method detection limit (0.012 µg/m3) in a limited number of samples of outdoor air of residential backyards in Regina in 2007; the median concentration was 0.006 µg/m3 (Health Canada 2010a). Another study was unable to detect 1,1-DCE in outdoor air in Windsor, Ontario during either the summer or winter season above detection limits of 0.152 µg/m3 and 0.046 µg/m3 for 2005 and 2006, respectively (Health Canada 2010b). In a study of indoor and outdoor air of residences in Ottawa, Ontario, with sampling conducted during the winter of 2002-03, the median outdoor air concentration was 0.005 µg/m3 (detection frequency: 18%) (Zhu et al. 2005). The highest median concentration of 1,1,-DCE in outdoor air in Windsor, Ontario for 2005 of 0.076 µg/m3 was used in deriving the intake estimate (Appendix 2) as it represents the highest median concentration across recent Canadian outdoor air studies.
In an indoor air study of residences in Windsor, Ontario during 2005 and 2006, 1,1-DCE was detected near the method detection limits (0.152 μg/m3 and 0.046 µg/m3 for 2005 and 2006, respectively) in a limited number of samples (i.e., detection frequencies between 0.4 to 4.3%). For the summer of 2005, the median and 95th percentile concentrations were both 0.076 µg/m3. Also, for the winter of 2005, the median and 95th percentile concentrations were both 0.076 µg/m3. For the summer and winter of 2006, the median and 95th percentile concentrations were both 0.023 µg/m3. (Health Canada 2010b). In a 2007 study in Regina, Saskatchewan, 1,1-DCE was detected near the method detection limit (0.012 µg/m3) in a limited number of samples of indoor air (i.e., detection frequency of 5.9 – 10.1%). For the summer of 2007, the median and 95th percentile concentrations of the 5-day samples were 0.006 µg/m3and 0.023 µg/m3, respectively. For the winter of 2007, the median and 95th percentile concentrations of the 5-day samples were 0.006 µg/m3 and 0.027 µg/m3, respectively (Health Canada 2010a). In a study of indoor and outdoor air of residences in Ottawa, Ontario, with sampling conducted during the winter of 2002-03, the median and 95th percentile indoor air concentrations were 0.005 µg/m3 and 0.99 µg/m3, respectively (detection frequency: 45%) (Zhu et al. 2005). The highest median concentration of 1,1-DCE in indoor air from recent studies in Canada was 0.076 µg/m3, based on Windsor, Ontario data collected during 2005; this value was used in deriving the environmental intake estimate (see Appendix 2). Also, the highest 95th percentile concentration in indoor air from recent studies in Canada was 0.99 µg/m3, based on Ottawa, Ontario data collected during the winter of 2002-03.
Among fourteen recent Canadian surveys of drinking water among different cities of the country between 2003 and 2008, 1,1-DCE was not detected (CBWO 2008; City of Victoria 2008; City of Vancouver 2008; TDWS 2008; City of Niagara Falls 2008; CSWTP 2008; City of London 2008; OME 2008; Utilities Kingston 2008; BCOS 2008; EPCOR 2008; Ville de Montréal 2006; CCW 2003; COWQS 2003). A summary of drinking water data obtained from sites distributed across the United States conducted by the United States Geological Survey over a sampling period of 1985 to 2001 revealed median 1,1-DCE levels in samples where it was detected at 0.20 μg/L and 0.026 μg/L for public wells and domestic wells respectively (Zogorski et al. 2006). Zogorski et al. (2006) also determined the detection frequencies of DCE, by percent of total samples at an assessment level of 0.2 μg/L, in public wells and domestic wells of 1.3% (n = 1 096) and 0.21% (n = 2 400) respectively. The highest detection limit of the recent Canadian surveys, 0.52 μg/L for 35 samples in Ottawa, Ontario, in 2003 was used in deriving the intake estimate (see Appendix 2).
Studies analyzing for 1,1-DCE in food items available in Canada were conducted by Enviro-Test Laboratories in the early 1990s in Ville-Mercier, Quebec (ETL 1993), Windsor, Ontario (ETL 1992) and Cayley, Alberta (ETL 1991). 1,1-DCE was not detected in any of four samples of 34 food composites in these studies. The foods that comprised these composites are presented for Ville-Mercier, Quebec in Appendix 1, Table A4. The detection limits in the ETL (1991, 1992) studies were 50 μg/kg and 1.0 μg/L in solids and liquids, respectively, and in the ETL (1993) study the detection limits were 5.0 μg/kg and 1.0 μg/L for solids and liquids, respectively. Residuals of 1,1-DCE in food items are not currently monitored by the Canadian Food Inspection Agency (2009 email from the Canadian Food Inspection Agency to Risk Assessment Bureau, Health Canada; unreferenced).
In terms of food packaging, 1,1-DCE may exist as a manufacturing impurity in wrap that contains PVDC. Packaging of snack products may involve a PVDC coating on cellulose or polypropylene (Gilbert et al. 1980). A copolymer of PVDC and polyvinyl chloride (PVC) may be used to package patés, cooked sausages and processed cheeses as part of “chub” packs (Gilbert et al. 1980). One brand name of household plastic wrapping film used in a food context was a copolymer composed of PVC (15–20% by weight) and PVDC (80–85% by weight) (Birkel et al. 1977). While the formulation of this brand name of food plastic wrap in North America was changed in 2004 from PVDC to low-density polyethylene, this does not preclude other food plastic wraps containing PVDC being marketed under different brand names in Canada (Allen and Albala 2007; Dow 2006b). For example, PVDC, listed as an acceptable polymer for food packaging applications by Health Canada, was present in two new packaging applications registered in Canada as of 2007 (Health Canada 2009b). One of these packaging applications is considered suitable for meat, cheese, sausage packaging, and flame-retardant fibres and filaments (Solvay 2010).
In a 2005 Japanese study, 1,1-DCE was not detected by headspace gas chromatography analysis at a detection limit of 0.06 µg/g in PVDC home wrapping film and casing film for sausage, cheese and uiro [Japanese steamed cake], or in PVC food containers, water pipes, home wrapping film and toys (Ohno et al. 2005). In a 1976 Japanese study, no 1,1-DCE monomer was detected at a detection limit of 1 ppm in PVDC house hold wrap and casing film for fish sausage (Motegi et al. 1976). A 1977 American study of food plastic wrap determined mean 1,1-DCE monomer concentrations in household film and industrial-purpose films at 8.8 ppm (6.5–10.4 ppm) and 18.4 ppm (10.8–26.2 ppm) respectively (Birkel et al. 1977). A 1978 American study revealed a mean concentration of 1,1-DCE in food plastic wrap at 5.9 ppm (2.4–12.7 ppm) (Hollifield and McNeal 1978). This study also examined the mean concentration of 1,1-DCE determined in three food-simulating solvents resulting from contact with two different thicknesses of film (0.5 mm and 6.0 mm). This migration study was allowed to proceed for varying lengths of time (0.5 to 39 days) at 49°C until each sample had either experienced full migration of 1,1-DCE into the solvent or had reached an apparent equilibrium level (Hollifield and McNeal 1978). For the 0.5 mm-thick film, mean concentrations of 1,1-DCE in heptane, corn oil and water were 39 ppb (34–44 ppb), 34 ppb (18–41 ppb) and 25 ppb (24–27 ppb), respectively (Hollifield and McNeal 1978). For the 6 mm-thick film, mean concentrations of 1,1-DCE in heptane, corn oil and water were 320 ppb (66–579 ppb), 255 ppb (12–627 ppb) and 177 ppb (90–211 ppb), respectively (Hollifield and McNeal 1978). For exposure assessment purposes, the results of the 0.5 mm-thick film experiment may be more relevant for food exposure, as one brand name of household food plastic wrap was marketed at 0.5 mm thickness (Birkel et al. 1977).
A 1980 survey of foodstuffs packaged in films containing PVDC purchased in Great Britain revealed a mean 1,1-DCE monomer concentration of 0.019 ppm (0.010–0.025 ppm) in potato crisps after 30 days of storage at ambient temperature (Gilbert et al. 1980). In addition, of the other foods tested for migration of 1,1-DCE (biscuits, cakes, snack products, cooked meats and cheeses), only two food items had concentrations exceeding the detection limit of 0.005 ppm: black pudding and liver pate both had concentrations ranging from 0.005 to 0.01 ppm, with levels detected at the outer edges of the product (Gilbert et al. 1980). These food items had been stored for 60 days at ambient temperature (Gilbert et al. 1980). This study also determined mean concentrations of 1,1-DCE monomer in the actual PVDC-containing films used to package the foods: 0.49 ppm (< 0.06–1.26 ppm) for PVDC/polypropylene, < 0.04 ppm for PVDC/cellulose, 0.11 ppm (< 0.02–0.28 ppm) for PVDC-PVC and 0.15 ppm (0.12–0.16 ppm) for the potato crisp bags. In a related study, at a lower detection limit of 0.001 mg/kg, 1,1-DCE was detected in food in contact with PVDC film wrap (MAFF 1980). A Japanese study conducted in August 2004 of 13 samples of various foodstuffs (sausage, fish sausage, boiled fish paste and cheese) revealed a concentration range of 0.003–0.0095 µg/g (Ohno and Kawamura 2006). The food concentration data are presented in Appendix 1, Table A4.
The United States Food and Drug Administration lists 1,1-DCE as an indirect food additive, as some preservative polymer mixtures coated onto fruits and vegetables (especially citrus fruits) may contain 1,1-DCE (US FDA 2006). Food migration studies for these preservative mixtures have not been identified, though this does not preclude the possibility of citrus fruits imported from the United States being potential sources of 1,1-DCE exposure.
In terms of deriving an intake estimate, for each of the twelve food categories that comprise Canadian consumption as specified in the Health Canada (1998) report, detection limits were used for food items where no 1,1-DCE was detected, while mean concentrations were used for food items where 1,1-DCE was detected. These detection limits and mean concentrations were selected from all the available studies listed in Appendix 1, Table A4, regardless of location. This conservative approach of using detection limits for some food items may lead to an overestimate of actual exposure. The food intake estimates and studies selected are presented in Appendix 6. Scenarios involving 1,1-DCE monomer migration from food packaging to food items were not conducted, as empirical data exist for 1,1-DCE concentrations in PVDC-packaged foodstuffs (as provided in Appendix 4). However, a submission to the United States Environmental Protection Agency, as part of the Voluntary Children’s Chemical Evaluation Program (VCCEP), provided a migration scenario for food wrap with typical and high-end oral intakes of 0.01 and 0.0375 μg/kg-bw per day respectively (US EPA 2002a; Williams et al. 2006). These intakes were substantially less than the estimated intakes from food and beverages presented in Appendix 2.
In terms of biomonitoring data, the National Health and Nutrition Examination Survey 2003–2004 conducted by the National Center for Health Statistics in the United States did not detect 1,1-DCE in any of 1367 samples of human blood from adults aged 20 to 59 years, detection limit of 0.009 ng/ml (NCHS 2009). 1,1-DCE was qualitatively detected in one of twelve breast milk samples derived from four cities in the United States, detection limit unspecified (Pellizzari et al. 1982). In a related study, 1,1-DCE was qualitatively detected in one of eight samples of breast milk derived from four cities in the United States, detection limit unspecified (Erickson et al. 1980).
In the most recent study identified for Canadian soil levels of 1,1-DCE, the maximum concentration detected in urban parkland and rural parkland soil of Ontario in the early 1990s was 0.12 ng/g solids and 0.098 ng/g solids, respectively (OMEE 1993). The weighted average of Ontario urban parkland, rural parkland (not including northwest region) and rural parkland (northwest region) soil of 0.046 μg/kg solids was used in deriving the intake estimate (see Appendix 2).
The maximum estimated intake for the general population was 1.34 μg/kg-bw per day for not formula-fed infants aged 0 to 6 months (Appendix 2). Indoor air and food and beverages are the main sources of estimated environmental exposure; however, the reliance on detection limits for several of the food items may indicate that the food and beverage intake estimate may exceed actual exposure. Food wrap containing PVDC is the expected major source of 1,1-DCE monomer in foodstuffs.
Residual amounts of 1,1-DCE can be present in carpet latex backing, industrial insulation adhesives, photographic film coatings, flame-retardant clothing, and PVDC/fluorinated copolymer oil- and water-repellent coatings of textiles (US EPA 2002a), but were not detected in a recent Japanese study of PVC water pipes, home wrapping film and toys (e.g. ball, soft toy, food toy and face mask), detection limit of 0.06 μg/g (Ohno et al. 2005). 1,1-DCE was detected in food plastic wrap at mean 1,1-DCE monomer concentrations in household film and industrial-purpose films at 8.8 ppm (6.5–10.4 ppm) and 18.4 ppm (10.8–26.2 ppm) respectively (Birkel et al. 1977).
As the 1,1-DCE is found bound within the polymer matrix, exposure resulting from use of these products is expected to be minimal (ATSDR 1994). However, as 1,1-DCE has a very high vapour pressure (600 mm Hg at 25°C), some of the residual monomer in consumer products may be released to indoor air. In an indoor air study of 75 residences in Ottawa, Ontario, conducted during the winter of 2002-2003, there was no correlation between the floor area carpeted (and backed with 1,1-DCE-containing latex) and measured values of 1,1-DCE in indoor air (Zhu et al. 2005). The arithmetic mean indoor air concentration was 0.27 μg/m3, with a range of not detected to 4.05 μg/m3, detection limit of 0.011 μg/m3(Zhu et al. 2005). The 75th percentile of the emission rate of 1,1-DCE in these 75 Ottawa homes was measured at 0.05 mg 1,1-DCE per hour (Zhu et al. 2005).
In a review of 50 indoor air studies of international locations conducted between 1978 and 1990, emission sources of 1,1-DCE that are unique to an indoor context (such as construction materials, including carpet latex backing) have resulted in concentrations of 1,1-DCE in indoor air that are elevated by an overall factor of approximately 13 compared to ambient air levels (Brown et al. 1994).
As the likely route of exposure to 1,1-DCE from consumer products is inhalation of indoor air containing 1,1-DCE emissions from construction materials, and as recent empirical data exist for indoor air of Canadian residences, use of the highest median concentration of 1,1-DCE across recent indoor air studies (i.e., 0.076 µg/m3 in Windsor, Ontario during 2005) (Health Canada 2010b) in deriving the daily intake estimate for the indoor air component of environmental exposure is considered to account for any potential inhalation exposure to end products. In addition, use of the highest 95th percentile concentration from recent Canadian indoor air studies (i.e., 0.99 µg/m3 in Ottawa, Ontario during the winter of 2002-03) (Zhu et al. 2005) would also account for any such inhalation exposure to end products when comparing to less than chronic non-cancer endpoints (see Characterization of Risk to Human Health section below).
Though end product exposure is considered accounted for in the overall intake estimate, a 1,1-DCE emission scenario from carpet latex backing to indoor air was presented in a report submitted by the Dow Chemical Company as part of the VCCEP program (US EPA 2002a). The submission estimated typical and high-end intakes of 0.023 and 0.027 μg/kg-bw per day for children, assuming that 80% of a 24-hour period is spent indoors (Williams et al. 2006). Newly constructed or renovated homes would presumably have the highest indoor air concentrations of 1,1-DCE from carpet latex backing, as the emission rate in the VCCEP scenario assumed first-order decay (Williams et al. 2006). These scenario-based inhalation intake estimates are roughly equivalent to the empirically based intake estimates presented in Appendix 2 (maximum inhalation intake estimate from indoor air of 0.04 μg/kg-bw per day) and indicate that use of recent empirical indoor air data identified in Windsor, Ontario in deriving the overall environmental exposure is protective of any potential exposures to end products through inhalation.
The VCCEP submission also concluded that dermal exposures to 1,1-DCE in textiles were insignificant or irrelevant due to the high-temperature processing of the fabrics that would drive off the minimal levels of 1,1-DCE present in coatings of textiles (< 5 ppm concentration in coating prior to processing) (US EPA 2002a). Therefore, dermal exposure to 1,1-DCE residuals in PVDC/fluorinated copolymer coatings of textiles in addition to flame-retardant and ignition-resistant clothing was not characterized on this basis. In addition, the VCCEP submission indicated that fire- and ignition-resistant clothing was typically used in industrial settings (US EPA 2002a). In addition, dermal exposures to 1,1-DCE in photographic paper and film are expected to be negligible, as the 1,1-DCE is contained in an internal latex layer and therefore essentially poses no potential for migration (US EPA 2002a). Dermal exposure to 1,1-DCE from its presence in textiles, flame retardants and ignition-resistant clothing, and photographic paper and film is considered to be negligible for the general population (i.e., primarily occupational). Therefore, the dermal route was not considered to be a significant route of consumer exposure relative to the inhalation route.
Confidence in Exposure Assessment
Confidence in the environmental exposure dataset is considered to be high. Empirical data specific to Canada were available for all environmental media and were recent for ambient air, indoor air and drinking water. The use of detection limits in deriving the overall intake estimate for drinking water and some food categories indicates that average concentrations in these media may have been substantially lower. As the main source of environmental releases of 1,1-DCE in the 2000 calendar year (point releases from facilities) is no longer ongoing, confidence that emissions to ambient air have decreased is high.
As for consumer product exposure, confidence is high that use of the highest median concentration for indoor air from Windsor, Ontario (2005) (0.076 µg/m3) in deriving the indoor air component of the daily intake estimates for environmental exposures, and use of the highest 95th percentile indoor air concentration (0.99 µg/m3, Ottawa, Ontario, 2002-03) when comparing to less than chronic non-cancer endpoints (see Characterization of Risk to Human Health section below), would account for any potential inhalation exposures through releases of 1,1-DCE from PVDC-containing construction materials. In addition, as the environmental intake estimates for indoor air are similar to the intake estimates derived from the carpet latex backing release scenario in the VCCEP submission, confidence is high that any potential inhalation exposure through end products is addressed in the intake estimates.
Health Effects Assessment
An overview of key toxicological studies is presented in Appendix 3.
Carcinogenicity bioassays reviewed in several identified assessments (IARC 1986, 1999; IPCS 1990, 2003; US EPA 2002b) include those conducted by oral, inhalation and subcutaneous routes of exposure as well as a dermal tumour initiation study. Many of these studies are limited by study design or conduct, including exposure durations of 1 year or less or administration of less than the maximum tolerable dose.
An exposure-related, increased incidence of tumours (renal adenocarcinomas) was observed in male (but not female) Swiss mice exposed by inhalation for 1 year to 0, 10 or 25 ppm 1,1-DCE (equivalent to 0, 40 and 100 mg/m3, respectively. This increase was only significant at the highest concentration (Maltoni et al. 1984, 1985; IPCS 1990) and it was the endpoint used by the U.S. EPA to develop an inhalation cancer potency factor (5 × 10-5 per μg/m3) (Roberts et al. 2002).
Similarly, a cancer potency estimate was derived by Health Canada. A lowest tumorigenic concentration 05 (TC05) of 4.2 mg/m3 (equivalent to 5.6 mg/kg-bw per day, based on the incidence of pulmonary adenomas in both male and female mice (in the same inhalation carcinogenicity study used by the EPA) was calculated. The TC05 is defined as the concentration, generally in air, associated with a 5% increase in incidence or mortality due to tumours (Health Canada 1996).
The renal tumours seen were suggested to be related to the toxicity of a metabolite following metabolism via CYP2E1 in the kidney of the mouse. Some researchers have reported an absence of CYP2E1 in human kidney (Amet et al. 1997; Cummings et al. 2000), suggesting that these tumours may not be relevant to humans. The incidence of other tumours, namely mammary carcinomas in female Swiss mice and pulmonary adenomas in male and female Swiss mice, was significantly increased, but without a clear exposure–response relationship. 1,1-DCE was also active as an initiator of lung papillomas in female Swiss mice (Van Duuren et al. 1979). There was also no evidence of carcinogenicity in rat or hamster studies.
1,1-DCE appears to be genotoxic in micro-organisms in the presence of an exogenous metabolic activating system, while mixed results were obtained in the absence of such an activating system. Mixed results have also been produced in mammalian cells in vitro. It is generally non-genotoxic in in vivoassays (chromosomal aberration, rat; dominant lethal, mouse and rat; micronucleus, mouse), although chromosomal aberrations in the bone marrow of Chinese hamsters and minimal DNA binding in the liver and kidneys of mice and rats have been reported (IPCS 1990; US EPA 2002b).
No increase in tumour incidence was associated with 1,1-DCE-exposed workers in any of the three epidemiology studies in the literature (Ott et al. 1976; Thiess et al. 1979; Waxweiler et al. 1981), but their small cohort size, short observation period and potentially confounding variables preclude any evaluation of 1,1-DCE’s carcinogenic potential in humans.
1,1-DCE has been classified by IARC (1999) as not classifiable as to its carcinogenicity to humans based on inadequate evidence in humans and limited evidence in experimental animals (Group 3); whereas the U.S. EPA (US EPA 2002b) concluded that there is suggestive evidence for the carcinogenicity of 1,1-DCE. In addition, it is classified in the European Union as Carcinogen Category 2, with the hazard statement “Suspected of causing cancer” (European Union 2008).
The target organs for non-cancer effects are the liver, the kidneys and the Clara cells of lungs. The lowest lowest-observed-adverse-effect concentration (LOAEC) identified was 10 ppm (40 mg/m3), based upon significant increases in kidney damage (regressive changes and/or abscesses and nephritis) in male Swiss mice exposed to 1,1-DCE for 52-weeks (Maltoni et al. 1984, 1985). The lowest oral lowest-observed-effect level (LOEL) was 5 mg/kg-bw per day, based upon chronic renal inflammation in male and female F344 rats in a 2-year gavage study (NTP 1982).
For non-cancer effects based on a less than chronic period, a lowest inhalation LOAEC of 60 mg/m3 was determined in a mouse developmental toxicity study. An increase in the mean number of fetuses with skeletal effects was observed in this study (Short et al. 1977).
Characterization of Risk to Human Health
In the critical carcinogenicity bioassay (Maltoni et al. 1984, 1985), renal adenocarcinomas were observed at the highest concentration (100 mg/m3) only. This concentration is approximately 1,300 000 times higher than the highest median concentration of 1,1-DCE in air in Canada (0.076 µg/m3in indoor air; Health Canada 2010b).
1,1-DCE is generally non-genotoxic in in vivoassays.
Using the cancer potency factor derived by Health Canada (i.e., lowest TC05 of 4.2 mg/m3; equivalent to 5.6 mg/kg-bw per day) for the same inhalation carcinogenicity study in mice mentioned above, and using the upper-bounding estimate of inhalation intake for adults in the Canadian population (i.e., ≤ 0.017μg/kg-bw per day for adults aged 20–59 and 60+ years; approximates lifetime exposures), the margin between these estimates is in the order of magnitude of 3.3 x 105. This margin is considered adequate to address uncertainties in the health effects and exposure databases.
A comparison of the lowest critical inhalation effect level for chronic non-cancer effects (40 mg/m3; Maltoni et al. 1984, 1985) with the highest median concentration of 1,1-DCE in air in Canada (0. 076 µg/m3 in indoor air; Health Canada 2010b) results in a margin of exposure of approximately 530 000. This margin is considered adequate to address uncertainties in the health effects and exposure databases.
A comparison of the lowest critical inhalation effect level for non-cancer effects based on a less than chronic period (60 mg/m3; Short et al. 1977), with the highest 95th percentile concentration of 1,1-DCE in air in Canada (0.99 µg/m3) results in a margin of exposure of approximately 61 000. This margin is considered adequate to address uncertainties in the health effects and exposure databases.
The margin of exposure between the highest upper-bounding estimate of intake from all sources of exposure (1.34 µg/kg-bw per day for not formula-fed infants between 0 and 6 months of age; Appendix 2and the critical oral effect level for non-cancer effects (5 mg/kg-bw per day; NTP 1982) is approximately 3700. This margin is considered adequate to address uncertainties in the health effects and exposure databases.
The non-cancer critical effects of 1,1-DCE associated with cytotoxicity of the liver, kidney and Clara cells of the lung in rats and mice are considered to result from the damage caused through the covalent binding of CYP2E1 activated metabolic products of 1,1-DCE to cellular macromolecules (US EPA 2002b).
Uncertainties in Evaluation of Risk to Human Health
Several of the carcinogenicity bioassays reviewed in identified assessments had study design limitations, including exposure durations of one year or less or administration of less than the maximum tolerable dose. Uncertainties exist regarding inter- and intra-species variation, extrapolation of data from animals to humans and a lack of data in humans for several endpoints.
There is uncertainty in current levels of 1,1-DCE in foodstuffs consumed by Canadians due to minimal recent data identified and the unknown degree to which food package formulations containing PVDC may have changed in Canada or some imported products in recent times. Without conducting a survey on a typical market basket of foodstuffs consumed by Canadians, it is uncertain as to the degree to which the intake estimate presented in Appendix 2 is representative of current Canadian consumption patterns. However, the use of detection limits and mean concentrations for the food categories would likely prevent any underestimation of exposure to 1,1-DCE in food items. There is uncertainty in the current quantities of 1,1-DCE in commerce in Canada, as the reporting year for the survey issued under section 71 of CEPA 1999 was for the 2000 calendar year. There is also uncertainty in the market penetration of PVDC-containing end products for sale to consumers in Canada among total similar products available.
On the basis of the information presented, it is concluded that 1,1-dichloroethene is not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity or that constitute or may constitute a danger to the environment on which life depends. Additionally, 1,1-DCE does not meet the criteria for persistence or bioaccumulation potential as set out in the Persistence and Bioaccumulation Regulations (Canada 2000).
On the basis of the adequacy of the margins of exposure between estimated exposures to 1,1-dichloroethene and critical effect levels, it is concluded that 1,1-dichloroethene is not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health.
It is therefore concluded that 1,1-dichloroethene does not meet the criteria in section 64 of CEPA 1999.
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- Appendix 1: Concentrations of 1,1-DCE in different media
- Appendix 2: Upper-bounding Deterministic Estimate of 1,1-DCE Daily Intake
- Appendix 3: Summary of health effects information for 1,1-dichloroethene
- Appendix 4: Robust Study Summaries for Ecotoxicity Studies
 A determination of whether one or more of the criteria of section 64 are met is based upon an assessment of potential risks to the environment and/or to human health associated with exposures in the general environment. For humans, this includes, but is not limited to, exposures from ambient and indoor air, drinking water, foodstuffs, and the use of consumer products. A conclusion under CEPA 1999 on the substances in the Chemicals Management Plan (CMP) is not relevant to, nor does it preclude, an assessment against the hazard criteria specified in the Controlled Products Regulations, which is part of the regulatory framework for the Workplace Hazardous Materials Information System [WHMIS] for products intended for workplace use. Similarly, a conclusion based on the criteria contained in section 64 of CEPA 1999 does not preclude actions being taken under other sections of CEPA 1999 or other Acts.
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