Draft Screening Assessment Report

Ethene, 1,1-dichloro-
(1,1-DCE)

Chemical Abstracts Service Registry Number
75-35-4

Environment Canada
Health Canada

December 2011

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(PDF Version - 368 KB)

Table of Contents

• Synopsis
• Introduction
• Substance Identity
• Physical and Chemical Properties
• Sources
• Uses
• Releases to the Environment
• Environmental Fate
• Persistence and Bioaccumulation Potential
• Potential to Cause Ecological Harm
• Potential to Cause Harm to Human Health
• Conclusion
• References
• Appendix 1: Concentrations of 1,1-DCE in Different Media
• Appendix 2: Upper-bounding Deterministic Estimate of 1,1-DCE Daily Intake (µg/kg-bw per Day by Various Age Groups)
• Appendix 3: Summary of health Effects Information for 1,1-dichloroethene
• Appendix 4: Robust Study Summaries

Synopsis

Pursuant to paragraphs 68(b) and (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, or 1,1-DCE), Chemical Abstracts Service Registry Number 75-35-4, which was a substance on the Domestic Substances List (DSL) selected for a pilot project for screening assessments. 1,1-DCE 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-DCE 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-DCE were manufactured and imported into Canada in 2000. However, it is no longer produced or imported into Canada for these uses. Small amounts of 1,1-DCE are created unintentionally in several industrial processes; most of this material is reformed into other substances within the facilities.

Globally, 1,1-DCE 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-DCE may persist as an unintended manufacturing residue in some of these items that may be present in Canadian commerce. 1,1-DCE 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-DCE 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-DCE to the NPRI.

1,1-DCE can also 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. Risk management activities have removed many of these solvent uses and poor disposal practices from Canadian society, such that significant new sources of 1,1-DCE in groundwater and soil should be less of an issue.

Based on experimental and modelled data for 1,1-DCE, 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-DCE 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-DCE is not highly hazardous to non-human organisms.

Recent monitoring data show that 1,1-DCE is present in urban air at very low concentrations, often just above analytical detection limits.

The ecological risk characterization considers Canadian monitoring data and the most sensitive non-human species to generate risk quotients. The risk quotients calculated for water and air, which are significantly less than 1, indicate that there is very low likelihood of ecological harm from the concentrations of 1,1-DCE found in the Canadian environment.

Based on the information available, it is proposed that 1,1-DCE 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.

The general population exposure to 1,1-DCE is mainly from indoor air and food and beverages. A comparison of the lowest critical inhalation effect level for non-cancer effects with the weighted average air concentration of 1,1-DCE in indoor and outdoor air 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 non-cancer effects.

A critical effect for characterization of risk to 1,1-DCE is carcinogenicity. Following lifetime inhalation of 1,1-DCE 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 proposed that 1,1-DCE 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 proposed that 1,1-DCE does not meet the criteria set out in section 64 of the Canadian Environmental Protection Act, 1999. Additionally, 1,1-DCE does not meet the criteria for persistence or bioaccumulation potential as set out in the Persistence and Bioaccumulation Regulations.

This substance will be included in the Domestic Substances List inventory update initiative. In addition, and where relevant, research and monitoring will support verification of assumptions used during the screening assessment.

Introduction

This screening assessment was conducted pursuant to section 74 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, as per 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.^[1]

A screening assessment was undertaken on Ethene, 1,1-dichloro- (1,1-DCE) (Chemical Abstracts Service Registry Number 75-35-4) on the basis that this compound was included in the Domestic Substances List (DSL) pilot project for screening assessments. 1,1-DCE was identified 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. 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 draft screening assessment was prepared by staff in the Existing Substances Programs at Health Canada and Environment Canada. This ecological assessment has undergone external written peer review/consultation. As mentioned above, the State of the Science Report for a Screening Health Assessment was previously externally reviewed. Although external comments were taken into consideration, the final content and outcome of the draft screening assessment remain the responsibility of Health Canada and Environment Canada.

The critical information and considerations upon which the draft assessment is based are summarized below.

Substance Identity

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.

Table 1. Substance identity for 1,1-DCE

CAS RN	75-35-4
DSL name	Ethene, 1,1-dichloro-
NCI names[1]	Ethene, 1,1-dichloro- (TSCA, DSL, AICS, SWISS, PICCS, ASIA-PAC, NZIoC) 1,1-Dichloroethylene (DSL, EINECS) 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
Chemical group (DSL Stream)	Discrete organics
Major chemical class or use	Alkenes
Major chemical sub-class	Halogenated alkene
Chemical formula	C2H2Cl2
Chemical Structure
SMILES2	C(=C)(Cl)Cl
Molecular mass	96.94 g/mol

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 K_ow and K_oc values indicate that it will not bind tightly to organic matter in the environment.

Table 2. Physical and chemical properties of 1,1-DCE

Property	Type	Value	Temperature	Reference
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) (dimensionless)	Experimental	2.13	–	PhysProp 2009
Organic carbon-water partition coefficient (log KOC) (dimensionless)	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

Sources

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). TheNotice 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 this year, 1,1-DCE was used in a sealer solvent and for the manufacture of 1,2-dichloroethane and 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 an environmental quality manager of the importing company to Existing Substances Branch, Environment Canada; unreferenced). In addition, use of 1,1-DCE in the manufacture of 1,2-dichloroethane and hydrochloric acid 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 31 000 kg (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 of “> 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 tested 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). As 60% of all annual production of sewage sludge in the United States is applied to land as a soil amendment, sewage sludge may represent a possible source of environmental 1,1-DCE in Canadian soil (Harrison et al. 2006).

Uses

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 to manufacture 1,2-dichloroethane and 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 an environmental quality manager of 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 theFood 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, risk management activities have removed many of these solvent uses and poor disposal practices from Canadian society, such that significant new sources of 1,1-DCE in groundwater and soil should be less of an issue.

Environmental Fate

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.

Table 3. Results of the Level III fugacity modelling (EQC 2003)

Substance released to:	Percentage of substance partitioning into each compartment (%)
	Air	Water	Soil	Sediment
Air (100%)	99.9	0.1	0.03	0
Water (100%)	4.4	95.3	0	0.2
Soil (100%)	55.3	0.7	43.9	0

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

Environmental Persistence

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 [1989]) 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 CEPA 1999’s Schedule 1 List of Toxic Substances.

Hydrolysis is not a significant degradation pathway (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).

Table 4a. Empirical data for degradation of 1,1-DCE

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

Table 4b. Modelled data for degradation of 1,1-DCE

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; HENRYWIN 2008
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

Air

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 2008a). In a similar study conducted in Windsor, Ontario, the Windsor Ontario Exposure Assessment Study of 2005 and 2006, a maximum concentration of 0.020 µg/m³ of 1,1-DCE was detected in the summer of 2005 (Health Canada 2008b).

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.

Groundwater

Concentrations of 1,1-DCE have also 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.

Surface Water

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.

Table 5. Empirical data for toxicity of 1,1-DCE to aquatic organisms

Test organism	Endpoint[2]	Value (mg/L)	Reference
Algae
Pseudokirchneriella subcapitata Green alga	EC50 growth, 24 to 96-hr Freshwater	> 560	US EPA 1978
	NOEC, 96-hr Freshwater	<56
Scenedesmus abundans Green alga	EC50 growth, 96-hr Freshwater	410	Geyer et al. 1985
Chlamydomonas reinhardtii Green alga[1]	72-hr EC10 growth 72-hr EC50 growth	3.94[*] 9.12	Brack and Rottler 1994
Skeletonema costatum Diatom	EC50 photosyn., 96-hr Saltwater	712	US EPA 1978
Aquatic invertebrates
Daphnia magna Water flea	LC50 24-hr, static	98	LeBlanc 1980
	LC50 48-hr, static	76
	NOEC 48-hr, static	< 2.4
Daphnia magna Water flea	LC50 24 to 48-hr, static	11.6	Dill et al. 1980
Americamysis bahia Oppossom shrimp	LC50 96-hr	224	US EPA 1978
Daphnia magna	48-hr EC50 immobilization	16	CHRIP c2008
Vertebrates (fish)
Cyprinodon variegates Sheepshead minnow	LC50 24 to 96-hr static	250	Heitmuller et al. 1981
	NOEC 96-hr static	80
Lepomis macrochirus Bluegill	LC50 96-hr static	74	Buccafusco et al. 1981
Lepomis macrochirus Bluegill	LC50 96-hr static	220	Dawson et al. 1977
Menidia beryllina Inland silverside	LC50 96-hr static	250	Dawson et al. 1977
Pimephales promelas Fathead minnow	LC50 24-hr static LC50 24-hr flowthrough	175 116	Dill et al. 1980
	LC50 48-hr static LC50 48-hr flowthrough	169 108
	LC50 96-hr static LC50 96-hr flowthrough	169 108
	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	97 74 29 29 29
Oryzias latipes Rice fish	LC50 96-hr	45	CHRIP c2008

 

Terrestrial Compartment

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 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^-3mg/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.

Table 6. Risk quotients calculated for 1,1-DCE

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 CEPA 1999’s Schedule 1 List of Toxic Substances.

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

Exposure Assessment

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). 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, detection limit of 0.012 µg/m³ (Health Canada 2008a). In a similar study conducted in Windsor, Ontario, the Windsor Ontario Exposure Assessment Study of 2005 and 2006, a maximum concentration of 0.020 µg/m³ of 1,1-DCE was detected in the summer of 2005 (Health Canada 2008b). The detection limit of 1,1-DCE in the NAPS survey, 0.026 µg/m³, was used in deriving the intake estimate (Appendix 2) as it contains 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 maximum and mean concentrations of 1,1-DCE in a recent indoor air study of residences in Windsor, Ontario, were 1.380 µg/m³ and 0.025 µg/m³, respectively, for the summer of 2005 (Health Canada 2008b). In addition, measurements of personal breathing-zone air of residents in Windsor, Ontario, in 2005 revealed a maximum concentration of 0.400 µg/m³ (Health Canada 2008b). In a 2007 study in Regina, Saskatchewan, the highest maximum and mean 1,1-DCE concentrations in indoor air were 0.038 µg/m³ and 0.014 µg/m³, respectively, for the winter season (Health Canada 2008a). A weighted average of 1077 samples of indoor air from the Health Canada (2008a and 2008b) studies of 0.0125 µg/m³ was used to represent indoor air levels in generating an intake estimate (see Appendix 2).

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 Environmental Health Directorate 1998 report (EHD 1998), 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 1.32 µg/kg-bw per day for non formula-fed infants aged 0 to 6 months (Appendix 2). 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.

Consumer Products

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 in the fall of 2002, 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 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 over all 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, the weighted average concentration of 1,1-DCE (0.0125 µg/m³) measured among 1077 samples during the summer and winter seasons of 2005 and 2006 in Windsor, Ontario, and 2007 in Regina, Saskatchewan, is considered representative of end product exposure. This value was used in deriving the intake estimate for the indoor air component of environmental exposure (see Appendix 2).

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 of 0.01 µg/kg-bw per day) and indicate that use of recent empirical indoor air data identified in Windsor, Ontario, and Regina, Saskatchewan, in deriving the overall environmental exposure is protective of 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, ambient air 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 a weighted average concentration for indoor air in Windsor, Ontario (2005 and 2006) and Regina, Saskatchewan (2007) is protective of releases 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 inhalation exposure to 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 (Health Canada 1994b). 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 oninadequate 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.

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).

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 over 7 040 000 times higher than the weighted average concentration of 1,1-DCE in air in Canada (0.0142 µg/m³: NAPS 2008; Health Canada 2008a, 2008b). This weighted average concentration of air exposure throughout the day assumes that, on average, 3 hours per day are spent outdoors (0.026 µg/m³; NAPS 2008) and 21 hours per day are spent indoors (0.0125 µg/m³; Health Canada 2008a, 2008b) as per EHD (1998).

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., < 0.001 µg/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 10⁶. 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 (40 mg/m³; Maltoni et al. 1984, 1985) with the weighted average concentration of 1,1-DCE in air in Canada (0.0142 µg/m³) results in a margin of exposure of approximately 2 820 000. 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 (1.32 µg/kg-bw per day for non formula-fed infants between 0 and 6 months of age; Table 1) and the critical oral effect level for non-cancer effects (5 mg/kg-bw per day; NTP 1982) is 3800. 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.

Conclusion

On the basis of the information presented, it is proposed that 1,1-DCE 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. 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-DCE and critical effect levels, it is proposed that 1,1-DCE 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 proposed that 1,1-DCE does not meet the criteria in section 64 of CEPA 1999.

This substance will be considered for inclusion in theDomestic Substances List inventory update initiative. In addition and where relevant, research and monitoring will support verification of assumptions used during the screening assessment.

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Appendix 1: Concentrations of 1,1-DCE in Different Media

Table A1. Concentration of 1,1-DCE in ambient air

Location	Sampling period	Number of samples	Detection limit (µg/m3)	Mean concentration[1](µg/m3)	Reference
Windsor, Ontario	January 23 to March 25, 2006 July 3 to August 26, 2006	219 214	0.003	ND (ND–0.005) ND (ND–0.013)	Health Canada 2008b
Windsor, Ontario	January 24 to March 19, 2005 July 4 to August 27, 2005	200 217	0.003	ND (ND–0.012) 0.005 (ND–0.020)	Health Canada 2008b
Regina, Saskatchewan	January 8 to March 16, 2007 June 20 to August 29, 2007	98 (winter; only 24-h canisters reported) 97 (summer; 5-day canisters)	0.012	ND ND (max 0.014)	Health Canada 2008a
Canada-wide sites (43 locations)	January to December 2008	1896	0.026	ND (0.002–0.013)	NAPS 2008
Ottawa, Ontario (residential areas)	Fall 2002	75	0.011	0.05 [ND–0.83] (detected in 13 of 75 samples)	Zhu et al. 2005
Canada-wide sites	1989–1996	9128	ns	0.06 [ND–0.78]; 8% > detection limit)[4]	NAPS 2008
Montréal, Quebec (urban)	1993	160	0.2 (0.05 ppbv)[2]	0.03 [ND–0.30] (14% > detection limit)	Environment Canada 1995
Montréal, Quebec (suburban)	1993	24	0.2 (0.05 ppbv)[2]	0.00 [ND–0.04] (0% > detection limit)	Environment Canada 1995
Sainte-Françoise, Quebec (rural)	1993	34	0.2 (0.05 ppbv)[2]	0.02 [ND–0.12] (6% > detection limit)	Environment Canada 1995
Montréal, Quebec (urban)	1992	166	0.2 (0.05 ppbv)[2]	0.00 [ND–0.02] (0% > detection limit)	Environment Canada 1995
Montréal, Quebec (urban)	1991	91	0.2 (0.05 ppbv)[2]	0.01 [ND–0.22] (4% > detection limit)	Environment Canada 1995
Montréal, Quebec (urban)	1990	110	0.2 (0.05 ppbv)[2]	0.00 [ND–0.11] (2% > detection limit)	Environment Canada 1995
Montréal, Quebec (urban)	1989	76	0.2 (0.05 ppbv)[2]	0.03 [ND–0.44] (13% > detection limit)	Environment Canada 1995
Greater Vancouver Regional District	1989–1992	473	0.2 (0.05 ppbv)[2]	0.05 (4% > detection limit)[3]	Environment Canada 1994
Canada (sites unspecified)	1989–1990	1100	0.2 (0.05 ppbv)[2]	0.06 (9% > detection limit)[3]	Environment Canada 1994
Windsor, Ontario	July 1987 to October 1990	124	ns	ns [ND–0.3] (10 of 124 samples > detection limit)	Environment Canada 1992
Walpole Island, Ontario	January 1988 to October 1990	61	ns	ns [ND–0.2] (8 of 61 samples > detection limit)	Environment Canada 1992
Toronto, Ontario (downtown)	June–August 1990	16	0.4 (MQL = 2.1)	1.9	OME 1991d
Toronto, Ontario (residential)	June–August 1990	7	0.4 (MQL = 2.1)	0.4	OME 1991d
Canada (residential homes)	February–March 1987	6	6 ng/tube (collection vial)	0.3 [ND–1]	Chan et al. 1990
Canada (residential homes)	November–December 1986	12	6 ng/tube (collection vial)	3.2 [ND–7]	Chan et al. 1990

Table A2. Concentration of 1,1-DCE in indoor air

Location	Sampling period	Number of samples	Detection limit (µg/m3)	Mean concentration[1](µg/m3)	Reference
Windsor, Ontario (personal breathing-zone air)	January 24 to March 19, 2005 July 4 to August 27, 2005	220 209	0.003	0.003 (ND–0.090) 0.009 (ND–0.400)	Health Canada 2008b
Windsor, Ontario	January 23 to March 25, 2006 July 3 to August 26, 2006	227 211	0.003	0.008 (ND–0.463) 0.012 (ND–0.103)	Health Canada 2008b
Windsor, Ontario	January 24 to March 19, 2005 July 4 to August 27, 2005	231 217	0.003	0.005 (ND–0.185) 0.025 (ND–1.380)	Health Canada 2008b
Regina, Saskatchewan[2]	January 8 to March 16, 2007 June 20 to August 29, 2007	90 101	0.012	0.014 (ND–0.083) 0.013 (ND–0.033)	Health Canada 2008a
Ottawa, Ontario (75 homes)	Fall 2002	75	0.011	0.27 [ND–4.05]; (detected in 34 of 75 homes)	Zhu et al. 2005
International locations (literature review of 50 studies)	1978–1990	n = 50 studies	ns	1–< 5	Brown et al. 1994
Toronto, Ontario (office)	June–August 1990	8	0.4 (MQL = 2.1)	5	OME 1991d
Toronto, Ontario (domestic)	June–August 1990	4	0.4 (MQL = 2.1)	5.4	OME 1991d
Canada (residential homes)	November–December 1986	12	6 ng/tube (collection vial)	8.4 [ND–77]	Chan et al. 1990
Canada (residential homes)	February/March 1987	6	6 ng/tube (collection vial)	3.8 [ND–13]	Chan et al. 1990
Woodland, California (residential homes)	June 1990	128	0.78 (MQL)	not quantifiable in any sample	CARB 1992
North Carolina (Research Triangle Park area – residential homes)	Summer	15	ns	ns (detected in 4 of 15 homes- mean of 12.06 and range of 0.46–23.9 µg/m3)	Pleil et al. 1985
North Carolina (Research Triangle Park area – residential homes)	Winter	16	ns	ns (detected in 4 of 16 homes- mean of 1.81 and range of 1.3–2.5 µg/m3)	Pleil et al. 1985
United States (various sites)	1970–1987	2120	ns	5.02 µg/m3	Shah and Heyerdahl 1988

Table A3. Concentration of 1,1-DCE in drinking water and groundwater

Location	Sampling period	Number of samples	Detection limit (µg/L)	Mean concentration[1](µg/L)	Reference
Drinking Water
Victoria, British Columbia	2008	2	0.1	ND	City of Victoria 2008
Vancouver, British Columbia	August 19, 2008	3	0.5	ND	City of Vancouver 2008
Toronto, Ontario	January–December 2008	ns	ns	ND	TDWS 2008
Niagara Falls, Ontario	November 6, 2008	1	0.41	ND	City of Niagara Falls 2008
Saskatoon, Saskatchewan	2008	1	0.2	ND	CSWTP 2008
London, Ontario	June 10, 2008	1	0.41	ND	City of London 2008
Kitchener, Ontario	January–November 2008	6	0.5	ND	OME 2008
Kingston, Ontario	2008	2	0.1	ND	Utilities Kingston 2008
Hamilton, Ontario	February–November 2008	ns	0.2	ND	BCOS 2008
Edmonton, Alberta	2008	ns	ns	ND	EPCOR 2008
Barrie, Ontario	2006	14	ns	ND –“< 0.41”	CBWO 2008
Montréal, Quebec	2006	ns	0.07	ND	Ville de Montreal 2006
Calgary, Alberta	2003	ns	0.5	ND	CCW 2003
Ottawa, Ontario	2003	35	0.52	ND	COWQS 2003
Québec, Quebec	February–November 2002	4	ns	< 0.2 (< 0.1 to < 0.4)	Ville de Québec 2002
United States	1985–2001	n = 1096 (public well samples)	MDL 0.047 (Connor et al. 1998)	< 0.16 (median of all samples) 0.20 (median of samples with detection)	Zogorski et al. 2006
United States	1985–2001	n = 2400 (domestic well samples)	MDL 0.047 (Connor et al. 1998)	<0.18 (median of all samples) 0.026 (median of samples with detection)	Zogorski et al. 2006
Toronto, Ontario	1986 and 1987	2 (tap water) 7 (bottled water)	0.04	tap water - ND bottled water - ND	City of Toronto 1990
Ontario (water treatment plants, various locations)	1987	44 treatment plants	0.1	raw - ND treated - ND distribution water - ND	OME 1988, 1989
29 Alberta municipal drinking water supplies	1978–1985	ns	ns	ns (detected in one of 29 municipal supplies at a max concentration of 1.4 µg/L	Health Canada 1994a
10 Ontario water treatment plants (Great Lakes locations)	July–August 1982 January–February 1983 April–May 1983	42 raw 42 treated	[0.1–0.4] [2]	raw - 0 [ND] treated- < 0.1 [ND–trace (1 sample at less than 0.1)]	Otson 1987
Canada-wide (29 municipalities, 30 water treatment plants)	August–September 1979	30 raw 30 treated	5.0 (MQL)	raw - 0 [ND] treated - < 1 [ND–~20]	Otson et al. 1982b
Canada-wide (29 municipalities, 30 water treatment plants)	November–December 1979	30 raw 30 treated	5.0 (MQL)	raw - 0 [ND] treated - 0 [ND]	Otson et al. 1982b
United States (EPA survey)	ns	ns	ns	detected in 3% of drinking water supplies; 0.3 µg/L (0.2–0.5 µg/L)	US EPA 1985 (cited in ATSDR 1994)
Groundwater
United States	1985–2001	3497	ns	< 0.20 (median for all samples) 0.068 (median for samples with detection)	Zogorski et al. 2006
Ottawa, Ontario	May 1988	37	ns	detected in 43% of samples [0.9–60]	Lesage et al. 1990
United States (community-based groundwater sources, nationwide survey)	ns	945	0.2 (MQL)	detected in 2.3% of samples (max. 6.3 µg/L, subset median values, 0.28–1.2 µg/L)	Rajagopal and Li 1991 [cited in ATSDR 1994] Westrick et al. 1984 [cited in ATSDR 1994]
United States, Ground Water Supply Survey	1982	466	ns	detected in 9 samples; 0.3 µg/L (median)	Cotruvo 1985 [cited in ATSDR 1994]

Table A4. Concentration of 1,1-DCE in foodstuffs

Item sampled	Sampling period	No. of samples	Detection limit	Mean concentration (µg/kg)	Reference
Food
Ville-Mercier, Quebec Ice cream Cheese and butter Beef and veal Pork/cured pork Lamb chops Poultry Eggs Organ meats Luncheon meats Canned meats Marine fish Freshwater fish Canned fish Shellfish Canned meat soups Canned pea and tomato soups Dehydrated soups Bread Flour and cakes Cereals Pies Pasta Potatoes and vegetables Rice and vegetables Beets and tomatoes Fruits Juices and canned fruit Oils and fats Peanut butter	January 1993	4	5.0 µg/kg	ND	ETL 1993
Ville-Mercier, Quebec Dairy Coffee and tea Soft drinks Alcohols Water	January 1993	4	1.0 µg/L	ND	ETL 1993
United Kingdom Biscuits Marshmallow Swiss roll Snack biscuits Crisps and snack foods Whole turkey Black pudding Smoked cheese Liver pate Cooked sausage	ns	ns	1 µg/kg	< 1 < 1 < 1 < 1 < 1 < 1 6 < 1 5 5	MAFF 1980
United States (food simulants) heptane (0.5-mm film) corn oil (0.5-mm film) water (0.5-mm film) Note: 0.5-mm film is equivalent thickness to plastic wrap used for food applications.	1977	n = 4 n = 5 n = 4	5–10 ppb	39 ppb (34–44 ppb) 34 ppb (18–41 ppb) 25 ppb (24–27 ppb)	Hollifield and McNeal 1978
Great Britain (potato crisps)	1979	n = 4	0.005 ppm	0.019 ppm (0.010–0.025 ppm)	Gilbert et al. 1980
Great Britain Biscuits Cakes Snack products Cheeses Cooked meats: Black pudding (n = 1) Liver pate (n = 1) Polony (n = 1) Bacon and liver pate (n = 1)	October 1978	n = 7 n = 1 n = 3 n =1	0.005 ppm	ND ND ND ND 0.005–0.01 ppm 0.005–0.01 ppm ND ns	Gilbert et al. 1980
Japan Sausage Fish sausage Boiled fish paste Cheese	August 2004	n = 13	0.001 µg/g	0.008 µg/g 0.005 µg/g 0.003 µg/g 0.0095 µg/g	Ohno and Kawamura 2006

Table A5. Concentration of 1,1-DCE in soil

Location	Sampling period	No. of samples	Detection limit (ng/g)[2]	Mean concentration[1](ng/g)	Reference
Soil
Ontario regions – urban parkland	ns (~1993)	59	MDL [4] = 2	0.074 [0.039–0.12][3]	OMEE 1993
Ontario regions – rural parkland (not including northwest region)	ns (~1993)	85	MDL [4] = 2	0.016 [0.010–0.024][3]	OMEE 1993
Ontario regions – rural parkland (northwest region)	ns (~1993)	17	MDL [4]= 2	0.097 [0.063–0.098][3]	OMEE 1993

Appendix 2: Upper-bounding Deterministic Estimate of 1,1-DCE Daily Intake (µg/kg-bw per Day by Various Age Groups)

Route of Exposure	0–6 months[1], [2], [3]		0.5–4 yr[4]	5–11 yr[5]	12–19 yr[6]	20–59 yr[7]	60+ yr[8]
	Formula fed	Not formula fed
Ambient air[9]	0.00		0.00	0.00	0.00	0.00	0.00
Indoor air[10]	0.00		0.01	0.01	0.00	0.00	0.00
Drinking water[11]	0.06	0.01	0.01	0.01	0.00	0.00	0.00
Food and beverages[12]		1.31	0.86	0.55	0.32	0.24	0.20
Soil[13]	0.00		0.00	0.00	0.00	0.00	0.00
Total intake	0.06	1.32	0.87	0.56	0.33	0.25	0.20
Maximum total intake from all routes of exposure:							1.32

Appendix 3: Summary of Health Effects Information for 1,1-dichloroethene

Endpoint	Lowest effect levels[1]/ Results
Acute toxicity	Lowest inhalation LC50(mouse) = 200 mg/m3 (Zeller et al. 1979a, 1979b, 1979c, 1979d) [Additional studies: Carpenter et al. 1949; Siegel et al. 1971; Jaeger et al. 1973, 1974; Klimisch and Freisberg, 1979a, 1979b; Zeller et al. 1979a, 1979b, 1979c, 1979d] Lowest oral LD50 (mouse) = 194 mg/kg-bw (Jones and Hathway 1978a) [Additional studies: Jenkins et al. 1972; Andersen and Jenkins 1977; Ponomarkov and Tomatis 1980]
Short-term repeated-dose toxicity	Lowest inhalation LOEC (rat) = 200 mg/m3: fatty changes and focal liver cell necrosis (4 weeks) (Plummer et al. 1990); changes to the liver and kidneys (7 days, with observation period to 28 days) (Maltoni and Patella 1983) [Additional studies: Gage 1970; Short et al. 1977; Oesch et al. 1983; Norris and Reitz 1984] Lowest oral LOEL (gavage) (rat) = 200 mg/kg-bw (2 times per week): increased serum sorbitol dehydrogenase and aminotransferases indicative of hepatotoxicity (4 weeks) (Siegers et al. 1983) [Additional studies: NTP 1982; Maltoni and Patella 1983]
Subchronic toxicity	Lowest inhalation LOEC (rat) = 100 mg/m3: minimal, reversible liver cell cytoplasmic vacuolation (90 days) (Norris 1977; Quast et al. 1977) [Additional studies: Lazarev 1960; Prendergast et al. 1967] Lowest oral LOEL (rat) = 19 mg/kg-bw per day: minimal, recoverable liver cell cytoplasmic vacuolation (90 days) (Norris 1977; Quast et al. 1977) [Additional studies: NTP 1982; Quast et al. 1983]
Chronic toxicity/ carcinogenicity	Lowest inhalation LOAEC (mice) = 40 mg/m3: significant increases in kidney damage (regressive changes and/or abscesses and nephritis in males) (52 weeks) (Maltoni et al. 1984, 1985) [Additional studies: Lee et al. 1977; Rampy et al. 1977, 1978; Viola and Caputo 1977; Hong et al. 1981; Quast et al. 1986; Cotti et al. 1988] Lowest oral LOEL (rat) = 5 mg/kg-bw per day: increased incidence of chronic renal inflammation in male and female F344/N rats, 2-year gavage study (NTP 1982) [Additional studies: Ponomarkov and Tomatis 1980; Quast et al. 1983; Maltoni et al. 1984, 1985] Inhalation study in Swiss mice: 0, 10 or 25 ppm (0, 40 or 100 mg/m3; conversion by IPCS 1990) for 52 weeks; significantly increased incidence of renal adenocarcinomas (0/126, 0/25 and 28/119 for the control, low and high concentrations, respectively) in males at 100 mg/m3; mammary carcinomas (3/185, 6/30 and 16/148 for the control, low and high concentrations, respectively) in females and pulmonary adenomas (12/331, 14/58 and 41/288 for the control, low and high concentrations, respectively) in males and females were not clearly exposure-related (Maltoni et al. 1984, 1985) No significant increases in tumours considered to be related to exposure were observed in rats or hamsters in inhalation bioassays or in any species in studies by oral, dermal or subcutaneous routes of exposure (Lee et al. 1977, 1978; Rampy et al. 1977, 1978; Viola and Caputo 1977; Van Duuren et al. 1979; Hong et al. 1981; NTP 1982; Quast et al. 1983, 1986; Maltoni et al. 1984, 1985). Dermal initiation–promotion study in female mice: initiation with 1,1-DCE; promotion by phorbol myristate acetate for 428–576 days, beginning 14 days after exposure to 1,1-DCE; 8/30 treated mice with lung papillomas versus 9/120 controls (Van Duuren et al. 1979)
Developmental toxicity	Lowest inhalation LOAEC (mouse) = 60 mg/m3: significant increase in the mean number of fetuses with an unossified incus and incompletely ossified sternebrae (gestation days 6–16); maternal LOEC = 119 mg/m3, based upon decrease in weight gain (Short et al. 1977) [Additional studies: Murray et al. 1979] Lowest oral LOEL (maternal, rat) = 14 mg/kg-bw per day; dams: minimal hepatocellular fatty change; reversible, accentuated hepatic lobular pattern; pups: no effects were observed (three-generation study) (Nitschke et al. 1983) Note: Although 0.02 mg/kg-bw per day (rat) was the lowest identified oral LOAEL (Dawson et al. 1993), based on several factors, the US EPA (2002a) could not conclude that exposure to 1,1-DCE caused these effects. [Additional studies: Murray et al. 1979]
Genotoxicity and related endpoints: in vivo	Chromosomal aberrations Positive results: Hamster, bone marrow (Hofmann and Peh 1976) [inhalation; 120 or 400 mg/m3, 6 hours/day, 5 days/week, 6 weeks] Negative results: Rat, bone marrow (Rampy et al. 1977) [inhalation; 100 or 300 mg/m3, 6 hours/day, 5 days/week, 6 months]; mouse, bone marrow (Cerna and Kypenova 1977) [intraperitoneal injection for 5 days] DNA adduct formation Positive results: CD1-mice [inhalation; 40 or 200 mg/m3, 6 hours], Sprague-Dawley rats [inhalation, 40 mg/m3, 6 hours], liver and kidney (Reitz et al. 1980) Dominant lethal test Negative results: Mouse (Andersen and Jenkins 1977) [inhalation; 50 ppm (198 mg/m3), 6 hours/day, 5 days]; rat (Short et al. 1977) [inhalation; 55 ppm (218 mg/m3), 6 hours/day, 5 days/week, 11 weeks] Micronuclei test Negative results: Mouse, bone marrow [oral, 200 mg/kg-bw]; mouse, fetal erythrocytes [oral, 100 mg/kg-bw] (Sawada et al. 1987) Non-mammalian sex-linked recessive lethal assay Negative results: Drosophila (Foureman et al. 1994) [oral, 20 000 or 25 000 ppm, 72 hours; or injection, 5000 ppm, 24 hours] Unscheduled DNA synthesis Positive results: CD-1 mice, liver and kidney (Reitz et al. 1980) [inhalation; 200 mg/m3, 6 hours]
Genotoxicity and related endpoints: in vitro	Aneuploidy Positive results: Saccharomyces cerevisiae, with and without activation (Koch et al. 1988) Chromosomal aberrations Positive results: Chinese hamster lung cells, with activation (Sawada et al. 1987) Negative results: Chinese hamster lung cells, without activation (Sawada et al. 1987); Chinese hamster fibroblast CHL cells (Ishidate 1983); Chinese hamster DON-6 cells (Sasaki et al. 1980) Gene conversion Positive results: S. cerevisiae, without activation (Koch et al. 1988);S. cerevisiae, with activation (Bronzetti et al. 1981) Negative reults: S. cerevisiae, without activation (Bronzetti et al. 1981); S. cerevisiae, with activation (Koch et al. 1988) Mutagenicity Positive results: Salmonella typhimurium BA13/BAL13, with activation (Roldan-Arjona et al. 1991) S. typhimurium TA100, with activation (Bartsch et al. 1975, 1979; Baden et al. 1976, 1978, 1982; Jones and Hathway 1978b; Simmon and Tardiff 1978; Waskell 1978; Oesch et al. 1983; Strobel and Grummt 1987; Malaveille et al. 1997) S. typhimurium TA100, without activation (Baden et al. 1976, 1978, 1982; Cerna and Kypenova 1977; Waskell 1978; Strobel and Grummt 1987) S. typhimurium TA1535, with activation (Baden et al. 1977; Jones and Hathway 1978b; Oesch et al. 1983) S. typhimurium TA1535, without activation (Cerna and Kypenova 1977) S. typhimurium TA1537, with activation (Oesch et al. 1983) S. typhimurium TA1538, without activation (Cerna and Kypenova 1977) S. typhimurium TA98, with activation (Oesch et al. 1983; Strobel and Grummt 1987) S. typhimurium TA98, without activation (Cerna and Kypenova 1977) S. typhimurium TA92, with activation (Oesch et al. 1983) S. typhimurium TA97, with activation (Strobel and Grummt 1987) Escherichia coli K12, with activation (Oesch et al. 1983) E. coli K12, without activation (Greim et al. 1975) E. coli WP2, with activation (Oesch et al. 1983) S. cerevisiae, with activation (Bronzetti et al. 1981; Koch et al. 1988); S. cerevisiae, without activation (Koch et al. 1988) Mouse lymphoma L5178Y T/K +/- cells, with activation (McGregor et al. 1991) Negative results: S. typhimurium BA13/BAL13, without activation (Roldan-Arjona et al. 1991) S. typhimurium TA100, with activation (Mortelmans et al. 1986) S. typhimurium TA100, without activation (Bartsch et al. 1975, 1979; Simmon and Tardiff 1978; Oesch et al. 1983; Mortelmans et al. 1986) S. typhimurium TA104, with and without activation (Strobel and Grummt 1987) S. typhimurium TA1535, with activation (Mortelmans et al. 1986) S. typhimurium TA1535, without activation (Baden et al. 1977; Oesch et al. 1983; Mortelmans et al. 1986) S. typhimurium TA1537, with activation (Mortelmans et al. 1986) S. typhimurium TA1537, without activation (Oesch et al. 1983; Mortelmans et al. 1986) S. typhimurium TA98, with activation (Mortelmans et al. 1986) S. typhimurium TA98, without activation (Oesch et al. 1983; Mortelmans et al. 1986; Strobel and Grummt 1987) S. typhimurium TA92, without activation (Oesch et al. 1983) S. typhimurium TA97, without activation (Strobel and Grummt 1987) E. coli K12, without activation (Oesch et al. 1983) E. coli WP2, without activation (Oesch et al. 1983) Chinese hamster lung V79 cells, hprt locus, with and without activation (Drevon and Kuroki 1979) Chinese hamster lung V79 cells, ouabain resistance, with and without activation (Drevon and Kuroki 1979) Sister chromatid exchange Positive results: Chinese hamster lung cells, with activation (Sawada et al. 1987); Chinese hamster ovary cells (McCarroll et al. 1983) Negative results: Chinese hamster lung cells, without activation (Sawada et al. 1987) Unscheduled DNA synthesis Positive results: Rat, hepatocytes (Costa and Ivanetich 1982)
Metabolism	1,1-DCE is rapidly absorbed following inhalation and oral exposures. The major route of excretion for unchanged 1,1-DCE is through the lung. Intraperitoneal (i.p.) administration of 125 mg/kg 14C-1,1-DCE to mice resulted in the highest concentrations of covalent binding (based on protein content) in the kidney, lung and liver. The covalent binding and cellular damage in kidney, lung and liver correlated with the high concentration of CYP2E1 Oxidation of 1,1-DCE by CYP2E1 should produce three metabolites: 1,1-DCE epoxide, 2-chloroacetyl chloride, and 2,2-dichloroacetaldehyde.The epoxide, and perhaps to a lesser extent the chloroacetaldehyde, are believed to be associated with the tissue reactivity and toxic effects in tissues that ensue after significant depletion of GSH. 1,1-DCE will not bioaccumulate in tissues to a significant extent. When the inhalation exposure was less than 100 ppm, the estimated amount of epoxide formed was fivefold lower in humans than in rats (US EPA 2002b).
Epidemiology	Cohort of 138 U.S. workers exposed to 1,1-DCE, where vinyl chloride was not used as a copolymer. Twenty-seven workers were lost to follow-up but considered to be alive in the analyses. Fifty-five people had less than 15 years since first exposure, and only five deaths were observed. The authors indicate no finding was statistically attributable to exposure to 1,1-DCE (Ott et al. 1976). Cohort of 629 males (447 German and 182 foreign workers) employed at two plants in the Federal Republic of Germany that had produced 1,1-DCE since 1955. Vital status was ascertained for 97% of the 447 German workers. Of the 182 foreign workers, 65 had worked for less than one year, and only 24% (44) were traced. Observed deaths were compared with local and regional rates, without making allowance for a latent period. Within the study period (approximately 20 years), 39 deaths were observed, where 57 [local] and 36 [regional] would have been expected. Five cases of lung carcinoma were observed, whereas 3.9 [local] and 2.2 [regional] were expected; this result was not statistically significant. Workers in the factory were also potentially exposed to vinyl chloride and acrylonitrile (Thiess et al. 1979). The International Agency for Research on Cancer Working Group noted that both Ott et al. (1976) and Thiess et al. (1979) suffered from the limited size of cohorts, the short observation period and the small numbers of deaths from specific causes. The fact that no allowance was made for latent period may have resulted in an overestimation of the expected numbers and an underestimation of risk. In an attempt to identify the specific exposure associated with an excess lung cancer risk noted previously in a U.S. synthetic chemicals plant, Waxweiler et al. (1981) considered 19 chemicals, one of which was 1,1-DCE. Company personnel assigned a rank of exposure to 1,1-DCE (from 0 to 5) to each job in the plant for each year since its opening in 1942. These exposure data were then linked with detailed, individual work histories to obtain an individual estimate for each of the 4806 male workers employed at the plant. The doses calculated were the product of the exposure rank of the job and the number of days worked at that job. Cumulative doses for 45 workers who had died of lung cancer during the study period of 1942–1973 were then compared to expected doses based on the cumulative exposure of subcohorts of fellow workers matched individually to the cases by year of birth and age of hire into the plant. This comparison failed to suggest any specific association between exposure to 1,1-DCE in the plant and excess lung cancer risk.

Appendix 4: Robust Study Summaries

Table A6. Robust Study Summary – Aquatic Toxicity – Alga

No.	Item	Weight	Yes/No	Specify
1	Reference: Brack W, Rottler H. 1994. Toxicity testing of highly volatile chemicals with green algae – a new assay. Environ Sci Pollut Res 1(4):223–228.
2	Substance identity: 75-35-4	n/a[1]	Y
3	Substance identity: 1,1-dichloroethylene	n/a	Y	The test substance name
4	Chemical composition of the substance	2	Y	The test substance name
5	Chemical purity	1	Y	> 99%
6	Persistence/stability of test substance in aquatic solution reported?	1	Y	Data are available but not included in the study. See Table 4a
Method
7	Reference	1	N	New approach test
8	OECD, EU, national, or other standard method?	3	Y	Based on OECD tests
9	Justification of the method/protocol if a non-standard method was used	2	Y
10	GLP (good laboratory practice)	3	N	n/a
Test organism
11	Organism identity: Chlamydomonas reinhardtii	n/a	Y	Green alga
12	Latin or both Latin and common names reported?	1	Y	Green alga
13	Life cycle age / stage of test organism	1		n/a
14	Length and/or weight	1		n/a
15	Sex	1		n/a
16	Number of organisms per replicate	1		n/a
17	Organism loading rate	1		NA[2]
18	Food type and feeding periods during the acclimation period	1	Y	Light, CO2 source
Test design / conditions
19	Test type (acute or chronic)	n/a	Y	Acute
20	Experiment type (laboratory or field)	n/a	Y	Lab
21	Exposure pathways (food, water, both)	n/a	Y	Water
22	Exposure duration	n/a	Y	72 hrs
23	Negative or positive controls (specify)	1	Y	Negative
24	Number of replicates (including controls)	1	Y
25	Nominal concentrations reported?	1	N
26	Measured concentrations reported?	3	Y
27	Food type and feeding periods during the long-term tests	1		n/a
28	Were concentrations measured periodically (especially in the chronic test)?	1	Y
29	Were the exposure media conditions relevant to the particular chemical reported? (e.g. for the metal toxicity – pH, DOC/TOC, water hardness, temperature)	3	Y
30	Photoperiod and light intensity	1	Y
31	Stock and test solution preparation	1	Y
32	Was solubilizer/emulsifier used if the chemical was poorly soluble or unstable?	1		n/a
33	If solubilizer/emulsifier was used, was its concentration reported?	1		n/a
34	If solubilizer/emulsifier was used, was its ecotoxicity reported?	1		n/a
35	Monitoring intervals (including observations and water quality parameters) reported?	1	Y
36	Statistical methods used	1	Y
Information relevant to the data quality
37	Was the endpoint directly caused by the chemical's toxicity, not by the organism’s health (e.g. when mortality in the control > 10%) or physical effects (e.g. “shading effect”)?	n/a	Y
38	Was the test organism relevant to the Canadian environment?	3	Y
39	Were the test conditions (pH, temperature, DO, etc.) typical for the test organism?	1	Y
40	Do system type and design (static, semi-static, flow-through; sealed or open; etc.) correspond to the substance's properties and organism's nature/habits?	2	Y
41	Was pH of the test water within the range typical for the Canadian environment (6 to 9)?	1	Y
42	Was temperature of the test water within the range typical for the Canadian environment (5 to 27°C)?	1	Y
43	Was toxicity value below the chemical’s water solubility?	3	Y
Results
44	Toxicity values (specify endpoint and value)	n/a	EC10, EC50	Growth
45	Other endpoints reported - e.g. BCF/BAF, LOEC/NOEC (specify)?	n/a	N
46	Other adverse effects (e.g. carcinogenicity, mutagenicity) reported?	n/a	N
47	Score: ... %	35/40 = 87.5
48	EC reliability code:	1
49	Reliability category (high, satisfactory, low):	High Confidence
50	Comments

Table A7. Robust Study Summary – Terrestrial Toxicity – Mammals

No.	Item	Weight	Yes/No	Specify
1	Reference: Prendergast J, Jones R, Jenkins Jr L, Siegel J. 1967. Effects on experimental animals of long-term inhalation of trichloroethylene, carbon tetrachloride, 1,1,1-trichloroethane, dichlorofluoromethane and 1,1-dichloroethylene. Toxicol Appl Phamacol 10:270–289.
2	Substance identity: CAS RN	n/a[1]	N
3	Substance identity: 1,1-dichloroethylene	n/a	Y	The test substance name
4	Chemical composition of the substance	2	Y	The test substance name
5	Chemical purity	1	Y	Reagent grade
6	Persistence/stability of test substance?	1	Y	Data are available but not included in the study. See Table 4a
Method
7	Reference	1	Y
8	OECD, EU, national, or other standard method?	3	N
9	Justification of the method/protocol if a non-standard method was used	2	Y
10	GLP (good laboratory practice)	3		NA[2]
Test organism
11	Organism identity: rats (Sprague-Dawley or Long-Evans), guinea pigs (Hartley), squirrel monkeys, rabbits (New Zealand albino), beagle dogs	n/a	Y
12	Latin or both Latin and common names reported?	1	N
13	Life cycle age / stage of test organism	1	N
14	Length and/or weight	1	Y	Trends noted
15	Sex	1		n/a to study
16	Number of organisms per replicate	1	Y
17	Organism loading rate	1	Y
18	Food type and feeding periods during the acclimation period	1	Y
Test design / conditions
19	Test type (acute or chronic)	n/a	Y	90-day inhalation or "work week" inhalation
20	Experiment type (laboratory or field)	n/a	Y	Lab
21	Exposure pathways (food, water, both)	n/a	Y	Air
22	Exposure duration	n/a	Y	90-day or 5-day
23	Negative or positive controls (specify)	1	Y	Negative
24	Number of replicates (including controls)	1	Y
25	Nominal concentrations reported?	1	N
26	Measured concentrations reported?	3	Y	Continuous
27	Food type and feeding periods during the long-term tests	1	Y
28	Were concentrations measured periodically (especially in the chronic test)?	1	Y	Continuous
29	Were the exposure media conditions relevant to the particular chemical reported? (e.g. for the metal toxicity – pH, DOC/TOC, water hardness, temperature)	3	Y
30	Photoperiod and light intensity	1		n/a to study
31	Stock and test solution preparation	1	Y
32	Was solubilizer/emulsifier used if the chemical was poorly soluble or unstable?	1		n/a
33	If solubilizer/emulsifier was used, was its concentration reported?	1		n/a
34	If solubilizer/emulsifier was used, was its ecotoxicity reported?	1		n/a
35	Monitoring intervals (including observations and water quality parameters) reported?	1	Y
36	Statistical methods used	1	Y
Information relevant to the data quality
37	Was the endpoint directly caused by the chemical's toxicity, not by the organism’s health (e.g. when mortality in the control > 10%) or physical effects (e.g. “shading effect”)?	n/a	Y
38	Was the test organism relevant to the Canadian environment?	3	Y
39	Were the test conditions (pH, temperature, DO, etc.) typical for the test organism?	1	Y
40	Do system type and design (static, semi-static, flow-through; sealed or open; etc.) correspond to the substance's properties and organism's nature/habits?	2	Y
41	Was pH of the test water within the range typical for the Canadian environment (6 to 9)?	1		n/a in air
42	Was temperature of the test water within the range typical for the Canadian environment (5 to 27°C)?	1		n/a in air
43	Was toxicity value below the chemical’s water solubility?	3		n/a in air
Results
44	Toxicity values (specify endpoint and value)	n/a	Y	90-day LC50
45	Other endpoints reported - e.g. BCF/BAF, LOEC/NOEC (specify)?	n/a	Y	NOEL = 101 mg/m3; LOEL = 189 mg/m3
46	Other adverse effects (e.g. carcinogenicity, mutagenicity) reported?	n/a	N
47	Score: ... %	30/36 = 83.3
48	EC reliability code:	1
49	Reliability category (high, satisfactory, low):	High Confidence
50	Comments

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Footnotes

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