Appendix B Toxicity Summaries - Orica

Consolidated Human Health Risk Assessment - 2017 Ref: OB/15/CHHRAR001-D

Appendix B Toxicity Summaries


Table of Contents

B1 Introduction ................................................................................................................................. B-1 B2 Carbon Tetrachloride .................................................................................................................... B-2 B3 Chloroform .................................................................................................................................... B-8 B4 Dichloromethane ........................................................................................................................ B-14 B5 1,1,2,2-Tetrachloroethane .......................................................................................................... B-20 B6 1,1,2-Trichloroethane ................................................................................................................. B-25 B7 1,2-Dichloroethane (EDC) .......................................................................................................... B-29 B8 Tetrachloroethene (PCE) ........................................................................................................... B-34 B9 Trichloroethene (TCE) ................................................................................................................ B-40 B10 1,1-Dichloroethane .................................................................................................................... B-46 B11 1,1-Dichloroethene .................................................................................................................... B-49 B12 cis- and trans-1,2-Dichloroethene ............................................................................................. B-53 B13 Vinyl Chloride ............................................................................................................................ B-58 B14 Hexachlorobenzene (HCB) ....................................................................................................... B-63 B15 Hexachlorobutadiene (HCBD) ................................................................................................... B-68 B16 Hexachloroethane (HCE) .......................................................................................................... B-73 B17 Mercury ...................................................................................................................................... B-77


Glossary of Terms

ADI Acceptable Daily Intake ADWG Australian Drinking Water Guidelines ANZECC Australia and New Zealand Environment and Conservation Council ATDS Australian Total Diet Survey ATSDR Agency for Toxic Substances and Disease Registry BMD Benchmark Dose BTEX Benzene, toluene, ethylbenzene and total xylenes CCME Canadian Council of Ministers of the Environment CICAD Concise International Chemicals Assessment Document CNS Central Nervous System EHC Environmental Health Criteria EPA Environment Protection Authority FSANZ Food Standards Australia New Zealand HEC Human Equivalent Concentration HED Human Equivalent Dose HIL Health Investigation Level HSDB Hazardous Substances Data Bank HSL Health Screening Level IARC International Agency for Research on Cancer IRIS Integrated Risk Information System JECFA Joint FAO/WHO Expert Committee on Food Additives JMPR WHO/FAO Joint Meeting on Pesticide Residues LOAEL Lowest-Observed-Adverse-Effect Level LOEL Lowest-Observed-Effect Level MF Modifying Factor MOA Mode (or Mechanism) of Action NEPC National Environment Protection Council NEPM National Environment Protection Measure NHMRC National Health and Medical Research Council NOAEL No-Observed-Adverse-Effect Level NOEL No-Observed-Effect Level NSW DECCW

New South Wales Department of Environment, Climate Change and Water

OCS Office of Chemical Safety PAH Polycyclic aromatic hydrocarbon PTDI Provisional Tolerable Daily Intake PTWI Provisional Tolerable Weekly Intake RAIS Risk Assessment Information System RfC Reference Concentration RfD Reference Dose SF Slope Factor


TC Tolerable Concentration TCE Trichloroethene TDI Tolerable Daily Intake TPH Total petroleum hydrocarbons TPHCWG Total Petroleum Hydrocarbon Criteria Working Group UF Uncertainty Factor UR Unit Risk USEPA United States Environmental Protection Agency VC Vinyl chloride VOC Volatile Organic Compound WHO World Health Organisation WHO DWG World Health Organisation Drinking Water Guidelines

Consolidated Human Health Risk Assessment - 2017 B-1 | P a g e Ref: OB/15/CHHRAR001-D

B1 Introduction This appendix presents toxicity summaries relevant to the CoPC identified in the 2017 CHHRA. The CoPC identified and considered in the 2017 CHHRA are the same as considered in the 2010 CHHRA.

It is noted that while much of the general chemical information remains the same, the review has considered all available toxicity data, including more recent information (where relevant) and ensured that the assessment of toxicity is consistent with guidance available in the NEPM (NEPC 1999 amended 2013a) and enHealth (enHealth 2012).


B2 Carbon Tetrachloride B2.1 General Carbon Tetrachloride (also known as carbona, carbon chloride, tetrachloromethane, carbon tet, methane tetrachloride, perchloromethane, tetrachlorocarbon and CTC) is predominantly a man-made compound, however, it has been detected in volcanic emission gases. It has also been suggested that carbon tetrachloride can be formed in the troposphere by solar induced photochemical reactions of chlorinated alkenes (WHO 1999a). Production of carbon tetrachloride began in about 1907 in the US. Since 1990 the production of carbon tetrachloride has dropped due to the Montreal protocol which established a phase-out by 1996 of the production of carbon tetrachloride and chloroflurocarbons (CFCs) by major manufacturing countries. Most of the carbon tetrachloride produced is used in the production of CFCs, which were primarily used as refrigerants, propellants, foam-blowing agents and solvents and in the production of other chlorinated hydrocarbons. Carbon tetrachloride has also been used as a grain fumigant, pesticide, solvent for oils and fats, metal degreaser, fire extinguisher and flame retardant, and in the production of paint, ink, plastics, semi-conductors and petrol additives. It was previously also widely used as a cleaning agent. All these uses have tended to be phased-out as production has dropped (WHO 1999a).

B2.2 Properties Carbon tetrachloride is a clear, colourless, volatile liquid with a characteristic, sweet odour. It is miscible with most aliphatic solvents and is itself a solvent. The solubility in water is low. Carbon tetrachloride is non-flammable and is stable in the presence of air and light. Decomposition may produce phosgene, carbon dioxide and hydrochloric acid. Key properties are presented below (ATSDR 2005; RAIS):

CAS No 56-23-5 Chemical Formula CCl4 Molecular Weight 153.8 Vapour Pressure 115 mmHg at 20oC Vapour Density 5.32 Density 1.59 g/ml at 25oC Solubility 793 mg/L at 20oC Air Diffusion Coefficient 0.057 cm2/s Water Diffusion Coefficient 8.788 x 10-6 cm2/s Henry’s Law Coefficient 0.03 atm.m3/mol

= 1.13 at 25oC (unitless) Koc 43.9 cm3/g Log Kow 2.83 Odour Threshold 10 - 71,000 mg/m3 Permeability Constant 0.0163 cm/hr Conversion Factors 1 ppm = 6.39 mg/m3 in air (25oC)


B2.3 Exposure Exposure of the general population to carbon tetrachloride maybe by inhalation, oral or dermal routes. Inhalation is expected to be the major route of exposure, particularly in occupational environment, but also in the general population. Dermal contact has not been shown to be a significant route of exposure to carbon tetrachloride (ATSDR 2005). NHMRC indicate that concentrations of carbon tetrachloride in major Australian reticulated supplies are significantly less than 0.001 mg/L (NHMRC 2011 Updated 2016).

In relation to the assessment of dermal absorption, dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance (USEPA 1995a) of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). No data is available specifically for the dermal absorption of CTC and it has a vapour pressure similar to that of benzene; hence a value of 0.05% is considered relevant.

If released into the environment the following can be noted with respect to carbon tetrachloride (WHO 1999a):

Air: Nearly all carbon tetrachloride released to the environment will ultimately be present in the atmosphere, due to its volatility. Since the atmospheric residence time of carbon tetrachloride is long, it is widely distributed. Estimates of atmospheric lifetime are variable, but 45-50 years is accepted as the most reasonable value. Carbon tetrachloride contributes both to ozone depletion and to global warming.

Soil and Water: Following releases to soil, most carbon tetrachloride is expected to evaporate rapidly due to its high vapour pressure. A small fraction of carbon tetrachloride may adsorb to organic matter. Carbon tetrachloride is expected to be moderately mobile in most soils, depending on organic carbon content, and leaching to groundwater may occur. Carbon tetrachloride introduced into water resources is transported by movement of surface water and groundwater. Because of its volatility, evaporation is considered to be the main process for the removal of carbon tetrachloride from aquatic systems. The amount of carbon tetrachloride dissolved in the oceans is reported to be less than 1-3% of that in the atmosphere.

Biodegradation: Carbon tetrachloride is very stable in the troposphere primarily because carbon tetrachloride, in contrast to most other volatile halocarbons, has low reactivity towards hydroxyl radicals. The principal degradation process for carbon tetrachloride occurs in the stratosphere, where it is dissociated by short wave length (190- 220 nm) UV radiation to form the trichloromethyl radical and chlorine atoms with an estimated a half-life of 18-80 years for this photo dissociation process. Carbon tetrachloride dissolved in water does not photodegrade or oxidize in any measurable amounts with the rate of hydrolysis calculated with a half-life of 7000 years (concentration of 1 ppm). Carbon tetrachloride has been shown to be resistant to aerobic biodegradation, however biodegradation may occur within 16 days under anaerobic conditions. Carbon tetrachloride may undergo reductive dechlorination to form chloroform and other products in the presence of free sulphide and ferrous ions


Carbon tetrachloride has a low tendency to bioconcentrate in aquatic or marine organisms. Most animals readily metabolise and excrete carbon tetrachloride following exposure and hence biomagnification is not expected.

B2.4 Background Exposures/Intake Intake of carbon tetrachloride from soil, water and food can be considered to be insignificant. Intakes from air can be calculated from urban air concentrations from a light industrial area in Brisbane (Hawas et al. 2001) which indicate a background concentration of 0.0025 mg/m3 (average) to 0.004 mg/m3 (max) which is approximately 40% to 65% of the tolerable concentration in air (equivalent to an ADI) as adopted from the WHO air quality guidelines (WHO 2000c). As the data from the Brisbane study is derived from an industrial area, it is considered relevant to use the average value measured rather than the maximum. However, if an area of interest is in an industrial area where background levels of carbon tetrachloride are expected to be elevated a higher background intake may be relevant. On the basis of average concentrations of carbon tetrachloride in air from this study, background intake can be assumed to be up to 40% of the TC (WHO 2000). As other sources of emission on the BIP have been included in the assessment separately, the average background intake is considered appropriate for other sources.

B2.5 Health Effects

General

The following information is available from (ATSDR 2005; WHO 1999a). There is no clinical disease which is unique to carbon tetrachloride toxicity.

Carbon tetrachloride is well absorbed from the gastrointestinal and respiratory tract in animals and humans. Dermal absorption of liquid carbon tetrachloride is possible, but dermal absorption of the vapour is slow. Carbon tetrachloride is distributed throughout the whole body, with highest concentrations in liver, brain, kidney, muscle, fat and blood. The parent compound is eliminated primarily in exhaled air, while minimal amounts are excreted in the faeces and urine.

Carbon tetrachloride has depressant effects on the central nervous system particularly following high levels of exposure. It can also produce irritation effects on the gastrointestinal tract and skin. Most other toxic effects associated with exposure to carbon tetrachloride are associated with it metabolism by mixed function cytochrome P-450 oxygenases.

The liver and kidney are target organs for carbon tetrachloride toxicity via oral and inhalation exposures. The severity of the effects on the liver depends on a number of factors such as species susceptibility, route and mode of exposure, diet or co-exposure to other compounds, in particular ethanol. Furthermore, it appears that pre-treatment with various compounds, such as phenobarbital and vitamin A, enhances hepatotoxicity, while other compounds, such as vitamin E, reduce the hepatotoxic action of carbon tetrachloride.

In humans, acute symptoms after carbon tetrachloride exposure are independent of the route of intake and are characterized by gastrointestinal and neurological symptoms, such as nausea, vomiting, headache, dizziness, dyspnoea and death. Liver damage appears after 24 h or more. Kidney damage is evident often only 2 to 3 weeks following the poisoning.


Epidemiological studies have not established an association between carbon tetrachloride exposure and increased risk of mortality, neoplasia or liver disease. Some studies have suggested an association with increased risk of non-Hodgkin's lymphoma and, in one study, with mortality and liver cirrhosis. However, not all of these studies pinpointed specific exposure to carbon tetrachloride, and the statistical associations were not strong.

Carcinogenicity and Genotoxicity

Human data on the carcinogenic potential of carbon tetrachloride are limited and there have been no conclusive associated between carbon tetrachloride exposure and cancer in humans. In experiments with mice and rats, carbon tetrachloride proved to be capable of inducing hepatomas and hepatocellular carcinomas. The doses inducing hepatic tumours were higher than those inducing cell toxicity. It is considered likely that the carcinogenicity of carbon tetrachloride is secondary to its hepatotoxic effects (WHO 1999a) and may be related to its metabolism (ATSDR 2005).

Carbon tetrachloride can induce embryotoxic and embryolethal effects, but only at doses that are maternally toxic, as observed in inhalation studies in rats and mice. In addition carbon tetrachloride is not teratogenic (WHO 1999a). (Baars et al. 2001) indicates that the available data show that carbon tetrachloride has no mutagenic end points, however, it does bind covalently to DNA in vitro. The data also indicate that the carcinogenic potency of carbon tetrachloride can only be noticed at dose levels with apparent hepatotoxicity.

Many genotoxicity assays have been conducted with carbon tetrachloride. On the basis of available data, carbon tetrachloride can be considered as a non-genotoxic compound (WHO 1999a). A detailed review of genotoxicity associated with carbon tetrachloride in the Stage 2 assessment (Woodward-Clyde 1996) supported this outcome.

Recent studies (Nagano, Kasuke et al. 2007; Nagano, K. et al. 2007), however, suggest clear evidence of carcinogenicity for carbon tetrachloride in rats and mice and suggest a cytotoxic-proliferative and genotoxic mode of action. However, these studies has not identified relevant approaches for quantifying such effects or the relationship with hepatotoxicty (noted previously). Hence it is considered relevant to assess exposures to carbon tetrachloride on the basis of threshold values derived to be protective of the most sensitive end-point, hepatotoxicity.

Sensitive Populations

Potential issues associated with exposures to CTC and susceptible population have been reviewed by the (USEPA 2010b). This review identified that the events involved in CTC liver toxicity and carcinogenicity involve metabolic and cellular processes common to cells at all life stages. Because metabolism is a hypothesised key event, heterogeneity in the human population in the microsomal enzymes responsible for CTC toxicity could affect susceptibility to CTC. There is no direct evidence for increased or decreased susceptibility to carbon tetrachloride in children. However, based on the level of hepatic enzymes in young children (lower) compared with adults, it is hypothesised that infants and children would be less susceptible to liver injury from CTC. No increased susceptibility has been identified for the developing foetus.


Fasting or food deprivation has been shown to increase the toxicity of carbon tetrachloride. Carbon tetrachloride toxicity is also affected by the level of antioxidants in the diet (USEPA 2010b).

Based on experimental findings from rodent studies, there is some reason to suspect that people with diabetes may have altered susceptibility to hepatotoxic effects from carbon tetrachloride (USEPA 2010b).

Factors that increase the expression of CYP2E1 or CYP3A are likely to increase susceptibility to carbon tetrachloride exposure (all other things being the same) because the relatively higher rate of metabolism on a per cell basis would significantly increase the rate of generation of trichloromethyl radicals in the liver and kidney. This includes heavy consumers of ethanol and co-exposure to other chemical inducers of CYP450 (such as acetone, methanol and aliphatic alcohols, MEK, MIBK, ketones, PCBs, DDT and some pesticides) (USEPA 2010b).

Classification

Carbon tetrachloride has been classified as a "probable" human carcinogen (Category B2) by the USEPA based on carcinogenicity in rats, mice and hamsters (USEPA 2010b).

IARC has classified carbon tetrachloride in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans and sufficient evidence in experimental animals for carcinogenicity (IARC 1999b).

The National Occupational Health and Safety Commission (NOHSC) as Category 2 carcinogen (probable human carcinogen) (Safe Work Australia). NICNAS has not classified carbon tetrachloride.

B2.6 Quantitative Toxicity Reference Values On the basis of the weight of evidence, CTC does not appear to have significant genotoxic potential so it is considered reasonable that a threshold approach is adopted for the characterisation of all health effects including carcinogenicity. The following quantitative values are available for CTC appropriate Australian and International sources:


Table B1 Summary of Published Toxicity Reference Values: Carbon Tetrachloride

Source Value Basis/Comments Australian ADWG (NHMRC 2011 Updated 2016)

TDI = 0.00086 mg/kg/day The current ADWG have derived a guideline of 0.003 mg/L for CTC based on a no effect level of 1.2 mg/kg/day based on a 90-day gavage study on mice and the application of 1000 safety factor and a 5/7 study duration adjustment factor. The NOEL adopted was consistent with that in the old WHO DWG (prior to revision to the values noted below). Hence the quantitative approach adopted in these guidelines is considered dated.

International WHO (WHO 1999a)

TDI = 0.00142 mg/kg/day TC = 0.0061 mg/m3

TDI based on a 12 week oral rat study (NOAEL of 1 mg/kg), 500 uncertainty factor and a 5/7 conversion. TC based on 90-day inhalation study on rats (NOAEL 6.1 mg/m3) and 100 uncertainty factor. Value is also published in the Air Quality Guidelines (WHO 2000a).

WHO DWG (WHO 2011)

TDI = 0.00142 mg/kg/day The WHO derives a guideline of 0.004 mg/L based on a TDI of 0.00142 mg/kg/day derived from a 12-week oral study in rats (as per (WHO 1999a)). It is noted that the guideline derived is lower than values calculated using linear extrapolation and a lifetime excess cancer risk of 10-4 to 10-6 and is considered to be protective of all health endpoints.

RIVM (Baars et al. 2001)


TDI derived based on a NOAEL of 1 mg/kg/day for hepatotoxic effects for semi-chronic oral exposures in rats and an uncertainty factor of 250. TC derived based on a NOAEL of hepatic effects for an inhalation exposure study over 200 days in rats and uncertainty factor of 100.

ATSDR (ATSDR 2005)

No chronic oral MRL Inhalation MRL =0.19 mg/m3

No chronic oral MRL has been established. The chronic inhalation MRL has been derived on the basis of a NOAEL of (adjusted) of 5.8 mg/m3 associated with liver effects in rats and an uncertainty factor of 30.

USEPA (USEPA 2010b)

RfD = 0.004 mg/kg/day RfC = 0.1 mg/m3

Oral reference dose (RfD) (revised in 2010) derived on the basis of a benchmark dose (adjusted) of 3.9 mg/kg/day associated with elevated serum SDH activity in rats and an uncertainty factor of 1000. The uncertainty factor includes a factor of 10 to protect sensitive individuals. Inhalation RfC (revised 2010) derived on the basis of a benchmark dose (human equivalent concentrations) of 14.3 mg/m3 associated with liver effects in a rat study and a 100 fold uncertainty factor. The uncertainty factor includes a factor of 10 to protect sensitive individuals. Non-threshold values are also available from the USEPA.

While most of the evaluations conducted have considered similar critical endpoints there are a wide range of toxicity reference values that have been derived, particularly for the assessment of inhalation exposures. While the USEPA values are based on the most recent review, the threshold values available from the WHO (WHO 1999a) are considered the most relevant for the assessment of both non-carcinogenic and carcinogenic endpoints.


B3 Chloroform B3.1 General Chloroform (also known as trichloromethane, methenyl chloride, methane trichloride, methyl trichloride and formyl trichloride, CFM) is both a synthetic and naturally occurring compound, with anthropogenic sources responsible for most of the chloroform in the environment. Chloroform is mainly used in the production of other materials, principally fluorocarbons, used in the synthesis of tetrafluoroethylene and polytetrafluoroethylene, and as a refrigerant and propellant. Chloroform is also widely employed as an organic solvent in industry and in analytical laboratories. It has also been used as an ingredient of pharmaceuticals, drugs, cosmetics, grain fumigants, dyes and pesticides.

In the past, chloroform has been extensively used as a surgical anaesthetic, but this use was discontinued because exposure to narcotic concentrations resulted in adverse side effects. The US Food and Drug Administration banned the use of chloroform as an ingredient in human drug and cosmetic products in 1976.

B3.2 Properties It is a colourless liquid with a pleasant, non-irritating odour and a slightly sweet taste. It is only slightly soluble in water, but is miscible with alcohol, benzene, ether, petroleum ether, carbon tetrachloride, carbon disulfide, and oils. Decomposition may produce phosgene, carbon dioxide and hydrogen chloride. Key properties are presented below (ATSDR 1997b; HSDB; RAIS):

CAS No 67-66-3 Chemical Formula CHCl3 Molecular Weight 119.38 Vapour Pressure 197 mmHg at 20oC Vapour Density 4.1 Density 1.48 g/ml at 25oC Solubility 7950 mg/L at 25oC Air Diffusion Coefficient 0.0769 cm2/s Water Diffusion Coefficient 1.09 x 10-5 cm2/s Henry’s Law Coefficient 0.00367 atm.m3/mol

= 0.15 at 25oC (unitless) Koc 31.8 cm3/g Log Kow 1.97 Odour Threshold 85 ppm (421 mg/m3 ) Permeability Constant 0.00683 cm/hr Conversion factor 1 mg/m3 = 0.202 ppm at 20oC and 101.3Pa

B3.3 Exposure Human exposure to chloroform can occur orally, dermally, or by inhalation. Chloroform is the principal trihalomethane generated as by-products during the chlorination of drinking water. The primary sources of chloroform in the environment are chlorinated drinking water and wastewater, pulp and paper mills, and chemical and pharmaceutical manufacturing plants. The general population is exposed to chloroform mainly in food, drinking-water and indoor air. Most of the chloroform released to the environment eventually enters the atmosphere, while much smaller amounts enter groundwater as a result of filtration through the soil. NHMRC indicate that


concentrations of total trihalomethanes (including chloroform) in major Australian reticulated supplies range up to 0.6 mg/L (NHMRC 2011 Updated 2016).

In relation to the assessment of dermal absorption, as noted in USEPA (2004), dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance (USEPA 1995a) of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). No data is available specifically for the dermal absorption of chloroform and it has a vapour pressure similar to that of benzene; hence a value of 0.05% is considered relevant.

If released into the environment the following can be noted with respect to chloroform (WHO 1994a, 2004b):

Air: Nearly all chloroform released to the environment will ultimately be present in the atmosphere due to its volatility. In the atmosphere, chloroform may be transported long distances before degrading via indirection photochemicals reactions with free radicals such as hydroxyl (which form low levels of phosgene and hydrogen chloride). Half-lives vary from 55 to 20 days.

Soil and Water: Following releases to soil, most chloroform is expected to evaporate rapidly due to its high volatility and low soil adsorption. Most of the remaining chloroform will travel through the soil because of its low adsorption onto soils with leaching of chloroform to groundwater considered to be a significant pathway. Because of its volatility, evaporation is considered to be the main process for the removal of chloroform from aquatic systems. Chloroform is not expected to adsorb significantly to sediment or suspended organic matter in surface water.

Biodegradation: Hydrolysis or direct photolysis are not considered to be significant degradation processes in water for chloroform. Chloroform is generally considered persistent in water and soils with a low potential for degradation. Under correct condition, chloroform may undergo anaerobic biodegradation. Concentrations of chloroform in soil or water above a certain threshold levels results in toxic conditions which inhibits bacteria, methane-fermenting bacteria under anaerobic conditions.

Chloroform is unlikely to bioaccumulate to any significant extent in aquatic biota.

B3.4 Background Exposures/Intake With respect to chloroform the average intake from food, water and air has been estimated (WHO 2004) to be between 0.6 to 10 μg/kg/day (WHO 2004b). Data available from Australia indicate a similar range of potential intakes from water and air as those considered by the WHO. Given the available TDI levels adopted, it is considered relevant to assume a 50% intake from background. On this basis, the suggested threshold values should be adjusted to account for background intakes.


B3.5 Health Effects

General

The following information is available from WHO and ATSDR (ATSDR 1997b; WHO 2004b). There is no clinical disease which is unique to chloroform toxicity.

Chloroform is rapidly absorbed through the lungs and the gastrointestinal tract, and to some extent through the skin. In humans, the respiratory absorption of chloroform ranges from 49 to 77% and absorption from the gastrointestinal tract approximates 100%, with peak blood levels being reached within 1 hour (ATSDR 1997b; WHO 2004b).

Following its absorption, chloroform is distributed to all organs. The distribution of chloroform in the body does not differ qualitatively between the various routes of exposure. A number of studies have shown that chloroform distributes to fat tissue. It is lipid soluble, readily passes through cell membranes, reaching relatively high concentrations in nervous tissue. Chloroform concentrations in tissues are dose-related and occur in the following order: adipose > brain > liver > kidney > blood. Chloroform passes through the placenta and has been detected in fresh cow’s milk and foetal blood at levels equal to or greater than that in maternal blood (ATSDR 1997b; WHO 2004b).

Both oxidative and reductive pathways of chloroform metabolism have been identified. Chloroform is metabolised by oxidative dehydrochlorination of its carbon-hydrogen bond to form phosgene (CCl2O), while the reductive pathway generates the dichloromethylcarbene free radical. Both oxidative and reductive metabolism proceeds through a cytochrome P450 (CYP)-dependant enzymatic activation step that occurs in both the liver and the kidney. The balance between oxidative and reductive pathways depends on species, tissue, dose and oxygen tension. The major end product of chloroform metabolism is carbon dioxide (CO2), most of which is eliminated via the lungs, but some is incorporated into endogenous metabolites and may be excreted as bicarbonate, urea, methionine and other amino acids, inorganic chloride ion, and carbon monoxide. Elimination of chloroform is not affected by the route of exposure. About 60 - 70% is eliminated unchanged in expired air; 30 - 40% is metabolised and excreted in urine and faeces. The extent of metabolism is dose-dependent (ATSDR 1997b; WHO 2004b).

Target organs for chloroform toxicity are the liver, kidneys, and central nervous system. The most universally observed toxic effect of chloroform is damage to the liver. Liver effects (hepatomegaly, fatty liver, and hepatitis) were observed in individuals occupationally exposed to chloroform. Several subchronic and chronic studies by the oral or inhalation routes of exposure documented hepatotoxic effects in rats, mice, and dogs. Renal effects have been reported in rats and mice following oral and inhalation exposures, but evidence for chloroform-induced renal toxicity in humans is sparse (ATSDR 1997b; WHO 2004b).

Chloroform is a central nervous system depressant, inducing narcosis and anaesthesia at high concentrations. Lower concentrations may cause irritability, lassitude, depression, gastrointestinal symptoms, and frequent and burning urination (ATSDR 1997b; WHO 2004b).



Human data on the carcinogenic potential of chloroform are limited and there have been no conclusive associated between chloroform exposure and cancer in humans. In experiments with mice and rats, chloroform induced liver and kidney tumours. The carcinogenic effects of chloroform on the mouse liver appear to be closely related to cytotoxic and cell replicative effects. Liver tumours in rat and mice studies have only occurred where signs of hepatoxicity have been seen. In the rat and mice studies, the development of renal tumours in males is a consequence of nephrototoxicity of chloroform (ATSDR 1997b; WHO 2004b).

The pattern of development of tumours following chloroform treatment in animals is consistent with a tumour promoting mechanism rather than a genotoxic one. On the basis of available evidence, a dose threshold for the development of liver tumours is considered appropriate. It was considered plausible by the WHO that kidney tumours in rats may be associated with a threshold mechanism; however there are some limitations of the database (ATSDR 1997b; WHO 2004b).

Review of chloroform by the USEPA indicates that it is considered likely to be carcinogenic to humans by all routes of exposure under high-dose conditions that lead to cytotoxicity and regenerative hyperplasia. Chloroform is not likely to be carcinogenic to humans by any routes of exposure at doses that do not cause cytotoxicity and cell regeneration. Hence the USEPA has concluded that the threshold effects value established is also protective against increased risk of cancer (USEPA 2001).

Similarly a review of the mode of action by the WHO has identified that mechanism for induction of tumours is consistent with a non-linear (or threshold) dose-response relationship for induction of tumours (WHO 2004b).

The weight of the available evidence indicates that chloroform has little, if any, capability to induce gene mutation, chromosomal damage and DNA repair (WHO 1994a, 2004b). However, there is some evidence of low-level binding to DNA. Chloroform does not appear capable of inducing unscheduled DNA synthesis in vivo. Review of chloroform by USEPA indicates that chloroform is not a mutagen and is not likely to cause cancer through a genotoxic mode of action (USEPA 2001).


Based on the review conducted by ATSDR (ATSDR 1997b), the following is noted:

Since the liver and kidney are the two main organs responsible for chloroform metabolism, individuals who have hepatic or renal impairment may be more susceptible to chloroform toxicity; one such population would be those who abuse alcohol.

Also, exhaustion and starvation may potentiate chloroform hepatotoxicity, as indicated in some human clinical reports and in animal studies.

Animal studies indicate that male mice and rats may be more susceptible to the lethal and renal effects of chloroform than female mice and rats. The greater susceptibility of adult male animals is associated with testosterone levels in the animals.


Evidence also exists for age-related effects; young male mice were less susceptible to the lethal effects of chloroform compared to adult males.

Classification

Chloroform has been classified as a "probable" human carcinogen (Category B2) by the USEPA based on carcinogenicity in animals. The USEPA review considered that chloroform is likely to be carcinogenic to humans by all routes of exposure under high-exposure conditions that lead to cytotoxicity and regenerative hyperplasia in susceptible tissues. Chloroform is not likely to be carcinogenic to humans by any route of exposure under exposure conditions that do not cause cytotoxicity and cell regeneration (USEPA 2001). In addition, “the weight-of-evidence of the genotoxicity data on chloroform supports a conclusion that chloroform is not strongly mutagenic, and that genotoxicity is not likely to be the predominant mode of action underlying the carcinogenic potential of chloroform. Although no cancer data exist for exposures via the dermal pathway, the weight-of-evidence conclusion is considered to be applicable to this pathway as well, because chloroform absorbed through the skin and into the blood is expected to be metabolized and to cause toxicity in much the same way as chloroform absorbed by other exposure routes.”

IARC has classified chloroform in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans and sufficient evidence in experimental animals for carcinogenicity (IARC 1999c).

The National Occupational Health and Safety Commission (NOHSC) as Category 3 carcinogen (possible human carcinogen) (Safe Work Australia). NICNAS has not classified chloroform.

B3.6 Quantitative Toxicity Reference Values On the basis of the weight of evidence, it is appropriate that chloroform is assessed on the basis of a threshold. The following quantitative values are available for chloroform from relevant Australian and International sources:


Table B2 Summary of Published Toxicity Reference Values: Chloroform


TDI = 0.07 mg/kg/day The current ADWG have derived a drinking water guideline for total trihalomethanes, which included chloroform (as well as bromodichloromethane, dibromochloromethane and bromoform) of 0.25 mg/L as a total or individually using a TDI of 0.07 mg/kg/day derived from a no effect level based on a 90-day gavage study on rats and the application of 100 safety factor.

International WHO (WHO 1994a)

TDI = 0.015 mg/kg/day for non-neoplastic effects TDI = 0.01 mg/kg/day for neoplastic effects

TDI = 0.015 mg/kg/day based on non-neoplastic effects (hepatoxicity) in a 7.5 year study on dogs (lowest identified effects level of 15 mg/kg), 1000 uncertainty factor. TDI = 0.01 mg/kg/day for neoplastic effects (liver tumours) based on a 3 week study in mice (NOAEL of 10 mg/kg), 1000 uncertainty factor. Based on induction of renal tumours in male rats a total daily intake associated with a 10-5 excess cancer risk (linearised multistage model) is 0.0082 mg/kg/day. This review has been superseded by more recent reviews below.

WHO DWG (WHO 2011)

TDI = 0.015 mg/kg/day Based on hepatoxicity in dogs (derived in 1994 from oral studies) generated using PBPK modelling and an uncertainty factor of 25.

WHO (WHO 2004b)


TDI and TC derived on the same basis as noted above from the DWG. The TDI and TC were considered to be protective against possible human carcinogenic risk from chloroform exposure as well as possible kidney damage that likely precede tumour development.



TDI derived based on a LOAEL of 30 mg/kg/day for hepatotoxic effects in mice and an uncertainty factor of 1000. TC derived based on a NOAEL of 110 mg/m3 for hepatic effects in rats and an uncertainty factor of 1000.

ATSDR (ATSDR 1997b)

Oral MRL = 0.01 mg/kg/day Inhalation MRL =0.1 mg/m3

MRL derived based on a LOAEL of 12.9 (adusted) mg/kg/day for hepatotoxic effects in dogs and an uncertainty factor of 1000. The chronic inhalation MRL has been derived on the basis of a LOAEL of 9.9 mg/m3 associated with liver effects in humans and an uncertainty factor of 100.

USEPA (USEPA 2001)

RfD = 0.01 mg/kg/day

Oral reference dose (RfD) (updated in 2001) has been derived for the assessment of both carcinogenic and non-carcinogenic effects of chloroform. The RfD is based on a BMD approach based on a study associated with liver effects in dogs (the same study considered by the WHO (WHO 2004b)) and application of a 100 fold uncertainty factor. The value has been derived following consideration of the mode of action where a non-linear dose response was considered relevant. The assessment of cancer risk from chloroform inhalation is noted on IRIS to be dated (from 1987) and has not incorporated newer data or the revised cancer guidelines. The old approach of considering a linear dose response for carcinogens regardless of the mode of action remains on the data base, however this is under review.

The threshold values available from the WHO are the most current and relevant values that are considered adequately protective of human carcinogenic risk from chloroform exposure as well as possible kidney damage that likely precede tumour development (WHO 2004b).


B4 Dichloromethane B4.1 General Dichloromethane (also commonly known as methylene chloride as well as methane dichloride, methylene bichloride, methylene dichloride or DCM) is a synthetic compound, which is not known to occur naturally in the environment. DCM is primarily used as a solvent, especially for grease, plastics and various paint-binding agents. Among its uses are: as a component of paint and varnish strippers, and adhesive formulations; solvent in aerosol formulations; extractant in food and pharmaceutical industries; process solvent in cellulose ester production and fibre and film forming; process solvent in polycarbonate production; blowing agent in flexible polyurethane foams; the extraction of fats and paraffins; plastics processing, and metal and textile treatment; a vapour degreasing solvent in metal-working industries. The main use in consumer products is in paint strippers, where DCM is the main constituent (70-75%). The second important use is in hairspray aerosols, where it acts as a solvent and vapour pressure modifier. Other types of DCM-containing products are household cleaning products and lubricating, degreasing and automotive products, some of which may be in aerosol form. DCM is produced by the reaction of methanol with hydrogen chloride which is then reacted with chlorine. Chloroform and, to a lesser extent, carbon tetrachloride are also produced (ATSDR 2000).

B4.2 Properties DCM is a non-flammable, colourless liquid with a penetrating ether-like odour. It is soluble in alcohol, ether, acetone, chloroform and carbon tetrachloride. The pure dry compound is very stable. DCM hydrolyses slowly in the presence of moisture, producing small quantities of hydrogen chloride. Commercial DCM normally contains small quantities of stabilisers to prevent decomposition. Key properties are presented below (ATSDR 2000; RAIS):

CAS No 75-09-2 Chemical Formula CH2Cl2 Molecular Weight 84.93 Vapour Pressure 435 mmHg at 20oC Vapour Density 2.9 Density 1.32 g/ml at 25oC Solubility 13000 mg/L at 20oC Air Diffusion Coefficient 0.099 cm2/s Water Diffusion Coefficient 1.25 x 10-5 cm2/s Henry’s Law Coefficient 0.00219 atm.m3/mol

= 0.133 at 25oC (unitless) Koc 21.7 cm3/g Log Kow 1.25 Odour Threshold 540 to 2160 mg/m3 Permeability Constant 0.00354 cm/hr


B4.3 Exposure Human exposure to DCM occurs principally through inhalation. However exposure may also occur via oral and dermal routes particularly during occupational or consumer use of DCM containing products. The chlorination of drinking water also produces DCM. NHMRC indicate that DCM has not been found in Australian drinking waters (NHMRC 2011 Updated 2016).

In relation to the assessment of dermal absorption, dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance (USEPA 1995a) of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). No data is available specifically for the dermal absorption of DCM and it has a vapour pressure greater than benzene; hence a value of 0.05% is considered relevant.

If released into the environment the following can be noted with respect to DCM (ATSDR 2000; WHO 1996):

Air: Nearly all DCM released to the environment will ultimately be present in the atmosphere due to its volatility, where it will degrade by reaction with photochemically produced hydroxyl radicals with a lifetime of 6 months. Transport can occur to regions far removed from the emission source. DCM is expected to have no significant impact on stratospheric ozone depletion nor will it contribute significantly to photochemical smog formation.

Soil and Water: Following releases to soil, most DCM is expected to volatilise and low soil adsorption. Most of the remaining DCM will travel through the soil because of its low adsorption onto soils (and hence high mobility) with leaching to groundwater considered to be a significant pathway. Volatilisation is considered to be the main process for the removal of DCM from aquatic systems. DCM is not expected to adsorb significantly to sediment or suspended organic matter in surface water.

Biodegradation: DCM undergoes slow hydrolysis in water and hence it is not considered to be a significant degradation process in water. Both aerobic and anaerobic biodegradation may be important for DCM in water. Degradation of DCM in soils was found to occur with the rate of degradation dependant on the soil type, concentration and redox state of the soil with degradation observed under both aerobic and anaerobic conditions. Biodegradation of DCM appears to be accelerated by the presence of elevated levels of organic carbon.

Bioaccumulation of DCM is not expected to be significant.

B4.4 Background Exposures/Intake With respect to DCM, intakes from soil, water and food can be considered to be insignificant. Based on data available from urban air in Brisbane and Perth (Hawas et al. 2001; WA DEP 2000) DCM intakes from air may contribute to approximately 20% of the TDI. Hence, the suggested TDI values presented for the evaluation of DCM should be adjusted to account for 20% intake from background.


B4.5 Health Effects

General

The following information is available from WHO and ATSDR (ATSDR 2000; WHO 1996). There is no clinical disease which is unique to DCM toxicity.

Humans and animals readily absorb DCM from the lungs and the gastrointestinal tract into systemic circulation. The compound is also absorbed to some extent through intact skin. Following absorption, DCM concentrations rapidly increase in the blood to reach equilibrium levels that depend primarily on exposure concentrations. A fairly uniform distribution to heart, liver, and brain is reported with increased concentrations also reported in adipose tissue. DCM is quite rapidly excreted, mostly via the lungs in the exhaled air. It can cross the blood-brain barrier and be transferred across the placenta, and small amounts can be excreted in urine or in milk (ATSDR 2000; WHO 1996).

Extensive toxicokinetic studies have shown that DCM is metabolised by two pathways: (1) a mixed function oxidase (MFO) pathway mediated by the P-450 system yielding CO and CO2 and (2) a glutathione-dependent (GST) pathway yielding only CO2. Other metabolites of DCM include formaldehyde and formic acid (ATSDR 2000; WHO 1996).

Tests involving acute exposure of animals have shown DCM to have moderate acute toxicity from oral and inhalation exposure. Case studies of DCM poisoning during paint stripping operations have demonstrated that inhalation exposure to extremely high levels can be fatal to humans. Acute inhalation exposure to high levels of DCM in humans has resulted in effects on the central nervous system (CNS) including decreased visual, auditory, and psychomotor functions, but these effects are reversible once exposure ceases. DCM also irritates the nose and throat at high concentrations (ATSDR 2000; WHO 1996).

The major effects from chronic inhalation exposure to DCM in humans are effects on the CNS, such as headaches, dizziness, nausea, and memory loss. Animal studies indicate that the inhalation of DCM causes effects on the liver, kidney, CNS, and cardiovascular system (ATSDR 2000; WHO 1996).

Animal studies have demonstrated that DCM crosses the placental barrier, however in the studies available DCM is not a reproductive toxicant nor is it a developmental toxicant via inhalation or oral pathways (ATSDR 2000; WHO 1996).


Review of carcinogenicity indicates that DCM has been found to be carcinogenic in mice, causing both lung and liver tumours, following exposure to high concentrations in air. These tumours were not seen in the rat or the hamster (WHO 1996).

Metabolism and biochemical studies, and mutagenicity assays in bacteria and B6C3F1 mice have provided a plausible explanation for the mechanism of action and the species differences in the carcinogenicity of DCM to the lung and liver. This explanation is based on the existence of an isoenzyme of glutathione-S-transferase which specifically metabolises DCM to the reactive intermediates responsible for tumour induction in the mouse. This is an important pathway only in


mice and then only at high doses. It is not a major pathway for rats, hamsters or humans. The mouse appears to be unique in its response to DCM and is not an appropriate model for humans. It is noted that the USEPA review has not dismissed these studies and considers them in their evaluation. Evaluation of the relevance of the mice studies notes their limitations but does not preclude them (USEPA 2011c).

Benign mammary tumours observed in rats exposed at high doses to DCM are the result of high serum prolactin levels which is not expected to occur at low levels of exposure and has not been observed in humans exposed to DCM (USEPA 2011c; WHO 1996).

On the basis of the available information, the carcinogenic potency of DCM in humans is expected to be low. This evaluation is considered to remain current despite the USEPA review which considers that there is sufficient data (primarily from the mice studies) to consider that DCM is “likely to be carcinogenic in humans” (USEPA 2011c; WHO 1996).

Review of genotoxicity indicates that the available data indicate that DCM or its metabolites are capable of interacting with DNA (WHO 1996). However, with the exception of mouse studies, in vivo studies using high levels of DCM exposure have not provided clear evidence of genotoxicity. The evidence suggests that DCM genotoxicity in the mouse results from the metabolism of DCM to genotoxic metabolites and this is a species-specific phenomenon which does not appear to occur in other species including humans. Therefore, the relevance of the genotoxicity in mice to humans is considered limited and there is no conclusive evidence that DCM in genotoxic (USEPA 2011c; WHO 1996).

The USEPA review has considered that the mode of action for DCM induced tumours in the lung and liver in mice) involves mutagenicity via reactive metabolites. As these endpoints are not considered relevant to humans, and no additional data is presented in the review, there is no strong evidence that DCM is genotoxic or mutagenic. However, if further data/peer-reviewed evaluations are available that supports this outcome then the assessment of DCM will need to be revised (USEPA 2011c; WHO 1996).


Insufficient information is available regarding potentially susceptible populations. The review conducted by the USEPA assumes a mutagenic mode of action base on mice data which does not appear to be supported by other reviews (USEPA 2011c; WHO 1996).

Classification

DCM has been classified as a "probable" human carcinogen (Category B2) by the USEPA based on increased incidence of hepatocellular neoplasms and alveolar/bronchiolar neoplasms in mice and benign mammary tumours in rats. It is noted that the USEPA review considered that DCM is “likely to be carcinogenic to humans” by all routes of exposure (USEPA 2011c).

IARC has classified DCM in Group 2A (probably carcinogenic to humans) based on inadequate evidence in humans and sufficient evidence in experimental animals for carcinogenicity (Benbrahim-Tallaa et al. 2014).


The National Occupational Health and Safety Commission (NOHSC) has classified DCM as Category 3 carcinogen (possible human carcinogen) (Safe Work Australia). NICNAS has not classified DCM

B4.6 Quantitative Toxicity Reference Values On the basis of the weight of evidence, it is appropriate that DCM is assessed on the basis of a threshold. The following quantitative values are available for DCM from relevant Australian and International sources:

Table B3 Summary of Published Toxicity Reference Values: Dichloromethane


TDI = 0.0012 mg/kg/day The current ADWG have derived a drinking water guideline for DCM of 0.004 mg/L using a TDI of 0.0012 mg/kg/day derived from a lowest effect level based on a 2 year drinking water study on rats and the application of 5000 safety factor. The 5000 fold safety factor included a 10 fold factor for potential genotoxicity.

International WHO DWG (WHO 2011)

TDI = 0.006 mg/kg/day Based on hepatotoxic effects in a 2-year drinking water study in rats and an uncertainty factor of 1000 including a factor of 10 to address concern about the carcinogenic potential.

WHO (WHO 2000c)

TC = 0.45 mg/m3 (1 week) The WHO (2000) review of DCM has identified that carcinogenicity is not the critical end point for risk assessment purposes. The formation of carbon monoxide in blood (COHb is a more direct indication of a toxic effect, it can be monitored and is a more suitable basis for the derivation of a guideline. A guideline value of 3 mg/m3 has been derived for the assessment of 24-hour exposures based on a 0.1% increase in COHb. In addition, it is noted that the weekly average concentration should not exceed one-seventh of this guideline (0.45 mg/m3).

EU (CSTEE 2000; TNO 1999)

TC = 1.25 mg/m3

Chronic inhalation exposures have been evaluated on the basis of liver effects in the rat, leading to a NOAEL of 125 mg/m3 continuous exposure, which, with a MOS of 100, reduces to a limit of 1.25 mg/m3. A chronic standard of 0.2 mg/m3 has been derived based on 1x10-4 incremental risk level, however the acceptable risk level adopted was is not considered to be acceptable by the EU. The review committee considered that the acceptable limit for long-term exposure of the general population should be based on consideration of the carcinogenic risk.


TDI = 0.06 mg/kg/day TC = 3 mg/m3

TDI derived based on a NOAEL of 6 mg/kg/day for hepatotoxic effects in rats and an uncertainty factor of 100. TC derived based on a LOAEL (adjusted) of 28 mg/m3 for CNS effects and increases in blood COHb levels in humans and an uncertainty factor of 10. Same as that derived by WHO for 24-hour exposures.

TCEQ (TCEQ 2011)

Chronic ESL = 0.35 mg/m3

Development of air criteria resulted in the derivation of a threshold criteria of 1.3 mg/m3 associated with hepatotoxic effects from a 2 year inhalation study in rats, and an uncertainty factor of 100. A lower criteria of 0.35 mg/m3 was derived on the basis of linear effects for liver and lung tumours in mice and a risk level of 1x10-5. The lower value from these evaluations listed in this table.

ATSDR (ATSDR 2000)

Oral MRL = 0.06 mg/kg/day Inhalation MRL =1.1 mg/m3

MRL derived based on a NOAEL of 6 mg/kg/day for hepatotoxic effects in rats and an uncertainty factor of 100. The chronic inhalation MRL has been derived on the basis of a NOAEL (adjusted) of 31 mg/m3 associated with liver effects in rats and an uncertainty factor of 30.

USEPA (USEPA 2011c)


Oral RfD based on a BMD from study associated with hepatotoxic effects in a rat study, use of a rat physiologically based PBPK model to estimate internal doses and an uncertainty factor of 30.


Source Value Basis/Comments Inhalation RfC based on a BMD from an inhalation study associated with hepatotoxic effects in a rat study, use of a rat physiologically based PBPK model to estimate internal doses, scaling factor to estimate human doses and an uncertainty factor of 100. The USEPA has also developed non-threshold toxicity reference values (not relevant for the assessment of threshold effects).

The above table presents a range of values for DCM, based on different studies but generally consistent health endpoints. In the evaluations available there remains some uncertainty in the potential for carcinogenic effects, hence, the oral value from the current ADWG (NHMRC 2011 Updated 2016) and the lower inhalation value available from TCEQ (TCEQ 2011), which is consistent with the inhalation value derived by the WHO (WHO 2000c) and USEPA (USEPA 2011c), have been adopted. It is noted that the use of these values provides a more conservative assessment than the use of non-threshold values from the USEPA review (USEPA 2011c). Hence these values are considered to be adequately protective of potential carcinogenic effects.


B5 1,1,2,2-Tetrachloroethane B5.1 General Tetrachloroethane (TCA) is a chemical compound occurring in two isomers: 1,1,1,2- and 1,1,2,2-Tetrachloroethane (also known as acetylene tetrachloride; sym-tetrachloroethane; s-tetrachloroethane; tetrachloroethane, 1,1-dichloro-2,2-dichloroethane and commonly abbreviated to 1,1,2,2-TCA) is a synthetic chemical not known to occur naturally. 1,1,2,2-TCA is widely used as an industrial solvent. It is also used as an intermediate in the production of trichloroethylene and tetrachloroethylene. Other minor uses of 1,1,2,2-TCA include the use as an insecticide (moth-proofing and as a fumigant). 1,1,2,2-TCA is synthesised by direct chlorination or oxychlorination of ethylene. 1,1,2,2-TCA can be a by-product generated during the production of vinyl chloride or ethylene dichloride (ATSDR 2008).

B5.2 Properties At room temperature both isomers of TCA are non-flammable, colourless, low to moderate volatile liquids with a sweetish, suffocating chloroform-like odour. It is miscible with ethanol, methanol, ether, acetone, benzene, petroleum, carbon tetrachloride, chloroform, carbon disulfide, dimethyl formamide and oils. Key properties are presented below (ATSDR 2008; RAIS):

CAS No. 79-34-5 Chemical Formula C2H2Cl4 Molecular Weight 167.85 Vapour Pressure 13.3 mmHg at 21oC Vapour Density 5.32 Density 1.59 g/ml at 20oC Solubility 2830 mg/L at 20oC Air Diffusion Coefficient 0.0489 cm2/s Water Diffusion Coefficient 9.29 x 10-6 cm2/s Henry’s Law Coefficient 0.00036 atm.m3/mol

= 0.015 at 25oC (unitless) Koc 94.9 cm3/g Log Kow 2.39 Odour Threshold 21 to 35 mg/m3 Permeability Constant 0.00694 cm/hr

B5.3 Exposure Exposure of the general population to both isomers of TCA may occur primarily through inhalation; however, exposure via oral or dermal routes may occur but are expected to be insignificant.

In relation to the assessment of dermal absorption, dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance (USEPA 1995a) for dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%, which includes compounds such as TCA. No data is available specifically for the dermal absorption of TCA; hence a default value of 3% is considered appropriate.


If released into the environment the following can be noted with respect to TCA (ATSDR 2008):

Air: Most of the TCA released to the environment enters the atmosphere where it is fairly stable in the lower atmosphere. The dominant process for removal of TCA is reaction with photochemically generated hydroxyl radicals with an estimated half-life of 53 days for 1,1,2,2-TCA and 550 days for 1,1,1,2-TCA. Removal may also occur through washout by precipitation, however most TCA is expected to re-enter the atmosphere via volatilisation. Atmospheric degradation of TCA is slow enough for it to be transported long distances with slow diffusion into the stratosphere expected to occur. TCA is not expected to contribute to the depletion of stratospheric ozone or to global warming.

Soil and Water: Following releases to soil, some TCA is expected to volatilise, with the remainder leaching into the subsurface soil profile and possibly to groundwater. If released to surface water, part would volatilise with the remainder dissolving in water where it would undergo degradation through hydrolysis. In groundwater, the major degradation processes involve anaerobic degradation and chemical hydrolysis.

Biodegradation: TCA may undergo degradation through hydrolysis and/or anaerobic degradation. Chemical hydrolysis is very sensitive to pH and is much more rapid under basic or neutral conditions. Trichloroethene is the major (if not only) product of chemical hydrolysis with half-lives reported in neutral pH from 29 to 102 days for 1,1,2,2-TCA. Biodegradation occurs through dehydrodehalogenation with the products of biodegradation including trichloroethylene, 1,2-dichloroethene and vinyl chloride.

Since TCA is volatile and can be expected to be transformed to other compounds such as TCE, TCA would not be expected to accumulate in sediments. TCA has a low tendency to bioconcentrate in aquatic or marine organisms.

B5.4 Background Exposures/Intake With respect to TCA, intakes from soil, water and food can be considered to be insignificant. Based on data available from urban air in Brisbane and Perth (Hawas et al. 2001; WA DEP 2000), TCA is generally not detected or rarely detected. Hence background intake of TCA can be considered to be negligible.

B5.5 Health Effects

General

The following information is available from WHO and ATSDR (ATSDR 2008; WHO 1998b). There is no clinical disease which is unique to TCA toxicity. Owing to the significant decline in the use of this substance, the toxicological profile of TCA has not been well characterised, with the available data being confined primarily to early limited studies.

TCA is well absorbed from the gastrointestinal and respiratory tract in animals and humans. Dermal absorption of TCA occurs. No studies are available on the distribution of TCA via inhalation, oral or dermal routes of exposure (ATSDR 2008; WHO 1998b).


The metabolism of 1,1,1,2-TCA proceeds through both oxidative and reductive pathways resulting in CO2 in exhaled air and trichloroethanol and trichloroacetic acid in urine as main metabolites. The data shows that there is a difference in the metabolism of 1,1,1,2- and 1,1,2,2-TCA. The amount of trichloro compounds in the urine of rats exposed to 1,1,1,2-TCA was 20 times higher than the amount of trichloro compounds in the urine of rats similarly exposed to the 1,1,2,2-isomer (Health Council of the Netherlands 2006).

Based on data on the metabolism of 1,1,2,2-TCA in mice, it is suggested that the principal pathway of degradation involves stage-wise hydrolytic cleavage of the carbon-chlorine bonds and oxidation to dichloroacetaldehyde hydrate, dichloroacetic acid (the major metabolite), and eventually glyoxylic acid. The glyoxylic acid is then metabolized to oxalic acid, glycine, formic acid, and carbon dioxide. A small proportion of the parent compound is probably non-enzymatically dehydrochlorinated to trichloroethene, which is further converted to trichloroacetic acid and trichloroethanol. In addition to the liver, metabolism may also occur in the epithelia of the respiratory tract and upper alimentary tract (WHO 1998b).

The metabolites of TCA are eliminated in the urine, faeces, skin and expired air (ATSDR 2008; WHO 1998b).

Based on the results of studies in experimental animals, the acute toxicity of TCA is slight to moderate. The chemical may induce skin, eye, and mucosal irritation (ATSDR 2008; WHO 1998b).

Limited long-term human data are not available; hence the evaluation of critical effects is based on animal studies. The results of available studies on the non-neoplastic effects of TCA in experimental animals exposed by ingestion or inhalation indicate that the central nervous system and liver are the principal target organs (ATSDR 2008; WHO 1998b).


Human data on the carcinogenic potential of TCA are limited; hence data available from animal studies have been used by various agencies in reviewing carcinogenicity.

Long-term exposure to 1,1,2,2-TCA resulted in a significantly increased incidence of hepatocellular carcinomas in both male and female mice. However, no significant increases in tumours were observed in similarly exposed rats. The relevance of increased incidence of liver tumours in mice to humans in the absence of a carcinogenic response in other species or other mouse tissues and without clear evidence of genotoxicity is highly questionable (ATSDR 2008; WHO 1998b).

The WHO has proposed “that the liver tumours in mice may be induced by mechanisms that may not be relevant to humans, for which humans are less susceptible, or for which there may be a threshold of exposure. In addition, it has been hypothesized that the carcinogenicity of 1,1,2,2-TCA may be associated with the formation of free radicals, lipid peroxidation, or hepatic damage (such as focal necrosis associated with intense cellular proliferation). Therefore, on the basis of data currently available, it is not possible to draw any firm conclusions with respect to the potential carcinogenicity of 1,1,2,2-TCA in humans” (ATSDR 2008; WHO 1998b).


There is limited evidence from studies in rats and mice that 1,1,2,2-TCA is carcinogenic. Increased incidence of fibroadenomas in female rats and hepatocellular adenomas in mice were observed in some studies (ATSDR 2008; WHO 1998b).

The weight of evidence of available in vitro and in vivo assays suggests that 1,1,2,2-TCA is not genotoxic or that it is, at most, weakly genotoxic (ATSDR 2008; WHO 1998b).


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. In relation to 1,1,2,2-TCA the following is noted from ATSDR (ATSDR 2008):

No direct evidence is available that suggests that children are more susceptible than adults to 1,1,2,2-TCA toxicity. In general, children may be more vulnerable to 1,1,2,2-TCA since intake dose per kilogram of body weight may be greater in early life than in mature humans, because children eat more food, drink more water, breathe more air, and ingest more soil/house dust per kilogram body weight than older age groups.

As metabolism is believed to play an important role in the toxicity of 1,1,2,2-TCA, particularly in the liver, individuals with elevated levels of cytochrome P450 enzymes may have an increased susceptibility to the compound. Since 1,1,2,2-TCA has been demonstrated to inhibit cytochrome P450 enzymes, it is also possible that people co-exposed to chemicals that are inactivated by cytochrome P450 enzymes will be more susceptible to those compounds.

Because the liver and nervous system are the main targets of 1,1,2,2-TCA toxicity, individuals with compromised function of liver or nervous system may be at increased risk from exposure to 1,1,2,2-TCA.

Classification

TCA (both isomers) has been classified as a "possible" human carcinogen (Category C) by the USEPA based on increased incidence of hepatocellular carcinomas in mice (USEPA 2008, 2009).

IARC has classified TCA (both isomers) in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans and sufficient evidence in experimental animals for carcinogenicity (IARC 2014).

The National Occupational Health and Safety Commission (NOHSC) or NICNAS have not classified the potential carcinogenicity of TCA (Safe Work Australia).

B5.6 Quantitative Toxicity Values On the basis of the weight of evidence, it is appropriate that 1,1,2,2-TCA is assessed on the basis of a threshold. The following quantitative values are available for 1,1,2,2-TCA from relevant Australian and International sources:


Table B4 Summary of Published Toxicity Reference Values: 1,1,2,2-TCA

Source Value Basis/Comments Australian – No guideline available International WHO (WHO 1998b)

Oral = 0.0012 – 0.0056 mg/kg/day Inhalation = 0.0034 – 0.016 mg/m3

Review of 1,1,2,2-TCA by the WHO was undertaken in 1998. Based on the limitations associated with the available studies a TDI cannot be derived with confidence for non-neoplastic effects or neoplastic effects. The toxicological end-point for which the dose-response relationship is best characterised is the increase in hepatocellular carcinomas observed in the long-term bioassay in mice, however as noted above, the relevant to humans is questionable. A range of guidance values have been derived using multistage modelling associated with the doses associated with a 5% increase in tumour incidence (TD0.05) and applying a safety margin. Guidance values derived are:

o Inhalation: 3.4 to 16 μg/m3 (applying 5000 margin) or 0.34 to 1.6 μg/m3 (applying 50000 margin) – derived from oral study. These values correspond to those considered by some agencies to represent "essentially negligible" risk (i.e. 10-5 to 10-6). This is associated with a slope factor that ranges from 0.0018 to 0.0083 (mg/kg/day)-1.

o Ingestion: 1.2 to 5.6 μg/kg/day (applying 5000 margin) or 0.12 to 0.56 μg/kg/day (applying 50000 margin) These values correspond to those considered by some agencies to represent "essentially negligible" risk (i.e. 10-5 to 10-6). This is equivalent to an inhalation unit risk of (0.6 to 3)x10-6 (μg/m3)-1

WHO (WHO 2000a)

NA The WHO (2000b) have derivation of an inhalation unit risk of (0.6 to 3)x10-6 (μg/m3)-1 for 1,1,2,2-TCA on the basis of the guidance values derived in the earlier review (WHO 1998b). It is noted that this evaluation is considered conservative as there is suggestive, but incomplete evidence that 1,1,2,2-TCA may induce tumours through a threshold mechanism. In addition, this value has been derived on the basis of an oral exposure study.

Health Canada (Health Canada 1993c)

NA Insufficient data was available to derive a TDI

ATSDR (ATSDR 2008)

NA The 2008 revision only included an intermediate oral MRL of 0.5 mg/kg/day based on liver effects in female rats using a BMD and a 100 fold uncertainty factor. The previous ATSDR review (1996) included a chronic oral MRL of 0.04 mg/kg/day.

USEPA (USEPA 2010c)


Oral RfD based on a BMD of 15 mg/kg/day associated with changes in lover weight in a sub-chronic rat study, with application of 1000 fold uncertainty factor. No inhalation criteria are available. The USEPA evaluation also provides non-threshold TRVs for oral exposures

The above table highlights the lack of quantitative data that is relevant to the assessment of 1,1,2,2-TCA as a threshold. The guidance values (upper values) developed by the WHO have been adopted in this assessment as these are considered to be adequately protective of all relevant health endpoints (WHO 1998b). While they have been derived on the basis of a non-threshold approach the values adopted for a 1 in 100,000 risk level are lower than the available threshold values based on other (non-cancer) health endpoints.


B6 1,1,2-Trichloroethane B6.1 General 1,1,2-Trichloroethane (also known as ethane trichloride; 1,1,2-TCE; beta-trichloroethane; 1,2,2-trichloroethane; vinyl trichloride; trichloroethane (non-specific name) and commonly abbreviated to 1,1,2-TCA) is a predominantly man-made chemical. 1,1,2-TCA is a chemical intermediate in the production of 1,1-dichloroethene. 1,1,2-TCA has limited use as a solvent for fats, oils, waxes and resins. It is also released to the environment as a result of anthropogenic activity and it has also been identified as an intermediate in the biodegradation of 1,1,2,2-tetrachloroethane (another man-made chemical). It is formed commercially by the chlorination of ethylene with chlorine or by the oxychlorination of ethylene with HCl and oxygen (ATSDR 1989).

B6.2 Properties 1,1,2-TCA is a non-flammable, colourless, volatile liquid with a pleasant, sweet odour. It is insoluble in water and miscible with alcohol, ether and many organic liquids. Key properties are presented below (ATSDR 1989; RAIS):

CAS No 79-00-5 Chemical Formula C2H3Cl3 Molecular Weight 133.41 Vapour Pressure 23 mmHg at 25oC Vapour Density 4.63 Density 1.4 g/ml at 20oC Solubility 4590 mg/L at 20oC Air Diffusion Coefficient 0.0669 cm2/s Water Diffusion Coefficient 1.0 x 10-5 cm2/s Henry’s Law Coefficient 0.0008 atm.m3/mol

= 0.0337 at 25oC (unitless) Koc 60.7 cm3/g Log Kow 1.89 Odour Threshold 2.8 - 926.8 mg/m3 Permeability Constant 0.00504 cm/hr

B6.3 Exposure Exposure of the general population to 1,1,2-TCA may occur primarily through inhalation, however exposure via oral or dermal routes may occur but are expected to be insignificant. Exposure may occur in the workplace where it is used as a solvent.

In relation to the assessment of dermal absorption, as noted in USEPA (USEPA 2004), dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant. Region III guidance (USEPA 1995a) of dermal absorption for volatile compounds (http://www.epa.gov/reg3hscd/risk/human/info/solabsg2.htm) suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%, which includes compounds such as 1,1,2-TCA. No data is available specifically for the dermal absorption of 1,1,2-TCA; hence a default value of 3% is considered appropriate.


If released into the environment the following can be noted with respect to 1,1,2-TCA (ATSDR 1989):

Air: Most of the 1,1,2-TCA released to the environment enters the atmosphere where it is fairly stable. In the atmosphere,1,1,2-TCA is degraded by photochemically-produced hydroxyl radicals with a half-life of approximately 49 days.

Soil and Water: Following releases to soil, 1,1,2-TCA is expected to partially volatilise, with the remainder leaching into the subsurface soil profile and groundwater. If released to surface water most would volatilise with the remainder dissolving in water. The chemical would not be expected to show appreciable adsorption to sediment or suspended organic material.

Biodegradation: 1,1,2-TCA may undergo slow biodegradation under anaerobic. Anaerobic degradation occurs predominantly through reductive dehalogenation which forms vinyl chloride. Aerobic degradation occurs via substitutive and oxidative mechanisms with the production of trichloroethyl alcohol. Aerobic degradation and hydrolysis are not likely to be an important fate processes for 1,1,2-TCA.

1,1,2-TCA has a low tendency to bioconcentrate in aquatic or marine organisms.

B6.4 Background Exposures/Intake With respect to 1,1,2-TCA, intakes from soil, water and food can be considered to be insignificant. Based on data available from urban air in Brisbane and Perth (Hawas et al. 2001; WA DEP 2000) 1,1,2-TCA is generally not detected in urban air and hence background intake can be considered to be negligible.

B6.5 Health Effects

General

There is no clinical disease which is unique to 1,1,2-TCA toxicity. 1,1,2-TCA is rapidly and extensively absorbed into the body following inhalation exposures (principal route of exposure) and dermal exposure (ATSDR 1989; US DOE 1995).

One absorbed, 1,1,2-TCA is distributed widely in body tissues (including the liver, fatty tissue, kidneys, blood and brains, heart, spleen and lungs). The primary metabolites identified are chloroacetic acid, S-carboxymethylcysteine, and thiodiacetic acid. Elimination occurs via exhalation and urine (including elimination of metabolites) (ATSDR 1989; US DOE 1995).

No information is available on the acute effects of 1,1,2-TCA in humans from inhalation or oral exposures. Tests involving acute exposure of mice and rats have shown 1,1,2-TCA to have moderate and high acute toxicity from inhalation and oral exposures, respectively. Studies on dermal exposure to 1,1,2-TCA in humans have reported stinging and burning sensations and transient whitening of the skin. Animal studies have reported effects on the liver, kidney, and central nervous system (CNS) from acute inhalation and oral exposure (ATSDR 1989; US DOE 1995).


No information is available on the chronic effects of 1,1,2-TCA in humans from inhalation or oral exposure. Animal studies have not observed adverse effects from chronic inhalation exposure to 1,1,2-TCA, however effects on the liver and immune system have been noted in chronic oral studies (ATSDR 1989; US DOE 1995).

No information is available regarding developmental or reproductive effects of 1,1,2-TCA in humans from inhalation or oral exposure. Animal studies have not reported developmental or reproductive effects from oral exposure to 1,1,2-TCA (ATSDR 1989; US DOE 1995).


No studies are available regarding cancer in humans from inhalation or oral exposure. A study reported liver tumours and adrenal tumours in mice, but no tumours in rats from exposure to 1,1,2-TCA by gavage. Initiation/promotion screening studies on male rat liver demonstrated that the chemicals has neither initiation nor promotion activity. A carcinogenic study in skin of rats indicted no chemical related changes (ATSDR 1989; US DOE 1995).

Potential for genotoxicity of 1,1,2-TCA was reviewed in the Stage 2 Assessment (Woodward-Clyde 1996) which indicated that the available data were inadequate to enable a proper evaluation of genotoxic potential. In particular, there were a lack of gene mutation assays using mammalian cells and in vivo chromosome damage assays. Review of genotoxicity by OECD (OECD 2000) recommended further work such as an in vivo genotoxicity study. This was undertaken and reported in 2003 (OECD 2003) which showed negative results. Review of available studies by ATSDR (ATSDR 2010) did not indicate the compound was mutagenic. Hence the weight of evidence suggests that 1,1,2-TCA is not genotoxic in vivo.


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. Insufficient information is available regarding 1,1,2-TCA, hence no specific susceptible populations can be identified.

Classification

1,1,2-TCA has been classified as a “possible” human carcinogen (Category C) by the USEPA on the basis of hepatocellular carcinomas and pheochromcytomas in one strain of mice.

IARC has classified 1,1,2-TCA in Group 3 (not classifiable as to its carcinogenicity to humans) based on no epidemiological data and limited evidence in experimental animals for carcinogenicity (IARC 1999a).

The National Occupational Health and Safety Commission (NOHSC) and NICNAS have not classified the potential carcinogenicity of 1,1,2-TCA.


B6.6 Quantitative Toxicity Values On the basis of the weight of evidence, it is appropriate that 1,1,2-TCA is assessed on the basis of a threshold. On this basis, the following quantitative values are available for 1,1,2-TCA from relevant Australian and International sources:

Table B5 Summary of Published Toxicity Reference Values: 1,2,2-TCA

Source Value Basis/Comments Australian – No guideline available International ATSDR (ATSDR 1989)

NA Only included an intermediate oral MRL of 0.04 mg/kg/day based on liver effects and an oral acute MRL of 0.3 mg/kg/day based on neurological effects.

USEPA (USEPA)


RfD (last updated in 1988) based on a NOAEL of 3.9 mg/kg/day associated with clinical serum chemistry changes in mice and an uncertainty factor of 1000. The USEPA also presents an oral slope factor of 0.057 (mg/kg/day)-1 based on a linear multistage model based on hepatocellular carcinomas; and an inhalation unit risk of 1.6x10-5 ( g/m3)-1 using a linear multistage model based on oral data used to derive the oral slope factor.

Only one threshold value is available, from the USEPA, which has been used in the assessment of oral and dermal exposures. No inhalation data, or evaluations, are available. It has therefore been assumed that the oral toxicity reference value is relevant for all pathways of exposure, with an inhalation value of 0.014 mg/m3 derived on the basis of a 70 kg body weight and inhalation of 20 m3/day.


B7 1,2-Dichloroethane (EDC) B7.1 General 1,2-Dichloroethane (also known as ethylene dichloride, ethylene chloride, glycol dichloride, freon 150, dutch liquid, 1,2-ethylene dichloride, alpha, beta-dichloride and commonly abbreviated to EDC) is a synthetic product which is primarily used in the production of the vinyl chloride monomer. It is also an intermediate in the manufacture of fluorocarbons and chlorinated solvents such as trichloethane, trichloroethylene, perchloroethylene and vinylidene. These solvents are used to remove dirt, grease, resins and glue as well as in the manufacture of polystyrene and latex. EDC was also added to leaded petrol as an anti-knock compound and has been used as a fumigant.

EDC is a volatile, colourless liquid at room temperature with a pleasant smell and sweet taste. EDC evaporates into air very quickly and is soluble in water and several organic solvents such as alcohols, chloroform and ether.

EDC is one on the most widely produced chemicals in the world. The majority of EDC released to the environment is in emissions to air. It is moderately persistent in the air, however it is not considered to be an ozone depleting substance.

B7.2 Exposure Exposure of the general population to EDC may by inhalation, oral or dermal routes. In most cases inhalation is the primary route of exposure. Exposure may also occur through oral ingestion and dermal contact with drinking/household water and/or soils. Children may exposed via the same pathways as for adults. EDC has been detected in human milk and hence infants could be exposed via breast-feeding. Intake from food sources is expected to be negligible. Occupational exposures (particularly inhalation and dermal contact) may occur in industries which handle the product (ATSDR 2001; HSDB).

In relation to the assessment of dermal absorption, as noted in USEPA (USEPA 2004), dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant. Region III guidance (USEPA 1995b) of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg). For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%, which includes compounds such as EDC. No data is available specifically for the dermal absorption of EDC; hence a default value of 3% is considered appropriate.

If released into the environment the following can be noted with respect to EDC (HSDB):

Air: EDC is expected to remain in vapour phase where it is moderately persistent with an estimated half-life of between 43 and 111 days. Once EDC reaches the troposphere, it undergoes photo-oxidation to produce formyl chloride, chloroacetyl chloride, hydrochloric acid, carbon monoxide and carbon dioxide EDC is transported to the stratosphere where photolysis may produce chloride radicals which may in turn reach the ozone layer. EDC is not expected to contribute to ozone depletion. Due to its persistence in the troposphere there is the potential for long-range transport of EDC.


Soil and Water: EDC is not expected to adsorb strongly in soils and may leach to groundwater where it has the potential to persist for years. EDC is expected to volatilise from surface soils and water.

Biodegradation: Biodegradation is expected to occur slowly with hydrolysis and photolysis not expected to be important fate processes. The potential for bioaccumulation in aquatic or terrestrial organisms appears to be low.

B7.3 Background Exposures/Intake For common contaminants, intakes from background sources such as food, water and/or air must also be considered in the evaluation and use of the ADI, TDI or RfD in assessing potential exposures to site related chemicals. However, as EDC has been evaluated to be a genotoxic carcinogen it is considered appropriate to evaluate exposure using a slope factor for oral, inhalation and dermal exposures. Hence background intake is not relevant to this assessment.

B7.4 Health Effects

General

There is no clinical disease which is unique to EDC toxicity. Primary effects are associated with the liver, kidneys and neurological, cardiovascular and immune systems.

EDC is readily absorbed into the body via inhalation, ingestion and dermal exposure. Following absorption into the body, EDC is widely distributed throughout the body. In animals the highest concentrations were generally within adipose tissue; however it is also distributed to the blood, liver, kidney, brain and spleen. EDC is metabolised extensively. Unmetabolised EDC is eliminated in expired air, while its metabolites (principally sulphur containing metabolites) are largely excreted in the urine. Although EDC is eliminated more slowly from adipose tissue than from blood or other tissues (lung and liver) following exposure, it is unlikely to bioaccumulate significantly.

The following summary has been derived from ATSDR (ATSDR 2001):

Acute toxicity: Breathing EDC can irritate the nose, throat and lungs causing coughing, shortness of breath and difficulty in breathing. Higher levels can cause a build-up of fluid in the lungs (pulmonary oedema). This can cause death. Exposure can cause nausea, vomiting, headaches, increasing drowsiness and then loss of consciousness. Over -exposure can also cause liver and kidney damage, and irritate the eyes. Contact can irritate the skin causing redness and a rash, and irritate the eyes. The lethal oral dose of ethylene dichloride in humans has been estimated to be between 20 and 50 mLs.

Short and long term toxicity: Other than carcinogenicity (discussed below) EDC can irritate the lungs. Repeated exposure may cause bronchitis to develop with cough, phlegm and/or shortness of breath. Repeated, prolonged contact can chronically irritate the skin causing dryness, redness and a rash. Repeated, prolonged exposure can cause loss of appetite, nausea and vomiting, trembling and low blood sugar (with weakness). It may damage the liver and kidneys.



Available data on the carcinogenicity of EDC in humans are limited. There are no epidemiological studies which show an association between EDC exposure and cancer. There is convincing evidence of increases in the incidence of both common and rare tumours in experimental animals at several sites (including squamous cell carcinomas of the stomach, haemangiosarcomas, fibromas of the subcutaneous tissue and adenocarcinomas and fibroadenomas of the mammary gland in rats; and alveolar/bronchiolar adenomas, mammary gland adenocarcinomas, endometrial stromal polyp or endometrial stromal sarcoma combined and hepatocellular carcinomas in mice) following oral exposure studies (WHO 1998a).

The incidence of benign lung papillomas was significantly increased in mice following long-term dermal application of EDC, while a non-significant increase in the number of pulmonary adenomas per animal was reported in a screening bioassay on mice and in the incidence of benign mammary gland tumours in rats exposed by inhalation for 2 years (WHO 1998a). In addition more recent studies (Nagano et al. 1998; Nagano et al. 2006) show carcinogenic outcomes associated with inhalation exposures.

The genotoxicity of EDC has been extensively investigated in non-mammalian and mammalian test systems. Following review of the available data by WHO (WHO 1998a), EDC has been identified as genotoxic in in vitro and in vivo assays, and binds to DNA in rodents in vivo.


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. In relation to EDC the following is noted from ATSDR (ATSDR 2001):

Considering the consistency of effects in acutely exposed humans and animals, and data showing that the liver, kidney, and immune system are sensitive targets of lower-dose and longer-term inhalation and oral exposures in animals, it is reasonable to assume that effects in these tissues would also be seen in similarly exposed adults and children. Some studies suggest a limited indication of the potential susceptibility of children to immunotoxic effects, particularly after bolus ingestion by children, which could occur, for example, with accidental ingestion of older household products that contain EDC. Similar effects are noted in relation to inhalation exposures.

The synergistic effect of disulfiram (tetraethylthiuram disulfide) on EDC hepatotoxicity and carcinogenicity in animal studies suggests that individuals exposed concurrently to EDC and disulfiram, either in the rubber industry or medically (disulfiram is used as an anti-alcohol-abuse drug), have increased risk for liver toxicity.

Inactivation of plasma alpha-1-proteinase inhibitor has been proposed to be an important factor in the development of lung emphysema. The occurrence of a synergistic inactivation of plasma alpha-1 proteinase inhibitor by EDC and cigarette smoke components (acrolein and pyruvic aldehyde) in vitro suggests that smokers as well as those exposed to passive smoke may be more susceptible to lung emphysema following repeated exposure to EDC.


Further, those with genetically reduced plasma alpha-1-proteinase inhibitor, who are predisposed to emphysema, may be at increased risk.

Classification

EDC was classified as a "probable" human carcinogen (Category B2) by the US EPA (IRIS) for all routes of exposure based upon evidence from animal studies.

IARC has classified EDC in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans for carcinogenicity and sufficient evidence in experimental animals (IARC 1999a).

NICNAS has not classified EDC.

B7.5 Quantitative Toxicity Values On the basis of the weight of evidence, EDC is a genotoxic carcinogen via all routes of exposure and hence it is relevant that it is assessed on the basis of a non-threshold approach. On this basis, the following quantitative values are available for EDC from relevant Australian and International sources:

Table B6 Summary of Published Toxicity Reference Values: EDC


SF = 0.01 (mg/kg/day)-1 The Australian Drinking Water Guidelines have derived a drinking water guideline of 0.003 mg/L for EDC using an oral slope factor of 0.01 (mg/kg/day)-1 and lifetime excess cancer risk of 1 in 1,000,000.


SF = 0.01 (mg/kg/day)-1 The WHO have established a guideline of 0.03 mg/L using a linearised multistage model and an excess lifetime cancer risk of 1 in 100,000. This corresponds to an oral slope factor of 0.01 (mg/kg/day)-1 (as used by NHMRC). It is noted that a lower guideline was proposed as part of the review process, however this has not been included in the final, and current, WHO drinking water guideline.

WHO (WHO 2000c)

NA WHO has undertaken a review of 1,2-dichloroethane for inhalation exposures in the late 1990’s. The review indicates that there is sufficient evidence of carcinogenicity in animals based on oral ingestion data. However, animal inhalation data reviewed at the time did not provide sufficient evidence of carcinogenicity and hence the WHO did not provide a carcinogenic assessment of inhalation exposures to EDC. Because of deficiencies in extrapolating oral data to inhalation, neither the oral slope factor nor any inhalation value have been recommended by the WHO in this assessment. A guideline value of 0.7 mg/m3 for a 24-hour average has been derived for non-carcinogenic endpoints by the WHO based on a lowest-observed-adverse-effect level from animal studies. It is noted that this guideline value recommended for the assessment of accidental release episodes or specific indoor pollution problems. The evaluation by WHO in this document was undertaken prior to the publication of studies by Nagano et al (Nagano et al. 1998; Nagano et al. 2006). The studies provided by Nagano et al show that inhalation exposures to EDC result in DNA damage and carcinogenic effects in animals. Hence it is considered reasonable to assume that EDC is a genotoxic carcinogen via inhalation routes of exposure as well as oral exposures.

WHO (WHO UR = (0.6 to 2.8)x10-6 A web-based source provides a summary of available inhalation


Source Value Basis/Comments 2000a) ( g/m3)-1 exposure assessments available from the WHO (EHC and CICAD

documents). For EDC the document references the CICAD published by the WHO (WHO 1998a) where a range of inhalation unit risk values for exposure to 1,2-dichloroethane in air have been presented based on tumour formation (based on oral studies). The range of inhalation unit risk values is (0.6 to 2.8)x10-6 ( g/m3)-1 (i.e. for an air concentration of 1

g/m3, the lifetime risk is estimated to be (0.6 to 2.8)x10-6). RIVM (Baars et al. 2001)

SF = 0.007 (mg/kg/day)-1 UR = 2x10-6 ( g/m3)-1

EDC is considered a genotoxic compound and as such oral and inhalation exposure is assessed on the basis of a non-threshold approach adopting an acceptable cancer risk level of 1 in 10,000 lifetime risk. On this basis, an oral cancer risk intake level of 0.014 mg/kg/day and a provisional inhalation cancer risk air concentration of 0.048 mg/m3 (based on oral data) have been established. If considered for a lifetime cancer risk level of 1 in 1,000,000, the air guideline would be equivalent to the lower guideline value established by the WHO (1998)

ATSDR (ATSDR 2001)

NA ATSDR does not address carcinogenic risks. A chronic and intermediate inhalation MRL, and an intermediate oral MRL, is available for the assessment of non-carcinogenic effects.

US EPA IRIS (USEPA)

SF = 0.091 (mg/kg/day)-1 UR = 2.6x10-5 ( g/m3)-1

The USEPA (IRIS, last uptated in 1987) has derived an oral slope factor of 0.091 (mg/kg/day)-1 for EDC based on a linear multistage model based on hemangiosarcomas in rats; and an inhalation unit risk of 2.6x10-5 ( g/m3)-1 using a linear multistage model based on oral data used to derive the oral slope factor. The USEPA does not present any data relevant to the assessment of non-carcinogenic effects for EDC.

OEHHA (OEHHA 1999, 2000b)

UR = 2.1x10-5 ( g/m3)-1 The California Air resources Board (OEHHA) has established inhalation unit risk value of 2.1x10-5 ( g/m3)-1 and a chronic reference exposure level for EDC of 0.4 mg/m3 based on hepatotoxicity (elevated liver enzyme levels in serum of rats). No acute reference exposure levels have been established.

TCEQ (TCEQ 2016)

TC = 0.44 mg/m4 UR = 3.4x10-6 ( g/m3)-1

TC based on liver and kidney toxicity (increased ALT and uric acid in serum) in rats, with use of a 180 fold uncertainty factor. Unit risk based on increased incidence of combined mammary gland tumours in female rats. Based on this study, a non-threshold air guideline of 0.0029 mg/m3 has been derived for a 1x10-5 risk.

While a range of values are available, those available from the ADWG (NHMRC 2011 Updated 2016) and WHO (WHO 2000a) have been adopted in this assessment. These values are current and provide a sound basis for the assessment of EDC, with the higher unit risk value (most conservative value) from the WHO adopted.


B8 Tetrachloroethene (PCE) B8.1 General Tetrachloroethene (also known as tetrachloroethylene, perchloroethylene, ethylene tetrachloride, per, perc, perchlor, 1,1,2,2-tetrachloroethylene and commonly abbreviated to PCE) is a synthetic chemical that is widely used for dry cleaning of fabrics and for metal-degreasing operations. It is also used as a building block for making other chemicals and is used in some consumer products. PCE manufacture in Australia ceased in 1991. Use in Australia has declined from 1995, consistent with declining use worldwide. PCE is primarily imported in its “pure” form with approximately 80 % used in the dry cleaning industry in Australia (NHMRC 2011 Updated 2016).

PCE is widespread in the environment and is found in trace amounts in water, aquatic organisms, air, foodstuffs, and human tissue. The highest environmental levels of PCE are found in the commercial dry-cleaning and metal-degreasing industries. The Australian Drinking Water Guidelines indicate that PCE has not been detected in Australian drinking water supplies (NHMRC 2011 Updated 2016).

Exposure to PCE may be derived from environmental and occupational sources as well as from consumer products. Common background levels of PCE in the environment are generally several thousand times lower than levels found in some workplaces. Background levels are found in the air, water, and food. The most significant exposure pathway is via the air, particularly in the workplace. PCE gets into air by evaporation from industrial or dry cleaning operations and released from stores of chemical wastes. It is frequently found in surface water (NHMRC 2011 Updated 2016).

Common consumer products that may contain PCE include water repellents, silicone lubricants, fabric finishers, spot removers, adhesives, and wood cleaners. Although uncommon, small amounts of PCE have been found in food, especially food prepared near a dry cleaning facility (NHMRC 2011 Updated 2016).

The respiratory tract is the primary route of entry for PCE. The chemical is rapidly absorbed by this route and reaches equilibrium in the blood within 3 hours after the initiation of exposure. PCE is also significantly absorbed by the gastrointestinal tract, but not through the skin. The chemical accumulates in tissues with high lipid content, where the half-life is estimated to be 55 hours (RAIS). PCE has also been detected in the breast milk of mothers who have been exposed to the chemical. PCE is considered to have a low potential for bioaccumulation (NICNAS 2001).

In relation to the assessment of dermal absorption, as noted in USEPA (2004), dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg) (USEPA 1995a). For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%, which includes compounds such as PCE. No data is available specifically for the dermal absorption of PCE; hence a default value of 3% is considered appropriate.

If released into the environment the following can be noted with respect to PCE:


Air: PCE is expected to remain in vapour phase. Removal is primarily through reaction with hydroxyl radicals, or chlorine atoms produced through photo-oxidation of PCE, which results in half-lives of 1 hour to 2 months.

Soil and Water: PCE is expected to volatilise from surface soils and water. PCE has a low to medium mobility in soil and may leach slowly through soil into groundwater where it may persist for years. Depending on conditions reductive dehalogenation to vinyl chloride may occur. Under anaerobic conditions PCE and TCE can be intrinsically biodegraded to form DCE and vinyl chloride.

B8.2 Background Exposures/Intake As PCE is highly volatile and not persistent, background intakes will be dominated by inhalation exposures. Concentrations of PCE in industrial, urban and regional areas are available in Australia. Data collected in NSW (DEC, 2003) from urban and regional areas in NSW report average concentrations of PCE of approximately 0.1 ppbv, or 0.0007 mg/m3 with a maximum concentration in Sydney CBD of 1.6 ppbv, or 0.01 mg/m3 (NSW DEC 2004). Concentrations in an industrial area in Brisbane have reported average and maximum concentrations reported of 0.015 mg/m3 and 0.085 mg/m3 respectively (Hawas et al. 2001). These concentrations are consistent with those reported in other cities in Australia (NICNAS 2001).

The average air concentration reported by DEC comprises less than 5% of the TC adopted for PCE (NSW DEC 2004).

Other significant exposures to the general public are likely to occur through the use of dry cleaning. Variable concentrations of PCE in homes and by personal samplers where dry-cleaned clothes are stored and worn are reported by NICNAS and WHO (NICNAS 2001; WHO 2000c). A study on the effect of wearing dry-cleaned clothes reported median personal air concentrations ranging of 0.032 mg/m3 to 0.22 mg/m3 depending on the garment. These exposures along with exposures to paint solvents and cleaning material containing PCE were considered potentially significant. No estimate of intake by the general public is provided in the NICNAS review. Median indoor air concentrations reported by WHO for homes not located in the same building as dry-cleaners was 0.004 mg/m3 (note that concentrations indoors were much higher in buildings with a dry-cleaners with indoor air levels ranging from 0.05 to 6.1 mg/m3) (WHO 2006). This value is also essentially negligible compared with the adopted TC. While there is the potential for increased background intakes depending on consumer use of products and frequency of dry-cleaning, average intakes are considered negligible.

To be conservative a background intake of 10% has been assumed in the quantification of risk. This adequately addresses intakes of PCE from other sources.

B8.3 Health Effects

General

There is no clinical disease which is unique to PCE toxicity. PCE is absorbed mainly through inhalation, causing both irritation and neurobehavioral effects. Skin burns, blistering and erythema can occur from severe direct contact with PCE. Some skin absorption can occur but does not appear to be of major significance. The amount of the chemical in the body increases with


increasing exposure level and with an increase in physical exercise during exposure. It accumulates to a limited extent in the fatty tissues of man and of animals. Because of its affinity for fat, PCE is found in milk. PCE has also been shown to cross the placenta and distribute to the foetus (WHO 2006).

PCE is eliminated slowly through the lungs. A small amount is metabolised to trichloroethanol and trichloroacetic acid. The concentrations of the compound in blood and breath can be used for estimating exposure levels in humans (WHO 2006).

The main targets of PCE toxicity are the liver and kidney by both oral and inhalation exposure, and the central nervous system by inhalation exposure (WHO 2006).

Short-term (one hour or less) inhalation exposure of humans to high concentrations of PCE can result in irritation of the upper respiratory tract and eyes and neurological effects such as headache, dizziness, sleepiness, impairment of coordination, and reversible mood and behavioural changes (WHO 2006).

Chronic exposure causes respiratory tract irritation, headache, nausea, sleeplessness, abdominal pains, constipation, cirrhosis of the liver, hepatitis, and nephritis in humans; and microscopic changes in renal tubular cells, squamous metaplasia of the nasal epithelium, necrosis of the liver, and congestion of the lungs in animals (WHO 2006).


Some epidemiological studies indicate a possible association between chronic exposure to PCE and an increased cancer risk; however the evidence provided is considered to be inconclusive (USEPA 2012). This is mainly due to concurrent exposure to other petroleum solvents as well as PCE, confounding factors (smoking, alcohol, socio-economic status) and small numbers of cancers in the studies (USEPA 2012).

An association between exposure to PCE (inhalation and ingestion) and an increased risk of cancer (mononuclear cell leukaemia and hepatic tumours) in animals has been suggested. Review of PCE by WHO indicates that PCE is a non-genotoxic animal carcinogen (WHO 2000c). Review of the possible mechanisms of tumour formation by PCE in animals suggests that the tumours observed may have little relevance for humans. This is subject to some debate, however, recent reviews by WHO and USEPA have noted that in the absence of suitable supporting evidence to the contrary, it must be concluded that the cancers produced by PCE in rodents are of potential relevance to humans (USEPA 2012; WHO 2006).

From the weight of evidence, PCE does not appear to have significant genotoxic potential, however some of the possible metabolites are recognised Ames bacterial mutagens (Baars et al. 2001; WHO 2000a, 2006). Review of the available studies by WHO suggests that non-genotoxic mechanisms have been recognised for the formation of kidney tumours in male rats and liver tumours in mice for some chemicals (WHO 2006). The available data on mechanism/mode of action (MOA) for PCE are limited, and the dose–response data related to these recognised mechanisms are not consistent with the dose–response relationships for cancer induction by PCE. WHO has derived a threshold inhalation value for PCE that is considered protective of key end-points including carcinogenicity


(WHO 2006). Hence, it may be considered appropriate that a threshold dose-response approach be adopted for PCE.

It is noted that the review of PCE by USEPA suggests that PCE has been shown to induce some genotoxic effects (USEPA 2012). There are a number of limitations noted in the assessment presented by the USEPA, in particular the factor that the MOA for PCE induces carcinogenesis is not yet fully characterised or understood and that the role of genotoxicity in hepatocarcinogenicity is uncertain. Where the USEPA lacks certainty the default position is to be conservative and as such they have suggested considering PCE having a mutagenic MOA, where a non-threshold approach is recommended for the assessment of carcinogenicity and mutagenicity. The assessment of PCE should be updated should additional data become available that supports the USEPA review.


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. In relation to PCE the following is noted from the USEPA review (USEPA 2012):

There is some evidence that certain subpopulations may be more susceptible to exposure to PCE. These subpopulations include early and later life stages, health and nutrition status, gender, race/ethnicity, genetics, and multiple exposures and cumulative risk.

Data-derived non-cancer outcomes of concern in early life stages are spontaneous abortion/foetal loss, mortality, and neurological impairment. As described above, the evidence for spontaneous abortion following prenatal exposures to PCE is well characterized in humans, and foetal loss is well characterised in experimental animals. However, the human epidemiological data that support this conclusion do not provide information on the maternal dose to PCE that may have resulted in spontaneous abortion. Further, data from the experimental animal studies suggest that this finding may be a high-dose effect. Together, this evidence suggests that reference values that are established on the basis of more sensitive neurological endpoints should mitigate potential risk of foetal lethality.

While limited studies show some early childhood responses associated with exposures to PCE these are not well characterised and cannot be related to the larger population.

In addition, although not definitive, studies further suggest that the developing foetus is susceptible to maternal organic solvent exposures.

There is suggestive evidence that there may be greater susceptibility for exposures to the elderly.

Diminished health status (e.g., impaired kidney liver or kidney) will likely affect an individual’s ability to metabolise PCE, whereas certain nutrients may have a protective effect on exposure.


Gender and race/ethnic differences in susceptibility are likely due to variation in physiology and exposure, and genetic variation likely has an effect on the toxicokinetics of PCE.

Multiple and cumulative exposures are likely to cause competition in metabolic capacity.

Classification

PCE was classified as a "likely" human carcinogen by the USEPA for all routes of exposure based upon evidence from animal studies (USEPA 2012).

IARC has classified PCE in Group 2A (probably carcinogenic to humans) based in limited evidence in humans (epidemiological studies showed elevated risks for oesophageal cancer, non-Hodgkin's lymphoma and cervical cancer) and sufficient evidence in experimental animals (induce peroxisome proliferation in mouse liver and induced leukaemia in rats) (IARC 2014).

NICNAS has classified PCE as a Carcinogen Category 3, which is a substance regarded as a possible risk of irreversible effects (NICNAS 2001).

B8.4 Quantitative Toxicity Values On the basis of the weight of evidence, PCE does not appear to have significant genotoxic potential. On this basis it is considered reasonable that a threshold approach is adopted for the characterisation of all health effects including carcinogenicity. On this basis, the following quantitative inhalation values are available for PCE from Level 1 Australian and International sources:

Table B7 Summary of Published Toxicity Reference Values: PCE


TDI = 0.014 mg/kg/day The current ADWG has derived a guideline of 0.05 mg/L for PCE based on a NOEL of 14 mg/kg/day from a 90-day drinking water study in rats and mice and an uncertainty factor of 1000. The uncertainty factor includes an additional 10 fold factor to address possible carcinogenicity.


TDI = 0.014 mg/kg/day WHO DWG TDI based on the same study and uncertainty factor as noted in the ADWG.

WHO (WHO 2006)

TC = 0.2 mg/m3

TC in air derived on the basis of the most sensitive endpoint, namely neurotoxicological effects, based on a LOAEC (adjusted) of 20 mg/m3 based on an occupational inhalation study (mean exposure of 10 years) and an uncertainty factor of 100. The TC derived lower that that derived on the basis of other key end-points such as kidney and liver effects and reproductive/developmental effects. Potential carcinogenic effects have been assessed on the basis of a benchmark dose approach with a BMCL10 of 20 mg/m3 and if a multistage model were considered the TC of 0.2 mg/m3 would be associated with a risk of 1 x10-3. The assessment presented by (WHO 2006) is an update of the earlier assessment presented in the WHO Air Quality Guidelines (2000) where a TC of 0.25 mg/m3 was derived (WHO 2000c). Concern about potential carcinogenic effects are noted (and should be addressed through an in-depth risk assessment) in the guideline. TDI derived using a PBPK model using the inhalation TC data.

WHO (WHO TDI = 0.05 mg/kg/day The available information on oral exposure was inadequate for derivation


Source Value Basis/Comments 2006) of a TDI by the oral route. However, as tetrachloroethene is well absorbed

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



TDI derived on the basis of a NOAEL of 16 mg/kg/day associated with liver effects in a 4-week oral study in rats and an uncertainty factor of 1000. TC adopted based on older (WHO 2000c) evaluation derived from a LOAEL (adjusted) of 23 mg/m3 from an occupational inhalation study and an uncertainty factor of 100.

Health Canada (Health Canada 1993b)


TDI derived on the same basis as noted for the WHO DWG and ADWG. TC derived from a LOAEL of 363 (adjusted) mg/m3 associated with multiple effects in mice and an uncertainty factor of 1000.

ATSDR (ATSDR 2014a)

No chronic oral MRL Inhalation MRL =0.24 mg/m3

The chronic inhalation MRL has been derived on the basis of a LOAEL (adjusted) of 24 mg/m3 associated with neurobehavioural effects in an occupational inhalation study and an uncertainty factor of 100.

USEPA (USEPA 2012)


RfD derived on the basis of neurotoxiocity in an inhalation occupational study, with route-extrapolation and an uncertainty factor of 1000. RfC derived on the basis of LOAELs of 15 and 56 mg/m3 associated with CNS effects in occupationally exposed adults and an uncertainty factor of 1000. The USEPA review has also presented non-threshold values for the assessment of carcinogenicity.

For oral exposures the TDI adopted is from the ADWG (NHMRC 2011 Updated 2016) and WHO DWG (WHO 2011). This value is consistent with threshold values derived from oral studies from RIVM and Health Canada. The most recent USEPA oral value is not derived on the basis of oral studies, rather it is based on route-extrapolation from the inhalation study. It is preferable to adopt TRVs that are based on studies relevant to the route of exposure assessed. This value has also been adopted for the assessment of dermal exposures.

For inhalation exposures the TC adopted is the value available from the more recent WHO review (WHO 2006). This is consistent with that derived by other agencies, with the exception of the USEPA value which is approximately 10 times lower. The USEPA review has considered the same key studies as the WHO, however, they have applied more conservative uncertainty factors, which the WHO did not consider appropriate. Hence the WHO inhalation value has been adopted.


B9 Trichloroethene (TCE) B9.1 General Trichloroethene (also known as 1,1,2-trichloroethylene, ethylene trichloride, and commonly abbreviated to TCE) is a synthetic product that was first prepared in 1864 by the reduction of hexachloroethane (HCE) with hydrogen. It is mainly used as a liquid or vapour degreasing solvent, particularly in the metal fabricating industry. International concern about the environmental and health and safety concerns of chlorinated hydrocarbons has reduced the use of TCE (ATSDR 2014b).

TCE was manufactured in Australia from the 1950’s to the early 1980’s, with current demand met by imports of the chemical. TCE is also recycled in Australia. TCE is used widely in both large and small industries in Australia for vapour degreasing, cold cleaning as well as use in adhesives, waterproofing agents, paint strippers, carpet shampoos and some other cleaning products. It is also an effective cleaning agent for organic materials as it has a low latent heat of vaporisation and is non-flammable (ATSDR 2014b).

Exposure of the general population to TCE may be by inhalation, oral or dermal routes. In most cases inhalation is the primary route of exposure. Exposure may occur through oral ingestion of drinking water or soil; however exposure to TCE in food is generally low. Apart from occupational exposures, the primary concern is inhalation indoors. TCE in the outdoor air may originate from indoor or outdoor sources. Outdoor sources include outdoor air, contaminated soils or groundwater. Indoor air sources include new building construction materials or home cleaning products. The potential for bioaccumulation of TCE is considered to be low (ATSDR 2014b).

With respect to the potential for dermal absorption, as noted by USEPA, dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg) (USEPA 1995a). For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%, considered to include compounds such as TCE. While dermal absorption of TCE from solvent products or an aqueous solution may be important, data on dermal absorption from soil is limited. Information presented by CCME notes that limited studies have shown that for soil the maximum amount absorbed via the skin is 0.8%, hence a default value of 3% is considered appropriate (CCME 2007).

If released into the environment the following can be noted with respect to TCE:

Air: TCE is expected to remain in vapour phase. Removal is primarily through reaction with hydroxyl radicals to produce low levels of phosgene, dichloroacetyl chloride, formyl chloride and other degradation products. Half-life pf TCE varies from 1 day to months.

Soil and Water: TCE is expected to volatilise from surface soils and water. TCE may leach through soil into groundwater where it may persist for years.


Water: Depending on conditions reductive dehalogenation to vinyl chloride may occur. Under anaerobic conditions TCE can be intrinsically biodegraded to form DCE and vinyl chloride.

B9.2 Background Exposures/Intake As TCE is highly volatile and not persistent, background intakes will be dominated by inhalation exposures. Concentrations of TCE in industrial, urban and regional areas are available in Australia. Data collected in NSW from urban and regional areas in NSW report average concentrations of TCE of approximately 0.1 ppbv, or 0.0005 mg/m3 with a maximum concentration in Sydney CBD of 3.6 ppbv, or 0.019 mg/m3 (NSW DEC 2004). Concentrations in an industrial area in Brisbane have reported average and maximum concentrations reported of 0.0002 mg/m3 and 0.0005 mg/m3 respectively (Hawas et al. 2001). Background air concentrations in Canada are considered to be approximately 0.0014 mg/m3, consistent with the range reported by NSW DEC (CCME 2007). Background intakes (dominated by inhalation) were estimated by the WHO Drinking Water Guidelines to be approximately 0.04 μg/kg/day for children and 0.01 μg/kg/day for adults (WHO 2011). On the basis of the available information, background intakes of TCE are expected to be low, with 10% assumed in the calculations.

B9.3 Health Effects

General

There is no clinical disease which is unique to TCE toxicity. In the past, TCE was used as a human anaesthetic. TCE has also been inhaled by people intentionally for its narcotic effect. Hence most toxicological data is associated with inhalation exposures. Primary effects are associated with the central nervous system (CNS) (ATSDR 2014b; USEPA 2011a).

TCE can be absorbed into the body via inhalation, ingestion and dermal exposure. Following absorption into the body, TCE is distributed to the blood, then transported to various tissues where it is metabolised. The toxicities associated with TCE are thought to be mediated by metabolites rather than the parent compound. Major sites of TCE distribution appear to be the body fat and liver (ATSDR 2014b; USEPA 2011a).

Humans and animals excrete un-metabolised TCE via expiration, while the metabolites are excreted primarily in urine. Urinary metabolites include trichloroacetaldehyde, trichloroethanol, and trichloroacetic acid; the reactive epoxide TCE oxide is an essential feature of the metabolic pathway (ATSDR 2014b; USEPA 2011a).

Human and animal data indicate that exposure to TCE can result in toxic effects on a number of organs and systems, including the liver, kidney, blood, skin, immune system, reproductive system, nervous system, and cardiovascular system. In humans, acute inhalation exposure to TCE causes central nervous system symptoms such as headache, dizziness, nausea, and unconsciousness. Among the reported effects during occupational exposure studies are: fatigue, light-headedness, sleepiness, vision distortion, abnormal reflexes, tremors, ataxia, nystagmus, increased respiration, as well as neurobehavioral or psychological changes. Cardiovascular effects include tachycardia, extrasystoles, EKG abnormalities, and precordial pain. The use of TCE as an anesthetic has been associated with cardiac arrhythmias (RAIS).


Cases of severe liver and kidney damage, including necrosis, have been reported in humans following acute exposure to TCE, but these effects generally are not associated with long-term occupational exposures. In animals, TCE has produced liver enlargement with hepatic biochemical and/or histological changes and kidney enlargement, renal tubular alterations and/or toxic nephropathy. Also observed in animals were hematological effects and immunosuppression. Inhalation studies with rats indicate that TCE is a developmental toxicant causing skeletal ossification anomalies and other effects consistent with delayed maturation. TCE may cause dermatitis and dermographism (RAIS).


Some epidemiological studies indicate a possible association between exposure to TCE and an increased cancer risk, with IARC (1995) noting elevated risk for cancer of the liver and biliary tract and a modestly elevated risk for non-Hodgkin’s lymphoma in three cohort studies. In animals TCE induces tumours at several sites and in different species. Tumours have been seen in mouse liver and lung and rat kidney and testis. On the basis of the available information, most current reviews by IARC, WHO, CCME and USEPA (2009) consider TCE to be carcinogenic (with responses tending to increase with dose) via all routes of exposure (CCME 2007; IARC 2014; USEPA 2011a; WHO 2011).

The potential mode of action (MOA) for TCE is reviewed and discussed in the current WHO drinking water guidelines and USEPA reviews (USEPA 2011a; WHO 2011).

The WHO DWG review concluded that the MOA for tumour induction by TCE may be attributed to non-genotoxic processes (related to cytotoxicity, peroxisome proliferation and altered cell signalling); genotoxic processes, (such as the production of genotoxic metabolites (e.g., chloral and DCVC1)); or the production of reactive oxygen species related to peroxisomal induction in the liver (WHO 2011). The potential role of several mutagenic or carcinogenic metabolites of TCE cannot be ignored. Hence TCE appears to be at least weakly genotoxic and evaluation of carcinogenicity on the basis of a non-threshold approach is considered appropriate (as is undertaken in the current WHO DWG and WHO Air Quality Guidelines (WHO 2000c, 2011)).

The most recent USEPA review provides a detailed assessment of genotoxicity (of TCE and metabolites) and mutagenicity. With respect to genotoxicity, although it appears unlikely that TCE, as a pure compound, causes point mutations, however there is evidence for TCE genotoxicity with respect to other genetic endpoints, such as micronucleus formation (USEPA 2011a). In addition, several TCE metabolites have tested positive in genotoxicity assays. It is noted that uncertainties with regard to the characterisation of TCE genotoxicity remain, particularly because not all TCE metabolites have been sufficiently tested in the standard genotoxicity screening battery to derive a comprehensive conclusion. However, the metabolites that have been tested particularly DCVC have predominantly resulted in positive data, supporting the conclusion that these compounds are genotoxic.

The MOA relevant to specific target organs in laboratory animals have been reviewed by the USEPA (USEPA 2011a). Only in the case of the kidney is it concluded that the data are sufficient to

1 DCVC is the abbreviation for the metabolite S-(1,2-dichlorovinyl)-L-cysteine.


support a particular MOA being operative. For the kidney, the predominance of positive genotoxicity data in the database of available studies of TCE metabolites together with toxicokinetic data supports the conclusion that a mutagenic MOA is operative in TCE-induced kidney tumours. Hence a linear (non-threshold) approach is recommended for the quantification of carcinogenic effects.

Classification

TCE has been classified by the USEPA as “carcinogenic to humans” by all routes of exposure based on convincing evidence of a causal association between TCE exposure in humans and kidney cancer, and evidence of TCE carcinogenicity in liver and lymphoid tissues. The data is supported by animal studies. IARC has classified TCE in Group 1 (carcinogenic to humans) based in limited evidence from several human epidemiological studies and on sufficient evidence from animal studies (IARC 2014; USEPA 2011a).

NICNAS has classified TCE as a Carcinogen Category 2, which is a substance regarded as if it is carcinogenic to humans, on the basis of the occurrence of tumours in experimental animals and limited evidence in workers (NICNAS 2000).


The USEPA review has concluded that there is sufficient weight of evidence that TCE operates through a mutagenic mode of action (MOA) for kidney tumours and there is a lack of TCE-specific quantitative data in relation to early lifetime susceptibility. Hence it may be appropriate to consider increased susceptibility associated with early lifetime exposures through the adjustment of exposure factors. This adjustment, however is noted to be relevant to the kidney cancer component of the total risk (note the inhalation unit risk includes a factor of 4 fold to address the risk of tumours at multiple sites). The effect of considering theses age adjusted exposure factors to only the kidney cancer portion of the unit risk has been evaluated by the USEPA and determined to be of minimal impacts to the total cancer risk, except when exposure only occurs during early life (if these effects occur). In addition to this evaluation a number of uncertainties have been identified in relation to applying the age adjustment factors for a more complex carcinogenic MOA as identified for TCE. Hence for the purpose of assessing long-term exposures, no further adjustments to account for potential early lifetime susceptibility requires consideration (USEPA 2011a).

B9.4 Quantitative Toxicity Values On the basis of the above it is reasonable to consider a non-threshold approach for the assessment of carcinogenicity in relation to TCE. It is noted that a number of guidelines (such as WHO DWG) have been derived on the basis of both carcinogenic and non-carcinogenic endpoints, with non-carcinogenic end-points noted to be more sensitive for at least oral intakes. Hence both non-threshold and threshold values available have been noted in the following.

The following quantitative inhalation values are available for TCE from relevant Australian and International sources:


Table B8 Summary of Published Toxicity Reference Values: TCE

Source Value Basis/Comments Australian ADWG No health based value

derived Not derived due to insufficient data.


SF = 0.00078 (mg/kg/day)-1 TDI = 0.00146 mg/kg/day

WHO DWG (provisional guideline) is based on the lower value derived from carcinogenic and non-carcinogenic endpoints. It is noted that the guideline derived on the basis of reproductive/developmental (threshold) effects was most conservative. The oral slope factor adopted is from Health Canada (range of values derived) and based on combined tubular cell adenomas and adenocarcinomas of the kidneys in rats following oral exposure to TCE for 103 weeks and a linear multistage model. The oral TDI derived from a BMDL10 of 0.146 mg/kg/day associated with reproductive/developmental effects in rats and an uncertainty factor of 100.

WHO (2000 and reviewed again in 2010) (WHO 2011)

UR = 4.3x10-7 ( g/m3)-1 Inhalation unit risk derived on the basis of Leydig-cell tumours in the testes of rats and a linear multistage model. Inhalation unit risk from rat study is supported by a similar unit risk of 9 x10-7 ( g/m3)-1 derived from increased incidence of hepatic tumours in a cohort study of occupationally exposed adults. The non-threshold approach was adopted by the WHO as TCE was considered genotoxic and carcinogenic.

EU (EU 2004) SF = 0.0019 (mg/kg/day)-1 TCE gives rise to concern for humans owing to possible mutagenic and carcinogenic effects and because it is not possible to identify a threshold exposure level below which these effects would not be expressed. For non-carcinogenic effects the most sensitive threshold effect evaluated was associated with CNS disturbance following repeated dose where a NOAEL of 38 mg/kg/day was considered. The EU has presented a calculation of lifetime cancer risk based on the T25 method in relation to non-Hodgkin lymphoma. From an inhalation study in female mice a HT25 dose descriptor for humans was derived as 130 mg/kg/day. Following the approach presented the EU calculated increased cancer risk for TCE for all groups using an equivalent slope factor of 0.0019 (mg/kg/day)-1. This value was used in the quantification of risk associated with exposure from oral, dermal and inhalation pathways for workers, consumers and environmental exposures.

European Chemicals Agency (ECHA 2014)

UR = 6.9x10-7 for exposures to 6.2 mg/m3 or greater UR = 6.4x10-8 for exposures to less than 6.2 mg/m3

The assessment of carcinogenic effects was determined to be appropriately addressed on the basis of a non-threshold approach. Human data on the formation of kidney cancer in occupational environments was determined to be relevant for establishing a dose-response relationship. Increased risks were only determined at high exposure concentrations, where the cytotoxic effects were considered to enhance the carcinogenic response. Below this level the risk of cancer is lower. Hence a sub-linear response has been considered for exposure to TCE.

Health Canada (Health Canada 2005a)

SF = 0.000811 (mg/kg/day)-1 UR = 1.2x10-7 ( g/m3)-1 TDI = 0.00146 mg/kg/day

Oral slope factor derived on the basis of the same study noted in the WHO DWG, however a slightly different value is quoted. Inhalation unit risk based on renal adenocarcinomas in rats following inhalation exposures for 104 weeks in males (a lower, less conservative value was derived for females). Note that the derivation of drinking water guidelines has also considered the oral TDI noted in the WHO DWG which results in a lower guideline than is derived on the basis of the oral slope factor.

CCME (CCME 2007)

SF = 0.000811 (mg/kg/day)-1 UR = 6.4x10-7 ( g/m3)-1 TDI = 0.00146 mg/kg/day TC = 0.04 mg/m3

Slope factor based on same study noted by Health Canada (2005). Inhalation unit risk based on older evaluation from Health Canada where a TC05 (concentration expected to cause a 5% incidence in cancer) of 0.082 mg/m3 and extrapolation based on an excess lifetime cancer risk of 10-6. TDI and TC values also presented for non-carcinogenic endpoints. TDI as noted by WHO DWG TC adopted from the USEPA RfC associated with effects on the CNS,


Source Value Basis/Comments kidney, liver and endocrine system in inhalation studies where a point of departure (POD) of 38 mg/m3 was identified and an uncertainty factor of 1000 adopted.


PTDI = 0.05 mg/kg/day PTC = 0.2 mg/m3

Provisional threshold values derived for TCE due to limited data and an assumption that the genotoxic mechanism for TCE (numerical chromosome aberration in vivo) exhibits a threshold. The basis for these values is not listed here as the evaluation is considered dated.

ATSDR (ATSDR 2014b)

No chronic MRLs derived No chronic oral or inhalation MRL has been established.

USEPA (USEPA 2011a)

SF = 0.046 (mg/kg/day)-1 RfD = 0.0005 mg/kg/day UR = 4x10-6 ( g/m3)-1 RfC = 2 μg/m3

Oral slope factor based on PBPK model-based route-extrapolation of the inhalation unit risk estimate for kidney cancer with a 5 fold factor included to address potential risk of tumours from multiple sites. RfD based on oral studies associated with heart malformations in rats and immunotoxicity in mice, use of a PBPK model to calculate a human equivalent dose and various uncertainty factors that ranged from 10 to 100. Inhalation unit risk derived on the basis of non-Hodgkin’s lymphoma , renal cell carcinoma and liver tumours from a human inhalation (epidemiology) studies, application of a PBPK model and adjustment (by a factor of 4) to address potential risk of tumours at multiple sites. The value is derived from linear extrapolation from the point of departure (LEC01). RfC based on route-extrapolation from oral studies for the critical effects of heart malformations in rats and immunotoxicity in mice and incorporation of uncertainty factors ranging from 10 to 100.

Based on the evaluation undertaken by the USEPA (USEPA 2011a) the health end-points associated with carcinogenic (non-threshold) and non-carcinogenic (threshold) effects are similar in sensitivity. This is outcome is unusual and not consistent with evaluations undertaken by other agencies. If this observation is appropriate, then it is important that the assessment consider both threshold and non-threshold endpoints to ensure all relevant additive effects (from a number of CHCs) are considered.

Many of the reviews conducted by the WHO, CCME, RIVM and ATSDR have considered limited and dated databases of information (as noted) (ATSDR 2014b; Baars et al. 2001; CCME 2007; WHO 2000c). The review of TCE toxicity conducted by the USEPA is the most current and comprehensive review (USEPA 2011a). The review by the WHO in relation to inhalation toxicity, considered some of the more recent studies, however the review has not considered non-carcinogenic endpoints and the key studies considered by the USEPA for the derivation of the inhalation unit risk (USEPA 2011a; WHO 2010).

On this basis it is considered appropriate that the more recent evaluation conducted by the USEPA is utilised for the purpose of quantifying risk (USEPA 2011a). This approach is consistent with that adopted in the development of the interim HIL for TCE (NEPC 1999 amended 2013b).

It is noted that there has been recent criticism of many of the key studies identified and relied upon, and the methodology adopted in the PBPK modelling, in the USEPA review. These issues are being further discussed with additional studies proposed to resolve some of the outstanding issues. The outcome of these discussions and studies are not yet available. Given the conservative nature of the USEPA toxicity values (in comparison to criteria developed by other agencies), adopting the USEPA evaluation in this assessment is expected to overestimate risk.


B10 1,1-Dichloroethane B10.1 General 1,1-Dichloroethane (ethylidene chloride or ethylidene dichloride and abbreviated to 1,1-DCA) is a manmade organic solvent. It does not occur naturally, but is manufactured for industrial uses. It is primarily used as an intermediate in the manufacture of other chemicals such as vinyl chloride (and associated byproducts such as EDC tars), 1,1,1-trichloroethane and to manufacture high vacuum rubber. 1,1-Dichloroethane is also used in paint removers and it is a coupling agent in antiknock fuel. In the past the chemical was used as an anaesthetic, however that use has been discontinued when effects on the heart, such as irregular heartbeats, were reported (ATSDR 1990).

At room temperature, 1,1-dichloroethane is a colourless, orderly liquid that is very volatile with an aromatic chloroform/ether like odour (HSDB).

If released to air, 1,1-dichloroethane will exist solely as a vapour phase in the ambient atmosphere and will be degraded by reactions with photo chemically produced hydroxyl radicals with a half-life of approximately 49 days. If released to soil, 1,1-dichloroethane is expected to be highly mobile with volatilisation expected to be an important fate process. If released to water the chemical is not expected to adsorb to suspended solids or sediments with volatilisation expected to be an important process. Halogenated aliphatic hydrocarbons, such as 1,1-dichloroethane, are generally considered to be resistant to biodegradation. The potential for 1,1-dichloroethane to bioaccumulate in aquatic organisms is low (ATSDR 1990).

B10.2 Exposure The main route of exposure to 1,1-dichloroethane is via inhalation of contaminated ambient air. Exposure may also occur from ingestion of contaminated drinking water and use of consumer products such as paint removers that may contain this compound. Occupational exposure to 1,1-dichloroethane may occur through inhalation and dermal contact (HSDB).

1,1-Dicloroethane can be absorbed via inhalation, ingestion and skin absorption. The chemical is believed to be rapidly absorbed. Information on the distribution of 1,1-dichloroethane to tissues and organs following exposure is limited. The anaesthetic effects of 1,1-dichloroethane indicate that the chemical reaches the central nervous system and is probably distributed throughout the rest of the body. Most of 1,1-dichloroethane is usually removed unchanged from the body within the breath within two days. A small part of the chemical is metabolised (with major metabolite being acetic acid, and minor metabolites include 2,2-dichloroethanol, chloracetic acid and dichloroacetic acid) with breakdown products quickly are removed in the breath or urine (ATSDR 1990).

B10.3 Background Exposures/Intake Limited data is available regarding environmental levels of 1,1-dichloroethane in Australia. It is noted in the Australian Drinking Water Guidelines (NHMRC 2011 Updated 2016) that dichloroethanes have not been found in Australian drinking waters. Data available from NSW (NSW DEC 2004) indicates 1,1-dichloroethane was rarely (<1% of samples) detected in air during sampling undertaken from 1996 to 2001 in a number of urban and rural areas. Where detected in Sydney's CBD, the maximum 24-hour average concentration was reported to be 0.9 ppbV (0.0036


mg/m3). Data available from the WA (WA DEP 2000) indicates 1,1-dichloroethane was never detected in air. Consideration of the maximum reported concentration of 1,1-dichloroethane in air suggests an intake of 1 μg/kg/day (assuming inhalation of 20 m3 per day and a 70 kg body weight) which is considered negligible compared with the adopted threshold levels adopted. No data is available regarding potential concentrations in food, with potential concentrations in water assumed to be negligible.

B10.4 Health Effects

General

Acute (short-term) inhalation exposure to high levels of 1,1-dichloroethane in humans results in central nervous system (CNS) depression and a cardio stimulating effect resulting in cardiac arrhythmias. Studies in animals have reported effects on the kidney (ATSDR 1990).

No information is available on the chronic (long-term), reproductive, developmental, or carcinogenic effects of 1,1-dichloroethane in humans. Some studies in animals have shown on that 1,1-dichloroethane can cause kidney disease after long term, high level exposure in air. An oral animal study reported a significantly positive dose-related trend in hemangiosarcomas, mammary tumours, liver tumours, and endometrial stromal polyps (ATSDR 1990).


There are no human cancer data available. Limited animal studies have shown 1,1-dichloroethane to be a potential carcinogen. No studies are available regarding in vivo genotoxic effects in humans. Data available from limited in vitro and in vivo (very few) assays in animals are conflicting with very few studies reporting conclusive results. In general 1,1-dichloroethane is considered to be non-genotoxic. This has also been noted in the more recent review by Ontario (Ontario 2007).


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. In relation to 1,1-dichloroethane insufficient data is available for the purpose of identifying relevant susceptible populations.

Classification

1,1-Dichloroethane is classified by the USEPA as Group C possible human carcinogen based on no human data and limited evidence in two animal species (rats and mice) from inhalation studies.

IARC has not evaluated 1,1-dichloroethane.


B10.5 Quantitative Toxicity Values Review of available data with respect to 1,1-dichloroethane indicates that the chemical is not considered to be genotoxic to humans (based on limited data) with limited data available with respect to carcinogenicity. On this basis it is considered relevant to consider potential exposures to 1,1-dichloroethane on the basis of a threshold approach. On this basis, the following quantitative values are available for 1,1-dichloroethane from relevant Australian and International sources:

Table B9 Summary of Published Toxicity Reference Values: 1,1-Dichloroethane

Source Value Basis/Comments Australian – no evaluations available International WHO DWG (WHO 2011)

NA WHO not that in view of the very limited database on toxicity and carcinogenicity, it was concluded that no guideline value should be proposed.

ATSDR (ATSDR 1990)

No chronic oral MRL No MRLs have been established.

USEPA (USEPA)

NA No threshold, or non-threshold, evaluation is available. A provisional value of 0.5 mg/m3 is noted, based on decreased body weight and renal histopathological changes in cats (same study considered by OEHHA).

USEPA PPRTV

RfD = 0.2 mg/kg/day Provisional peer reviewed toxicity value is available, the basis of which is not available.

OEHHA (OEHHA 2003)

RfD = 0.04 mg/kg/day Derived for the purpose of establishing a drinking water guideline, based on a NOAEL of 40 mg/kg/day for kidney damage in cats and an uncertainty factor of 1000. The guidelines also considered carcinogenic effects using a non-threshold approach.

Ontario (Ontario 2007)

TC = 0.165 mg/m3 After reviewing additional toxicological information, the Ontario Ministry decided to derive air quality standards for 1,1-dichloroethane based on the rationale supporting the former RfC of 483 μg/m3 developed by the U.S. EPA. In a view of the possible carcinogenic effect of 1,1-dichloroethane, an additional margin of safety uncertainty factor of 3 has been applied to this RfC in order to limit exposure levels to this compound. As a result, a criterion of 165 μg/m3 is achieved (rounded up from 161 μg/m3).

Limited data is available for 1,1-dichloroethane, hence the only available published peer-reviewed values have been adopted. These are the oral RfD available from OEHHA (OEHHA 2003) and the inhalation value available from Ontario (Ontario 2007).


B11 1,1-Dichloroethene B11.1 General 1,1-Dichloroethene (vinylindene chloride, 1,1-dichloroethylene or 1,1-DCE) does not occur naturally. It is produced commercially by the dehydrochlorination of 1,1,2-trichloroethane in the presence of excess base or by thermal decomposition of methyl chloroform. 1,1-DCE is used as an intermediate in the production of hydrochloroflurocarbons (HCFCs), in the production of chloroacetyl chloride and in the production of latex and resins (also known as PVDC polymers). These polymers are produced as emulsion polymers, as solvent-soluble powders for coating applications, and as resins for extrusion and co-extrusion. PVDC co-polymers containing 79–90% 1,1-DCE are used to form moisture and vapour barrier coatings and films, with applications as food packaging products. PVDC co-polymers containing 10–70% 1,1-DCE are used to improve flame and ignition resistance properties. Residual 1,1-DCE in PVDC used for food packaging products typically ranges from 5 to <1 mg/kg, the limit of detection of the method. Other consumer products containing PVDC include PVDC-latex for carpet backing (<2 mg/kg residual 1,1-DCE), PVDC-latex for photographic film coating (<100 mg/kg residual 1,1-DCE), PVDC for flame retardant fibres for clothing and outdoor awnings (<100 mg/kg residual 1,1-DCE), and PVDC-fluorinated copolymers for application on textiles (<100 mg/kg residual 1,1-DCE). Further processing decreases the residual 1,1-DCE in the final consumer product (HSDB; WHO 2003c).

B11.2 Exposure 1,1-DCE can be found in the environment from release during manufacture and use all from breakdown of polyvinylindene (PVDC) products, all from the biotic or abiotic breakdown of 1,1,1-trichloroethane, tetrachloroethene, 1,1,2-trichloroethene and 1,1-dichloroethane.

The main route of exposure to 1,1-DCE is via inhalation of contaminated ambient air. Exposure may also occur from ingestion of food, contaminated drinking water and use of consumer products containing 1,1-DCE such as plastic wrap which contains residual polymers. Occupational exposure to 1,1-DCE may occur through inhalation and dermal contact (ATSDR 1994a; HSDB)..

If released to air, 1,1-DCE will exist solely as a vapour phase in the ambient atmosphere and will be degraded by reactions with photo chemically produced hydroxyl radicals with a half-life of approximately 16 hours. The major reaction products are formaldehyde, phosgene and hydroxyacetyl chloride. If released to soil, 1,1-DCE is expected to be highly mobile with volatilisation expected to be an important fate process. If released to water the chemical is not expected to adsorb to suspended solids or sediments with volatilisation expected to be an important process. 1,1-DCE is generally considered to be resistant to biodegradation. The potential for 1,1-DCE to bioaccumulate in aquatic organisms is low (ATSDR 1994a; HSDB).

B11.3 Background Exposures/Intake Limited data is available regarding environmental levels of 1,1-DCE in Australia. It is noted in the Australian Drinking Water Guidelines (NHMRC 2011 Updated 2016) that DCEs have not been found in Australian drinking waters. Data available from the WA (WA DEP 2000) indicates 1,1-dichloroethane was rarely detected in air during sampling. Where detected in Duncraig (WA), the maximum concentration was reported to be 0.2 ppbV (0.00079 mg/m3). Data available from the


NSW (NSW DEC 2004) indicates 1,1-DCE was never detected in air during sampling undertaken from 1996 to 2001 in a number of urban and rural areas. This is well below the adopted threshold tolerable concentration in air adopted below. Data suggest (WHO 2003c) that the mean exposure from drinking-water will not exceed 6 to 9×10–5 mg/kg body weight per day for a 70-kg individual consuming 2 litres per day; oral exposure from food and soil is most likely negligible (WHO 2003c); and data suggest that the upper end of the range for the mean concentration of 1,1-DCE in air will not exceed 0.004 mg/m3 (consistent with data available from Australia). Thus, human exposure is expected to be far below the threshold levels adopted and are considered negligible.


General

1,1-DCE is rapidly absorbed following inhalation and oral exposure. Because of its low relative molecular mass and hydrophobic nature, dermal absorption is also likely, but there are no relevant published data. Although 1,1-DCE is rapidly distributed to all tissues, most of the free 1,1-DCE, its metabolites, and covalently bound derivatives are found in the liver and kidney. 1,1-DCE is rapidly oxidized to 1,1-dichloroethene oxide (DCE-epoxide), 2-chloroacetyl chloride, and 2,2-dichloroacetaldehyde. The major metabolites, DCE-epoxide and 2-chloroacetyl chloride, can react with glutathione (GSH), water, or tissue macromolecules. It is not known if the metabolism of 1,1-DCE is the same in humans, although in vitro microsomal preparations from human liver and lung form the same initial products (ATSDR 1994a; WHO 2003c).

The primary effect of acute exposure to high concentrations (approximately 4000 ppm) of 1,1-DCE vapour in humans is central nervous system (CNS) depression which may progress to unconsciousness. Occupational exposure has been reported to cause liver dysfunction in workers. 1,1-DCE is irritating when applied to the skin and prolonged contact can cause first degree burns. Direct contact with the eyes may cause conjunctivitis and transient corneal injury (ATSDR 1994a; WHO 2003c).


The following can be summarised 1,1-DCE (WHO 2003c):

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

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

Bioassays for cancer by the oral route of exposure have been conducted in rats, mice, and trout. Although these bioassays have protocol limitations, none provides any significant evidence that 1,1-DCE is a carcinogen by the oral route of exposure.

Bioassays for cancer by the inhalation route of exposure have been conducted in rats, mice, and hamsters. Most of these bioassays also have protocol limitations. One bioassay in male mice showed an increase in the incidence of kidney adenocarcinomas at one exposure level. There is evidence that the induction of kidney adenocarcinomas is a sex- and species-specific response. The results of the one bioassay showing an increase in tumours in one


sex and at one exposure level in a single rodent species are not sufficient to justify an exposure–response assessment.

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

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

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


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. In relation to 1,1-DCE the following is noted (USEPA 2002):

There are no studies that have demonstrated increased susceptibility of children to 1,1-DCE. Because of the role of CYP2E1 and GSH in the expression of toxicity of 1,1-DCE, individuals

with high levels of CYP2E1 (e.g., abusers of ethanol and individuals routinely exposed to ketones and heterocyclic compounds and other inducers of CYP2E1) could be more sensitive to the adverse effects of 1,1-DCE.

There is some evidence, however, that the rate of hepatic blood flow is an important limiting factor in the metabolism of 1,1-DCE. This effect would reduce the importance of the variability among individuals in concentration of CYP2E1 in the liver as a determinant of susceptibility to the adverse effects of 1,1-DCE. Individuals at risk from exposure to 1,1-DCE would also include those who have an extremely low level of GSH, for example, individuals who are malnourished or fasting or who are poisoned from acetaminophen.

Classification

1,1-DCE is classified by the US EPA as Group C possible human carcinogen based on no human data and limited animal data, in particular tumours observed in one mouse strain after inhalation exposure.

IARC has classified 1,1-DCE as Group 3 not classifiable as to human carcinogenicity based on limited evidence in animals (IARC 1999a).


B11.5 Quantitative Toxicity Values Review of available data with respect to 1,1-DCE indicates that the chemical is not considered to be genotoxic to humans (based on limited data) with limited data available with respect to carcinogenicity. On this basis it is considered relevant to consider potential exposures to 1,1-DCE on the basis of a threshold approach. The following quantitative values are available for 1,1-DCE from appropriate Australian and International sources:

Table B10 Summary of Published Toxicity Reference Values: 1,1-Dichloroethane


TDI = 0.009 mg/kg/day The Australian Drinking Water Guidelines have derived a drinking water guideline of 0.03 mg/L on the basis of a LOAEL of 9 mg/kg/day from a 2-year drinking water study in rats and a 1000 fold safety factor. This is the same as the previous evaluation provided by the WHO, prior to update in 2005, as noted below.


TDI = 0.046 mg/kg/day WHO derived a DWG of 0.14 mg/L based on a TDI of 0.046 mg/kg/day derived from a BMD10 of 4.6 mg/kg/day associated with hepatocellular mid-zonal fatty changes in female Sprague-Dawley rats and an uncertainty factor of 100. This update replaced the previous evaluation adopted in the ADWG (NHMRC 2011 Updated 2016).

WHO (WHO 2003c)


TDI derived on the same basis as noted in the WHO DWG (above). TC derived from a BMCLHEC of 6.9 mg/m3 associated with hepatocellular mid-zonal fatty changes in female Sprague-Dawley rats and an uncertainty factor of 30.

ATSDR (ATSDR 1994a)

Oral MRL = 0.009 mg/kg/day

Oral chronic MRL based on the same study and approach outlined in the ADWG, which has since been reviewed further been revised by the WHO.

USEPA (USEPA 2002)


Oral RfD is based on the same study and approach outlined by WHO (2003), just the threshold value has been rounded up from 0.046 to 0.05 mg/kg/day. Inhalation RfC derived on the same basis and approach as in the WHO (2003).

On the basis of the above the most current evaluations available from the WHO (WHO 2003c) which are consistent with that presented by the US EPA (USEPA 2002) has been adopted in this assessment.


B12 cis- and trans-1,2-Dichloroethene B12.1 General 1,2-Dichloroethene (also known as 1,2-DCE, 1,2-dichloroethylene, acetylene dichloride and dioform) exists in two isomeric forms, cis-1,2-dichloroethene and trans-1,2-dichloroethene, that are colourless, volatile liquids with a slightly acrid odour. 1,2-DCE is prepared commercially by either the direct chlorination of acetylene or by the reduction of 1,1,2,2-tetrachloroethane with fractional distillation used to separate the two isomers. 1,2-DCE can also be formed as a by-product during the manufacture of other chlorinated compounds. Commercial use is not extensive, but trans-1,2-DCE and mixtures of cis- and trans-1,2-DCE have been used as intermediates in the production of other chlorinated solvents and compounds, as well as low temperature extraction solvents for dyes, perfumes, and lacquers. Although not used extensively in industry, 1,2-DCE is used in the production of other chlorinated solvents and as a solvent for dyes, perfumes, and lacquers (ATSDR 1996; HSDB).

Both cis- and trans-1,2-DCE are colorless, volatile liquids with ethereal and slightly acrid odours. 1,2-DCE is slightly soluble in water, but is very soluble in alcohol, ether, acetone and most other organic solvents. Both forms are moderately flammable and react with alkalies to form chloracetylene gas, which spontaneously ignites in air. Additionally, cis- and trans-1,2-DCE react violently with potassium hydroxide, sodium, and sodium hydroxide and form shock-sensitive explosives when combined with dinitrogen tetraoxide (HSDB). 1,2-DCE emits chlorine gas when heated to decomposition (US DOE 1994).

B12.2 Exposure Because of its volatility, the primary route of 1,2-DCE exposure to humans is by inhalation, however exposure via oral or dermal routes may occur but are expected to be insignificant. Exposure may occur in the workplace (ATSDR 1996; HSDB).

In relation to dermal absorption, USEPA (USEPA 2004) notes that dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant. Region III guidance (USEPA 1995a) of dermal absorption for volatile compounds (http://www.epa.gov/reg3hscd/risk/human/info/solabsg2.htm) suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg), which would include DCE. For other volatiles (lower vapour pressure than benzene) dermal absorption is assumed to be 3%. No data is available specifically for the dermal absorption of DCE, hence a reasonable default value needs to be adopted. It is recommended that the dermal absorption value adopted for benzene (1%) be adopted for DCE. This is within the range of values recommended in the USEPA, but is more conservative than the default of 0% or 0.05%.

Cis-1,2-DCE may be released to the environment in emissions and wastewater during its production and use. Under anaerobic conditions that may exist in groundwater, landfills or sediment, 1,2-DCE can be formed as breakdown products from the reductive dehalogenation of trichloroethene, tetrachloroethene and 1,1,2,2-tetrachloroethane. The cis-1,2-DCE isomer is the more common isomer found although it is commonly mistakenly listed as the trans-isomer. The trans-isomer, is


more commonly analysed for and the analytical procedures generally used do not distinguish the isomers (ATSDR 1996; HSDB).

If released into the environment the following can be noted with respect to 1,2-DCE (ATSDR 1996):

Air: In the atmosphere 1,2-DCE will be lost by reaction with photochemically produced hydroxyl radicals (half life 8 days for cis-isomer and 3.6 days for trans-isomer) and scavenged by rain. Most of the 1,2-DEC removed by rain will probably re-enter via volatilisation. Because it is relatively long lived in the atmosphere, considerable dispersal from source areas should occur.

Soil: If 1,2-DCE (cis or trans) is released on soil, it should evaporate and/or leach into the groundwater where very slow biodegradation should occur. Adsorption of 1,2-DCE to soil, sediment or suspended solids in water is not a significant fate process.

Water: If released into water, 1,2-DCE (cis or trans) will be lost mainly through volatilization (half life 3 hr in a model river).

Biodegradation: Biodegradation, adsorption to sediment, and bioconcentration in aquatic organisms should not be significant. In groundwater 1,2-DCE may undergo anaerobic biodegradation with a biodegradation half life of approximately 13-48 weeks. Aerobic biodegradation processes have also been observed. 1,2-DCE is commonly found in mixtures with other chlorinated solvents and hence half-lives can only be approximated.

1,2-DCE has a low tendency to bioconcentrate in aquatic or marine organisms.

B12.3 Background Exposures/Intake As DCE is highly volatile and not persistent, background intakes will be dominated by inhalation exposures. DCE is not considered to be a typical urban air contaminant and little data is available from data collected in Australian cities. Cis-1,2-DCE has been detected in VOC sampling from Perth (WA DEP 2000) with average concentrations of 0.2 ppb (0.8 μg/m3) and a maximum reported concentration of 2.1 ppb (8.3 μg/m3). These values were comparable to average concentrations reported in air in the US and used by RIVM (Baars et al. 2001) to estimate background intake of 1,2-DCE (both isomers) of approximately 0.13 μg/kg/day. Based on the TDI adopted for DCE, this intake is less than 5% which is essentially negligible.



General

The following information is available from WHO (WHO 2003b), USDOE (US DOE 1994) and ATSDR (ATSDR 1996). There is no clinical disease which is unique to 1,2-DCE toxicity.

1,2-DCE is rapidly absorbed by the lungs. Once absorbed, the chemical is metabolised by the liver to form dichloroethanol and dichloroacetic acid via the epoxide intermediate. Animal studies indicate that the metabolism of the cis-isomer occurs faster than that of the trans-isomer. As the cis and trans-isomers are lipid soluble of low molecular weights, they would be expected to be readily absorbed by the oral or dermal routes.

Toxicokinetic data are very limited for both human and animals exposures to 1,2-DCE. Although the compound is relatively lipophilic, there is no good evidence of accumulation in the liver, brain, kidney and adipose tissue. 1,2-DCE is likely to be metabolised to more hydrophilic by-products and therefore eliminated quickly by the kidney as metabolites.

Workers exposed to 1,2-DCE have been reported to suffer from drowsiness, dizziness, nausea, fatigue, and eye irritation. Acute and sub-chronic oral and inhalation animal studies of trans-1,2-DCE and acute inhalation animal studies of cis-1,2-DCE suggest that the liver is the primary target organ. The toxicity is expressed in increased activities of liver associated enzymes, fatty degeneration, and necrosis. Secondary target organs include the central nervous system and lung. No information is available concerning the chronic, developmental, or reproductive toxicity of cis-1,2-DCE or trans-1,2-DCE.


No studies are available regarding carcinogenicity.

Review of available genotoxicity studies by the WHO (WHO 2003b, 2011) provided equivocal results. Review by RIVM (Baars et al. 2001) suggested that cis-1,2-DCE could be considered genotoxic in vivo, producing gene mutations and chromosome aberrations. However, no carcinogenic toxicity values have been derived for the cis- isomer. A more recent review of genotoxicity provided by the USEPA (USEPA 2010a) suggested that overall, data for 1,2-DCE (both isomers) are not positive for genotoxicity an mutagenicity. The positive results (considered by RIVM) are considered inconsistent by the USEPA and need further confirmation.

Susceptible Populations

Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. Limited information is available in relation to 1,2-DCE, however the following is noted by the USEPA (USEPA 2010a):

On the basis of a limited study, trans-1,2-DCE is not expected to cause fetotoxicity or developmental effects in humans; however, the limited information does not support an


assessment of potential developmental toxicity. No studies were conducted that would address childhood susceptibility to either cis- or trans-1,2-DCE.

Acute toxicity studies in animals provide suggestive evidence that males may be more sensitive than females to either cis- or trans-1,2-DCE. However, conclusions about gender differences in response to 1,2-DCE exposure cannot be drawn based on this limited information.

Four specific enzymes have been associated with the metabolism of cis- or trans-1,2-DCE: CYP2E1, CYP3A4 (relevant in rats, not humans), ADH (not well understood), and GSTZ.

o CYP2E1 is constitutively expressed in human liver but is inducible by a variety of factors, prominently by ethanol consumption, diabetes, or hunger, with in vivo activity levels varying up to 20-fold. Therefore, it is possible that a portion of the population may experience susceptibility towards the toxic effects of cis- or trans-1,2-DCE. In addition, the possibility exists that polymorphism of the CYP2E1 gene may affect the susceptibility of humans to the toxic effects of cis- and/or trans-1,2-DCE.

o Although DCA is likely a minor metabolite of cis- and trans-1,2-DCE, it is considered a likely human carcinogen, and therefore genetic polymorphism of the enzyme that metabolizes DCA, GSTZ, may play a role in human susceptibility.

Classification

USEPA has placed both cis-1,2-DCE and trans-1,2-DCE as “inadequate information to assess the carcinogenic potential of cis- or trans-1,2-DCE”. This description reflects the lack of human epidemiological investigations or chronic animal bioassays (USEPA 2010a).

The International Agency for Research on Cancer (IARC) has not classified DCE.

1,2-DCE has not been evaluated by the IARC or the National Occupational Health and Safety Commission (NOHSC) or NICNAS with respect to carcinogenicity.

12.5 Quantitative Toxicity Values On the basis of the available information it is considered appropriate that a threshold dose-response approach be adopted for 1,2-DCE. Few quantitative toxicity values are available; however the following are available from relevant Australian and International sources:


Table B11 Summary of Published Toxicity Reference Values: 1,2-Dichloroethene


TDI = 0.017 mg/kg/day for trans- isomer

The Australian Drinking Water Guidelines have derived a drinking water guideline of 0.06 mg/L for 1,2-DCE (relevant to the sum of both isomers) following guidance from the WHO (refer below).

International WHO (WHO 2011)

TDI = 0.017 mg/kg/day for trans- isomer

The WHO DWG has derived a guideline of 0.05 mg/L based on a TDI of 0.017 mg/kg/day based on a NOAEL of 17 mg/kg from a 90 day study in mice administered trans-1,2-DCE in drinking water and an uncertainty factor of 1000. This guideline is relevant to the sum of both cis- and trans- isomers, however this is due to the WHO adopting a conservative approach where there is no data being available for the derivation of a cis- isomer value.

RIVM (Tiesjema & Baars 2009)

cis- and trans- TDI = 0.03 mg/kg/day TC = 0.06 mg/m3

A TDI of 0.03 mg/kg/day has been established based on a NOAEL of 32 mg/kg/day from a 90 day oral rat study (using the cis- isomer) and an uncertainty factor of 1000. TC based on a LOAEL of 780 mg/m3 (continuous exposure) derived from liver and lung effects from an inhalation study on rats (using the trans- isomer) and applying an uncertainty factor of 3000. Value is considered provisional due to a poor database.

ATSDR (ATSDR 1996)

No chronic MRLs derived No chronic values available, however acute and intermediate duration MRLs are available. Acute and intermediate inhalation MRL for trans- isomer is 0.8 mg/m3. Acute oral MRL for cis- isomer is 1 mg/kg/day while the intermediate oral MRL is 0.2 mg/kg/day for trans- and 0.3 mg/kg/day for cis-.

USEPA (USEPA 2010a)

RfD = 0.002 mg/kg/day for cis- isomer RfD = 0.02 mg/kg/day for trans- isomer

RfD for cis- derived on the basis of a BMDL10 of 5.1 mg/kg/day associated with increased kidney weight in male rats and a 3000 fold uncertainty factor (includes 3 fold factor for database deficiencies). RfD for trans- derived on the bases of a BMDL1SD of 65 mg/kg/day associated with immune system effects in a 90 day study in mice with an uncertainty factor of 3000.

While there is limited data, most of the evaluations provide similar toxicity reference values, particularly for oral exposures. The oral TRVs derived by the USEPA (USEPA 2010a) have been adopted for the assessment of all exposures. For inhalation exposures an inhalation criteria of 0.007 mg/m3 for cis- and 0.07 mg/m3 for trans- (similar to that derived by RIVM based on a study with the trans-isomer) have been calculated (assuming 70 kg body weight and inhalation of 20 m3/day).


B13 Vinyl Chloride B13.1 General Vinyl chloride (also known as, chloroethene, 1-chloroethylene, ethylene monochloride and vinyl chloride monomer and commonly referred to as VC) is man-made or results from the breakdown of other manmade substances, such as trichloroethene, trichloroethane, and tetrachloroethene. It is used mainly for the manufacture of polyvinyl chloride (PVC) plastics and resins, and VC copolymers. It is used as a monomer with vinyl acetate or vinylidene chloride in the production of resins. It is also used for the production of chlorinated hydrocarbons, such as methyl chloroform and 1,1,1-trichloroethane, and other chemicals; and in the production of adhesives. Other uses include furniture, automotive upholstery, wall coverings, house wares and automotive parts. Up until the mid-1970’s it was used as a coolant, propellant in spray cans and in some cosmetics (ATSDR 2006).

VC is a volatile, colourless gas with a pleasant, sweet, ethereal odour. It is a colourless liquid below -14oC. VC is slightly soluble in water and highly soluble in diethyl ether, soluble in ethanol, benzene and most organic liquids (HSDB).

B13.2 Exposure The main route of exposure for the general public to VC is via inhalation (since VC commonly exists as a gas). Atmospheric concentrations of VC are generally low resulting in very little exposure to the general public. Similarly, the main route of occupational exposure is via inhalation. Dermal exposure is generally considered to be low.

With respect to dermal absorption, as noted by the USEPA, dermal absorption of highly volatile chemicals is expected to be negligible as it is not expected to remain on the skin long enough for absorption to be significant (USEPA 2004). Region III guidance of dermal absorption for volatile compounds suggests that a value of 0.05% be adopted for benzene and other volatiles with a vapour pressure similar to or greater than benzene (VP = 95.2 mmHg), relevant to vinyl chloride (USEPA 1995a). No data is available regarding dermal absorption of vinyl chloride from soil, however studies associated with dermal absorption of vapour phase vinyl chloride show low levels of dermal absorption (0.023 – 0.031%) (ATSDR 2006). It is recommended that the dermal absorption value adopted for benzene (1%) be adopted for vinyl chloride. This is within the range of values recommended in (USEPA 1995a), but is more conservative than the default of 0% or 0.05%.

In some countries exposure may also occur via ingestion of contaminated drinking water. In Australia, there are stringent requirements on the maximum permissible residual VC concentrations in PVC pipes and fittings used to carry potable water. Hence VC has not been reported in Australian drinking waters (NHMRC 2011 Updated 2016).

In the past, VC had been detected in food that was stored in materials that contained PVC. Many countries now regulate the amount of VC in food packaging materials.

If released into the environment, the following can be noted with respect to VC:


Air: Reaction with photochemically produced OH* radicals is the dominant atmospheric transformation process, which results in half-lives of 1 to 4 days in the troposphere. Several critical compounds, such as chloroacetaldehyde, formaldehyde and formyl chloride, are generated during experimental reactions.

Soil: Volatilisation half-lives are approx. 0.2-0.5 days. VC has a low soil sorption potential and therefore a high mobility in soil. VC may leach through soil into groundwater where it may persist for years.

Water: With few exceptions, VC is not easily degraded. However under anaerobic conditions PCE and TCE can be intrinsically biodegraded to form DCE and VC.

B13.3 Background Exposures/Intake As vinyl chloride is highly volatile and not persistent, background intakes will be dominated by inhalation exposures. Concentrations of vinyl chloride in industrial, urban and regional areas are available in Australia. Data collected in NSW from urban and regional areas in NSW note that vinyl chloride was rarely detected (<1% of samples) with the maximum reported from the Sydney CBD of 0.3 ppbv (0.0008 mg/m3) (NSW DEC 2004). Vinyl chloride was not detected in ambient air sampling undertaken in Perth (WA DEP 2000). Vinyl chloride has not been detected in drinking water and low levels are expected in food. Low levels have been historically reported in some consumer products. Background intakes expected from vinyl chloride are expected to be low, with conservative intakes estimated by Health Canada (Health Canada 2013) of approximately 0.005 mg/kg/day and RIVM (Baars et al. 2001) of approximately 0.00006 mg/kg/day (predominantly from inhalation). It is noted that, as the most sensitive endpoint is carcinogenicity which is assessed on the basis of a non-threshold approach, hence background intakes are not used the quantification of an incremental lifetime cancer risk.


General

Numerous human population studies and reports have led to the identification of significant long-term health effects which are sometimes collectively referred to as “vinyl chloride disease” and characterised by Raynaud’s Phenomenon, acroosteolysis, joint and muscle pain, enhanced collagen deposits, stiffness of the hands and scleroderma-like skin changes. Most of these effects are associated with inhalation exposures in the workplace (particularly during the 1970’s) (ATSDR 2006).

Primary effects are associated with the liver/spleen, vascular, skeletal, immune system, skin, respiratory and higher central nervous system (CNS) effects. The liver is the most sensitive target organ for vinyl chloride toxicity for both intermediate and chronic duration in halation and chronic duration oral exposures. The sensitivity of the liver to vinyl chloride exposure is consistent with the proposed mechanism of action (metabolism). It is well recognised that VC is a genotoxic carcinogen (ATSDR 2006).

VC is rapidly and well absorbed after inhalation or oral exposure. The primary route of exposure to VC is inhalation. Dermal absorption of VC in the gaseous state is not significant. Following exposure VC is distributed rapidly throughout the body. Placental transfer of VC occurs (ATSDR 2006).


The main route of metabolism of VC after inhalation or oral uptake involves oxidation by cytochrome P-450 (CYP2E1) to form chloroethene oxide (CEO), a highly reactive, short-lived epoxide which rapidly rearranges to form chloroacetaldehyde (CAA). These metabolites have been shown to bind to DNA and hepatocellular proteins. Thus, the prevalence of a mixed function oxidases activity in the liver and resulting production of reactive metabolites results in the observed sensitivity of the liver to cancer and non-cancer effects (ATSDR 2006).

After inhalation or oral exposure to low doses, VC is metabolically eliminated and non-volatile metabolites are excreted mainly in the urine. CEO is thought to be the most important metabolite in vivo, concerning the mutagenic and carcinogenic effects of VC (ATSDR 2006).


Exposure to vinyl chloride via inhalation has been associated with increase in liver cancer including a rare form of angiosarcoma and biliary tract cancer. Other studies have indicated increase incidence of CNS and brain cancer. While most data is associated with inhalation exposures, ingestion studies suggest evidence of carcinogenicity via oral exposure (ATSDR 2006; WHO 1999b).

Vinyl chloride has been identified as genotoxic and mutagenic (ATSDR 2006; USEPA 2000; WHO 1999b). The USEPA (USEPA 2000) review notes that vinyl chloride toxicity occurs via a genotoxic pathway (identified from a number of lines of evidence) that is understood in some detail. On this basis the assessment of carcinogenicity on the basis of a non-threshold (linear) approach is appropriate.

Susceptible Populations

The USEPA (USEPA 2000) review notes that chemically induced human liver carcinogenicity is associated with mutational alteration of multiple genes, consistent with a mutagenic mode of action. In addition several studies of partial lifetime exposure suggest that the lifetime cancer risk depends on age at exposure, with higher lifetime risks attributable to exposures at younger ages (USEPA 2000). This is also noted by the WHO (WHO 1999b, 2000c). Consistent with USEPA guidance the derivation of non-threshold values for vinyl chloride by the USEPA has incorporated factors that address early life susceptibility and hence if the USEPA non-threshold values are adopted (also considered in the WHO values) no additional adjustment is required in the quantification of exposure

Classification

VC is classified as a known human carcinogen (Category A) by the USEPA based upon sufficient evidence from animal studies. VC is a known human carcinogen via the inhalation and oral routes of exposure and a highly likely carcinogen via the dermal route of exposure (USEPA 2000).

IARC has classified VC in Group 1 (carcinogenic to humans) based in sufficient evidence from animal studies (IARC 2012).

NICNAS (2003) has classified VC as a Carcinogen Category 1, which is a substance regarded as carcinogenic to humans.


B13.5 Quantitative Toxicity Values The most sensitive end-point for vinyl chloride is carcinogenicity (noting that in the derivation of the ADWG both carcinogenic and non-carcinogenic effects were considered as sensitive for the oral pathway). Hence the selection of appropriate non-threshold values for the assessment of VC exposure is relevant.

The following quantitative non-threshold values are available for vinyl chloride from appropriate Australian and International sources:

Table B12 Summary of Published Toxicity Reference Values: Vinyl Chloride


Adopted WHO non-threshold approach.

Current Guideline derived on the basis of the WHO non-threshold value and additional consideration of non-carcinogenic effects with a TDI of 0.00013 mg/kg/day


SF = 2.3 (mg/kg/day)-1 (for exposures from birth) SF = 1.15 (mg/kg/day)-1 (for exposures as adults)

WHO DWG derived on the basis of linear extrapolation from dose response data for all liver tumours from an oral exposure study in rats and assuming a doubling of the risk of exposure from birth (incorporating the 2-fold uncertainty identified by the USEPA (2000) review to address early life sensitivity. Exposures by workers (only adults) can be calculated on the basis of a slope factor that is 2 times lower.

WHO (WHO 2000c)

UR = 1x10-6 ( g/m3)-1 Inhalation unit risk derived on the basis of occupational exposures studies associated with haemangiosarcoma and a linear multistage model. The value derived is noted to be limited as it does not address early life sensitivity identified in newborn animals (relevant to exposures by children to 10 years).

Health Canada (Health Canada 2013)

SF = 0.45 (mg/kg/day)-1

Slope factor based on PBPK modelling based on liver cancer. Used the USEPA approach for considering early life stage risk that early life stage exposure resulted in double the risk as an adult.


SF = 0.17 (mg/kg/day)-1 UR = 2.8x10-5 ( g/m3)-1

Slope factor derived on the basis of hepatocellular carcinomas, angiosarcomas and neoplastic nodules in female rats as markers for carcinogenic response, and a linear extrapolation model. Inhalation unit risk derived on the basis liver effects in an inhalation study on female rats and mice and an extrapolation model. No consideration of early-life sensitivity was considered by RIVM. Threshold values were also derived for non-carcinogenic effects with a TDI = 0.0013 mg/kg/day which is based on the same study as considered in the ADWG, but with a less conservative uncertainty factor of 100. An inhalation TC = 0.056 mg/m3 was derived based on an inhalation study. RIVM notes that the carcinogenic endpoints are most sensitive.

ATSDR (ATSDR 2006)

No quantitative assessment of carcinogenic effects

ATSDR does not provide quantitative estimates of carcinogenic effects. However for non-carcinogenic effects a chronic oral MRL = 0.003 associated with non-neoplastic effects in livers from a chronic oral rat study was derived.

USEPA (USEPA 2000)

SF = 1.5 (mg/kg/day)-1 (for exposures over lifetime) SF = 0.75 (mg/kg/day)-1 (for exposures as adult)

Slope factor derived on the basis of hepatocellular carcinomas, angiosarcomas and neoplastic nodules in female rats as markers for carcinogenic response, a PBPK model to estimate human equivalent dose and linearised multistage model. Based on animal evidence of age-dependent sensitivity an additional 2-fold uncertainty has been included to address early-life sensitivity in exposures from birth.

USEPA (USEPA 2000)

UR = 8.8x10-6 ( g/m3)-1 for exposures over lifetime) UR = 4.4x10-6 ( g/m3)-1 for exposures as adult)

Inhalation unit risk derived on the basis liver angiosarcomas, angiomas, hepatomas and neoplastic nodules in an inhalation study on female rats and mice and an extrapolation model. Based on animal evidence of age-dependent sensitivity an additional 2-fold uncertainty has been included to address early-life sensitivity in exposures from birth. The USEPA review also identified threshold values for the assessment of


Source Value Basis/Comments non-carcinogenic effects with an oral RfD = 0.003 mg/kg/day (same as

derived by ATSDR) and an RfC = 0.1 mg/m3 based on route-extrapolation from the oral value.

Both the WHO and USEPA recognise age-sensitivity is important with respect to the assessment of exposure to vinyl chloride and hence it is appropriate to adopt toxicity values that take these issues into consideration. On this basis the non-threshold values available in the current WHO DWGs (oral) and presented by the USEPA (inhalation) have been adopted for the quantitative assessment of exposure (USEPA; WHO 2011).


B14 Hexachlorobenzene (HCB) B14.1 General Hexachlorobenzene (also known as perchlorobenzene, pentachlorophenyl chloride and commonly abbreviated to HCB) is a synthetic organic compound that does not naturally occur. HCB can be produced by reacting benzene with excess chlorine in the presence of ferric chloride at 150-200oC. HCB was primarily used as a grain fumigant on wheat, barley, oats and rye for the control of bunts. In most countries its use as a fungicide has been discontinued. HCB was also used in the production of pyrotechnic and military ordinance and in the manufacture of nitroso rubber for car tyres (ATSDR 2002; HSDB).

HCB is a clear white crystalline solid at room temperature that is practically insoluble in water. When heated to decomposition, it emits toxic fumes of chlorides. HCB is slightly soluble in ethanol, soluble in ethyl ether and very soluble in benzene.

HCB has also been incidentally produced as a by-product in the manufacture of chlorinated solvents such as tetrachloroethylene and carbon tetrachloride, chlorinated pesticide and other chlorinates compounds. Small amounts of HCB can also be produced during combustion processes such as burning of city wastes.

B14.2 Exposure Exposures to HCB may occur and in the workplace and within the general environment. Work place exposure may occur through inhalation and dermal contact with this compound at workplaces where HCB is produced or used (ATSDR 2002; HSDB).

The general population may be exposed to HCB via inhalation of ambient air, ingestion of food and drinking water. The general population is not likely to be exposed to large amounts of HCB, however trace amounts have been reported in food and air with HCB reported in most people tested for HCB or its metabolite (ATSDR 2002; HSDB).

HCB is readily absorbed by the oral route and poorly via the skin, with a dermal absorption factor of 0.1 (10%) (RAIS). Intake from dietary sources is estimated to be the most significant intake mechanism in the general population; however intake from ambient air or drinking-water may increase in areas closer to emission sources.

The results of most studies of the levels of HCB in foods and human tissues over time indicate that exposure of the general population to HCB declined from the 1970s to the mid-1990s in many locations. Infants may be exposed to HCB from their mother in utero or via human milk (WHO 1997).

If released into the environment the following can be noted with respect to HCB (ATSDR 2002; HSDB; WHO 1997):

Air: HCB is expected to exist in both the vapour and particulate-phase in the ambient atmosphere. Vapour-phase HCB is degraded in the atmosphere by reaction with photochemically-produced hydroxyl radicals with an estimated atmospheric half-life of about


2.6 years. Particulate-phase HCB may be physically removed from the air by wet and dry deposition. Due to its persistence in the troposphere HCB meets the criteria for long-range transport in the atmosphere.

Soil and Water: HCB is considered to be immobile in soils and sediment. Volatilisation of HCB from dry soil surfaces is not expected on the basis of the vapour pressure of the chemical. Volatilisation from moist soil surfaces may occur, however this process may be attenuated due to adsorption on to soil particles. HCB is not expected to biodegrade based on a measured half-life in soil of 3 to 6 years. In water, HCB is expected to adsorb to sediment or particulate matter. This compound may volatilise from the surface of water bodies, however adsorption may attenuate this process. The volatilisation half-life is estimated to be approximately 5 years if adsorption is considered.

Biodegradation is not expected in water on the basis of biodegradation half-lives which have been estimated to be in the order of several years (2 to 10 years) in fresh waters.

Bioconcentration in aquatic organisms is estimated to be very high on the basis of bioconcentration factors (BCF) in the range of 1,600 to 20,000 measured in fish.

On the basis of the potential for long-range transport, persistence in air, water, soil and sediment, bioaccumulation, toxicity and ecotoxicity, HCB meets the UN-ECE Persistent Organic Pollutant (POP) criteria (UNECE 1998b). HCB is listed under Schedule X within the National Strategy for the Management of Scheduled Waste. In addition, HCB has a National Waste Management Plan endorsed by ANZECC in 1996 (ANZECC 1996).

B14.3 Background Exposures/Intake For the general population, away from areas where significant stores of HCB waste have been identified (ANZECC 1996), background intakes would be expected to be primarily associated with residues in food. Food Standards Australia New Zealand have not detected HCB in any sample in the 18th, 19th or 20th food surveys (FSANZ 2003). WHO (WHO 1997) calculated that the total background intake by an adult of HCB is between 0.0004 and 0.003 μg/kg body weight per day mostly derived from dietary exposures. This intake is essentially negligible compared to the adopted TDI. Hence background intakes would be expected to be negligible. This is consistent with reviews of background intakes estimated by RIVM (Baars et al. 2001).


General

The following information is available from WHO (WHO 1997, 2011) and ATSDR (ATSDR 2002). There is no clinical disease which is unique to HCB toxicity.

There is a lack of toxicokinetic information for humans. HCB is readily absorbed by the oral route in experimental animals and poorly via the skin (there are no data concerning inhalation). In animals and humans, HCB accumulates in lipid-rich tissues, such as adipose tissue, adrenal cortex, bone marrow, skin and some endocrine tissues, and can be transferred to offspring both across the placenta and via mothers' milk. HCB undergoes limited metabolism, yielding pentachlorophenol, tetrachlorohydroquinone and pentachlorothiophenol as the major metabolites in urine. Elimination


half-lives for HCB range from approximately one month in rats and rabbits to 2 or 3 years in monkeys.

Acute toxicity of HCB is considered to be low via the oral and inhalation exposure pathways. In humans, toxicity has been observed following short-term repeated ingested exposure: with the liver, immune system, skin, thyroid and nervous systems the target organs of toxicity. In animals, similar effects have been noted. The most pronounced effect in both humans and animals is liver toxicity. HCB accumulates in the body over time.

Most data on the effects of HCB on humans originate from accidental poisonings that took place in Turkey in 1955-1959, in which more than 600 cases of porphyria cutanea tarda were identified from oral ingestion of HCB in bread. In this incident, disturbances in porphyrin metabolism, dermatological lesions, hyperpigmentation, hypertrichosis, enlarged liver, enlargement of the thyroid gland and lymph nodes, and (in roughly half the cases) osteoporosis or arthritis were observed, primarily in children. Breast-fed infants of mothers exposed to HCB in this incident developed a disorder called pembe yara (pink sore) and most died within a year. Animal studies have shown that HCB causes reproductive toxicity and increases the risk of cancer formation.

The primary systems for HCB are hepatic toxicity (porphyria), reproductive toxicity, developmental toxicity and carcinogenicity.


No association has been found between HCB levels in humans and the incidence of breast or other cancers. Several animal studies have demonstrated an increase in the incidence of tumour formation following oral exposure to HCB. Evidence of carcinogenicity is strongest in the liver (hyperplasia, benign tumours and malignant tumours). In addition HCB has been shown to induce renal metaplasia, adenomas and renal cell carcinomas; lymphosarcomas; adrenal hyperplasia; parathyroid adenomas and hemangioendothelioma and thyroid tumours (WHO 1997, 2011).

HCB has little capability to induce directly gene mutation, chromosomal damage and DNA repair. It exhibited weak mutagenic activity in a small number of the available studies on bacteria and yeast, although it should be noted that each of these studies has limitations. There is also some evidence of low-level binding to DNA in vitro and in vivo, but at levels well below those expected for genotoxic carcinogens.

On the basis of available metabolic and toxicological information the WHO (WHO 1997, 2011) considered that a threshold/ TDI approach was appropriate for the assessment of non-neoplastic effects and neoplastic effects.


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. Limited information is available in relation to HCB, however the following is noted from ATSDR (ATSDR 2002):


Infants and young children appeared to be especially sensitive to the effects of HCB in the Turkish grain poisoning epidemic. During the epidemic, there was an extremely high rate of mortality (close to 100% in some villages) in breast fed infants (under 2 years of age) of mothers known to have ingested the contaminated bread. No quantitation of exposure (dose and duration) was presented in any of these clinical reports. However, an estimated dose of 0.05–0.2 g/day (0.7–2.9 mg/kg/day for a 70-kg person) is considered to be reliable by the original investigators of the Turkey epidemic. 20-30 years follow-up studies identified a number of adverse effects related to the developmental and reproductive toxicity of HCB.

Animal studies have also confirmed that the developing organism is an especially sensitive target for HCB. Animal studies have also shown that HCB mediates toxicity through the neuroendocrine axis, with multiple effects on the thyroid gland (hypothyroidism), parathyroid gland (hyperparathyroidism), adrenal gland, and ovaries. Because the hormones produced by these endocrine organs play a crucial role in growth and development of the organism, it is not surprising that HCB interferes with these processes.

Classification

HCB (USEPA) has been classified as a "probable" human carcinogen (Category B2) by the USEPA on the basis of the induction of tumours in the liver, thyroid and kidney in three rodent species following oral exposure.

IARC (IARC 2001) has classified HCB in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans and limited evidence in experimental animals for carcinogenicity.

The National Occupational Health and Safety Commission (NOHSC) not evaluated HCB. NICNAS has classified not classified HCB.

B14.5 Quantitative Toxicity Values HCB has been associated with carcinogenic effects however the mode of action is of prime importance for determining the most appropriate dose-response approach for assessing risk. The available data (Baars et al. 2001; WHO 1997, 2011) shows that the weight-of-evidence does not suggest that HCB is genotoxic and hence a threshold approach is considered appropriate. Further review of HCB by the WHO (WHO 1997, 2011) considered the derivation of a threshold TDI based on both non-neoplastic and neoplastic effects (similar threshold values), hence appropriate threshold values are available that adequately address neoplastic (carcinogenic) effects.

The following are available from relevant Australian and International sources:


Table B13 Summary of Published Toxicity Reference Values: Hexachlorobenzene

Source Value Basis/Comments Australian – No evaluations available International JMPR (JMPR 1974)

ADI =0.0006 mg/kg/day An ADI of 0.0006 mg/kg/day was estimated for HCB by JMPR in 1969 (reaffirmed in 1974) based on a dietary study in rats.

WHO DWG (WHO 2011)

TDI = 0.00016 mg/kg/day No guideline has been established by WHO as concentrations of HCB occur in drinking water well below those of health concern. However the review conducted (in 2003) has indicated a health based guideline of 0.0005 to 0.001 mg/L can be derived on the basis of a linear multistage low-dose extrapolation model or a TDI of 0.00016 mg/kg/day.

WHO (WHO 1997)

TDI = 0.00016 mg/kg/day TDI presented is the lowest TDI derived on the basis of non-neoplastic and neoplastic effects. A TDI of 0.00017 mg/kg/day was derived on the basis of a NOEL of 0.05 mg/kg/day associated hepatic effects in pigs and rats and an uncertainty factor of 200 (the evaluation presented included consideration of the study used by ATSDR in the derivation of the chronic MRL). The lower TDI presented was derived on the basis of a Tumorigenic Dose (TD5) associated with a 5% excess incidence in tumours in experimental animals. The TDI was derived on the basis of results from a 2-generation study in rats where the TD5 values ranged from 0.81 mg/kg/day to 2.01 mg/kg/day. The lower value associated with neoplastic effects in the liver was adopted with an uncertainty factor of 5000 to derive the TDI of 0.00016 mg/kg/day. It is noted that review also derived a TDI for non-neoplastic effects of 0.00017 mg/kg/day based on a NOEL of 0.05 mg/kg/day based on hepatic effects in a subchronic study in pigs and a chronic study in rats. This evaluation considered the same studies as ATSDR and US EPA and the non-neoplastic TDI is similar to that derived for neoplastic effects.

Health Canada (Health Canada 1993a)

TDI = 0.0005 mg/kg/day TDI derived on the basis of effects on the liver in a subchronic dietary pig and rat study where a NOEL of 0.05 mg/kg/day was determined and a 100 fold uncertainty factor applied.

ATSDR (ATSDR 2002)


Chronic oral MRL derived on the basis of a LOAEL of 0.016 mg/kg/day for peribiliary lymphocytosis and fibrosis of the liver in a 2-generation study on rats and an uncertainty factor of 300.

ATSDR (ATSDR 2013)


Draft evaluation provides a chronic MRL based on a LOAEL of 0.022 mg/kg/day associated with liver effects in rats and a 300 fold uncertainty factor.

USEPA (USEPA)


Oral RfD (last reviewed in 1988) based on a NOAEL of 0.08 mg/kg/day associated with liver effects in a rat study and an uncertainty factor of 100. The USEPA has also derived non-threshold oral and inhalation values which are not presented here as they are not considered relevant.

Based on the available reviews presented above, a range of oral criteria are available. The oral value adopted by the WHO (WHO 1997) is relevant to both neoplastic and non-neoplastic effects (note the ATSDR and USEPA threshold values relate to non-neoplastic effects only) and is therefore considered suitable and appropriate.

No inhalation data, or evaluations, are available. It has therefore been assumed that the oral toxicity reference value is relevant for all pathways of exposure, with an inhalation value of 0.00056 mg/m3 derived on the basis of a 70 kg body weight and inhalation of 20 m3/day.


B15 Hexachlorobutadiene (HCBD) B15.1 General Hexachlorobutadiene (also known as perchlorobutadiene; 1,1,2,3,4,4-hexachloro-1,3-butadiene; 1,3-hexachlorobutadiene; dolen-pur; GP-40-66:120 and commonly abbreviated to HCBD) is a synthetic organic compound that does not naturally occur. HCBD is used as an intermediate in the production of rubber compounds. It is also used as a solvent, a fluid for gyroscopes, a heat transfer fluid, hydraulic fluid and has been used as a fumigant. HCBD has also been used as a means of recovering chlorine containing gas (snift) in chloride production plants. It is a by-product in the manufacture of chlorinated solvents such as PCE and carbon tetrachloride (ATSDR 1994b; HSDB).

HCBD is a colourless, oily liquid at room temperature with a turpentine like, pungent odour. HCBD is non-flammable, non-combustible, poorly soluble in water but miscible with ethanol and ether (ATSDR 1994b; HSDB).

B15.2 Exposure Exposure of the general population to HCDB may by inhalation, oral or dermal routes. Exposure is most likely to occur in occupational environments which handle or produce the chemical. Other environmental exposures may be associated with inhalation, ingestion of HCBD in drinking water or ingestion of fish or other foods. HCBD has not been found in Australian drinking waters (NHMRC 2011 Updated 2016).

HCBD is readily absorbed by the oral route and poorly via the skin, with a dermal absorption factor of 0.1 (10%)(RAIS).

If released into the environment the following can be noted with respect to HCBD (UNECE 2002; WHO 1994b):

Air: Inter-compartmental transport of HCBD will occur by volatilisation (limited), adsorption to particulate matter, and subsequent deposition or sedimentation. In addition to deposition, reaction with hydroxyl radicals is assumed to be an important sink of HCBD in the troposphere with an estimated atmospheric half-life of up to 2.3 years. Due to its persistence in the troposphere HCBD meets the criteria for long-range transport in the atmosphere.

Soil and Water: HCBD is expected to bind with soil and sediments. In water, HCBD is considered persistent unless there is high turbulence. Information available on the persistence of HCBD in water, sediment and soil shows conflicting results, however expert judgement has identified HCBD as persistent. Half-lives in natural waters and soils have been reported to be 4-52 and 4-26 weeks respectively. There is conflicting data available about biodegradation. Based on the Structure of HCBD it can be expected that dechlorination is required before aerobic biodegradation can occur. Model calculations indicate that HCBD does not biodegrade fast.

HCBD has a high bioaccumulation potential as has been confirmed by both laboratory and field observations. Average steady-state bioconcentration factors of 5800 and 17 000, based


on wet weight, have been determined experimentally in rainbow trout. Biomagnification has not been observed either in the laboratory or in the field.

HCBD is not listed as a key persistent organic pollutant under the Stockholm Convention. However, on the basis of the potential for long-range transport, persistence in water, soil and sediment, bioaccumulation, toxicity and ecotoxicity, HCBD meets the UN-ECE Persistent Organic Pollutant (POP) criteria (UNECE 2002). Further review of the potential persistence and bioaccumulation potential of HCBD has been undertaken (WCC 2005) that indicates that the persistence and bioaccumulation potential of HCBD does not necessarily meet the requirements as set out by the UNECE. However, while further data is obtained and debate continued, HCBD can be considered persistent in the environment and has the potential for bioaccumulation in the food chain.

B15.3 Background Exposures/Intake No data is available regarding environmental levels of HCBD in Australia, other than noting that HCBD has not been found in drinking water in Australia (NHMRC 2011 Updated 2016). HCBD is not a common urban air contaminant and as such background intakes of HCBD are considered to be negligible. Intakes of HCBD estimated by Health Canada (Health Canada 2000) suggested background intakes (from air, water and food) comprised approximately 5% to 20% of the TDI, however the evaluation was based on limited data with few detections, hence it is not considered reliable. On this basis, the assessment of risk associated with potential intake of HCBD does not need to be adjusted account for background unless other sources of HCBD are present in the study area. These are addressed in the assessment itself; hence no further consideration is warranted here.


General

The following information is available from WHO (WHO 1994b) and ATSDR (ATSDR 1994b). There is no clinical disease which is unique to HCBD toxicity. As there have been very few human studies, the evaluation of toxicity is mainly based on studies in experimental animals. However, limited human in vitro data suggest that the metabolism of HCBD in humans is similar to that observed in animals.

There is limited data available on the absorption of HCBD in animals. Oral experiments indicate that HCBD absorption is rapid and complete, however little data are available concerning absorption following dermal and inhalation exposures.

When orally administered, HCBD or its metabolites have been observed to be distributed primarily in the kidney (outer medulla) and adipose tissue. HCBD has also been located in the liver following intraperitoneal administration. HCBD and its metabolites are excreted in exhaled air, urine and faeces.

HCBD vapour is considered to be irritating to the mucous membranes of humans, and the liquid is corrosive. The compound should also be regarded a sensitising agent.


The main target organs for toxicity are the kidney and, to a much lesser extent, the liver. Reduced birth weight and neonatal weight gain has only been observed at maternally toxic doses, as was developmental toxicity.

Biotransformation to a reactive sulphur containing metabolite probably accounts for the observed nephrotoxicity, genotoxicity and carcinogenicity.


There is limited evidence for carcinogenicity in animals and insufficient evidence in humans. Review of carcinogenicity (OEHHA 2000a) indicated that there is sufficient evidence for the carcinogenicity of HCBD, based on the development of renal tubular neoplasms in rats. Review of HCBD by the WHO (WHO 2011) also notes the development of kidney tumours in a long-term oral study in rats. HCBD has not been shown to be carcinogenic by other routes of exposure. On the basis of available metabolic and toxicological information the WHO considered that a TDI approach was appropriate for the derivation of an oral drinking water guideline.

HCBD has been found to induce gene mutations, chromosomal aberrations, increased sister chromatid exchanges and unscheduled DNA synthesis, although some studies have reported negative results. OEHHA undertook a review of HCBD (OEHHA 2000a). The review suggested that HCBD and its metabolites were genotoxic in bacteria and mammalian cell cultures. The studies on pharmacokinetics and metabolism suggest that the carcinogenicity and genotoxicity of HCBD is primarily due to metabolism through a pathway that includes GSH, enzymes of the mercapturate pathway, deacylase activity and -lyase activity. Both trichloroethene (TCE) and tetrachloroethene (PCE) have been shown to induce renal tubular neoplasms in long-term studies in rats. Similarities in the site and type of tumour induced in rodents, the genotoxic activity in short-term test systems, and the metabolism of HCBD, TCE and PCE suggest that the three chlorinated alkenes may share a common mechanism of action in the induction of kidney tumours. However, oxidative metabolism via CYP-dependent monooxygenases appears to play a greater role in the bioactivation of TCE and PCE than HCBD.

Health Canada (Health Canada 2000) indicates that both genotoxic and non-genotoxic steps may be involved in the induction of tumours by HCBD in laboratory animals. Health Canada noted that observations in the single adequate carcinogenesis bioassay identified that tumours occur only at doses greater than those that induce non-neoplastic effects in the kidney. These degenerative effects and resulting regeneration are likely requisite in the induction of tumours and are considered, therefore, to be the critical endpoint.

On this basis it is considered appropriate that a threshold approach is adopted.


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. Limited information is available in relation to HCBD, however the following is noted from USEPA (USEPA 2003):


The primary target organ for HCBD is the kidney. Individuals with pre-existing kidney damage may represent a potentially sensitive subpopulation for HCBD health effects.

Studies in animals showed that the young rats and mice were more sensitive to the acute effects of oral HCBD than adults. Those data may suggest that infants may potentially be more susceptible to HCBD toxicity.

Classification

HCBD has been classified as a "possible" human carcinogen (Category C) by the USEPA (USEPA).

IARC (IARC 1999c) has classified HCBD in Group 3 (not classifiable as to its carcinogenicity to humans) based on inadequate evidence in humans and limited evidence in experimental animals for carcinogenicity.

The National Occupational Health and Safety Commission (NOHSC) as Category 3 carcinogen (possibly carcinogenic to humans). NICNAS has classified not classified HCBD.

B15.5 Quantitative Toxicity Values On the basis of the available information a threshold approach is considered relevant for the assessment of HCBD. The following threshold values are available from relevant Australian and International sources:

Table B14 Summary of Published Toxicity Reference Values: Hexachlorobutadiene


TDI =0.0002 mg/kg/day The Australian Drinking Water Guidelines have derived a drinking water guideline of 0.0007 mg/L for HCBD using a TDI of 0.0002 mg/kg/day based on a NOAEL of 0.2 mg/kg/day based on renal effects in rats and a 1000 fold safety factor. An additional 10 fold safety factor has been included to address potential carcinogenic effects and genotoxicity of some metabolites.


TDI = 0.0002 mg/kg/day Value derived on the same basis as above in the ADWG.

Health Canada (Health Canada 2000)

TDI = 0.00034 mg/kg/day TDI derived on the basis of a BMD of 0.034 mg/kg/day for effects on the kidney in a mouse study and a 100 fold uncertainty factor. It is noted that the review considered that non-neoplastic renal effects in animals is critical and appropriate for the derivation of a guideline.

ATSDR (ATSDR 1994b)

No chronic MRL, however 0.0002 mg/kg/day considered protective of chronic exposures

No chronic MRL is available however a intermediate duration oral MRL has been derived and is 0.0002 mg/kg/day using a LOAEL of 0.2 mg/kg-day based on the presence of kidney damage in female mice (NTP, 1991). ATSDR did not derive a chronic duration MRL because no data were located on the effects of chronic exposure in humans, and because the intermediate-duration MRL protects against chronic exposures (based on effects seen in chronic animal studies).

USEPA (USEPA)

NA The USEPA (available from IRIS) has derived an oral slope factor of 0.078 (mg/kg/day)-1 for HCBD based on a linear multistage model based on renal tubular adenomas and adenocarcinomas in rats; and an inhalation unit risk of 2.2x10-5 ( g/m3)-1 using a linear multistage model based on oral data used to derive the oral slope factor. The USEPA does not present any data relevant to the assessment of non-carcinogenic effects for HCBD. An oral reference dose of 0.0002 mg/kg/day was derived by the USEPA, however it was withdrawn in 1993.


On the basis of the available information the oral TDI presented in the current ADWG (NHMRC 2011 Updated 2016) is consistent with other evaluations and has been adopted in this assessment for the assessment of oral and dermal exposures. It is noted that use of this threshold value provides an evaluation that is a sensitive as the use of non-threshold values from the USEPA. Hence the threshold value adopted is considered adequately protective of all health endpoints.

No inhalation data, or evaluations, are available. It has therefore been assumed that the oral toxicity reference value is relevant for all pathways of exposure, with an inhalation value of 0.0007 mg/m3 derived on the basis of a 70 kg body weight and inhalation of 20 m3/day.


B16 Hexachloroethane (HCE) B16.1 General Hexachloroethane (also known as perchloroethane; carbonhexachloride, 1,1,1,2,2,2-hexachloroethane; hexachloroethylene, phenohep and commonly abbreviated to HCE) is a synthetic chlorinated alkane compound that does not naturally occur. HCE is not produced in large quantities; however it is used in a number of industrial processes and products. Common uses include the manufacture of smoke candles and grenades, as a plasticiser for cellulose esters, as a minor component of rubber and insecticide formulations, as a moth repellent, retardant in industrial fermentation processes, camphor substitute in nitrocellulose solvent, flame retardant and in the manufacture of aluminium alloys. Historically HCE was also used as an antehelminthic agent (flukicide) in veterinary medicine. HCE can be released into the environment as a by-product of industrial processes, chlorination of drinking water (very small quantities) or incineration of chlorinated waste materials such as polyvinyl chloride (PVC) (ATSDR 1997a; HSDB).

HCE is a colourless, non-flammable solid (rhombic crystals) at room temperature with a camphor-like odour. HCE is insoluble in water, bur soluble in ethanol, diethyl ether, chloroform, benzene and oils (ATSDR 1997a; HSDB).

B16.2 Exposure The primary routes of potential human exposure to HCE are through inhalation or drinking contaminated water. Occupational exposure of workers in industrial facilities manufacturing or using HCE may occur through inhalation or dermal absorption. HCE exposure to the general public is expected to be relatively low (ATSDR 1997a; HSDB).

HCE is readily absorbed by the oral route and poorly via the skin, with a dermal absorption factor of 0.1 (10%) (RAIS).

If released into the environment the following can be noted with respect to HCE (ATSDR 1997a; HSDB):

Air: HCE is quite stable in air and hence atmospheric transport of HCE may occur. HCE is expected to diffuse slowly with a half-life of about 30 years (noted within the stratosphere) and 73 years (within the troposphere). Deposition of HCE from air to water, plants and soil has been reported. It is not expected to react with hydroxyl radicals or ozone or to degrade in the troposphere. Degradation by photolysis may occur.

Soil and Water: HCE released to water or soil may volatilise or adsorb to soil or sediments. Volatilisation appears to be the major removal mechanism from surface waters. HCE is expected to have a medium to low mobility in soil, hence leaching to groundwater may occur. Sorption of HCE to aquifer materials has been found to retard migration in groundwater. In aquatic systems moderate to slight adsorption to suspended solids and partitioning to sediments is likely. HCE is relatively resistant to degradation in aquatic environments. HCE may be reduced to tetrachloroethylene in the presence of sulphide and ferrous ions. HCE may biodegrade in soil but abiotic degradation processes are not


expected to be significant. HCE is biotransformed in soil under aerobic and anaerobic conditions, but it is more rapid in anaerobic soils.

HCE is expected to bioconcentrate to a moderate degree. However HCE appears to be rapidly metabolised and given low ambient levels of HCE in the environment, bioaccumulation and biomagnification in the food chain is not expected to be significant.

HCE is not listed by the UNEP as a persistent organic pollutant. However there is the potential for HCE to bioaccumulate and due to its persistence in the atmosphere there is the potential for long-range transport of HCE to occur.

B16.3 Background Exposures/Intake No data is available regarding environmental levels of HCE in Australia, other than noting that HCE is not a common urban air or water contaminant and as such background intakes of HCE are considered to be negligible. On this basis, the assessment of risk associated with potential intake of HCE does not need to be adjusted account for background.


General

The following information is available from ATSDR (ATSDR 1997a). There is no clinical disease which is unique to HCE toxicity. There is limited data available on the absorption of HCE. Once HCE is ingested it is rapidly absorbed. The degree of absorption following inhalation or dermal exposure is unknown.

Following absorption, HCE is rapidly distributed throughout the body via the systemic circulation. HCE is lipophilic and hence most partitions to fat. Degradation/detoxification of HCE is thought to occur primarily in the liver.

Acute Effects: HCE acts primarily as a central nervous system depressant (possibly resulting in mild paralysis) in humans acutely exposed to it and in high concentrations it causes narcosis. HCE is moderately irritating to the skin, mucous membranes, and liver in humans. Liver and kidney effects have been observed in animals acutely exposed to HCE by ingestion. Based on animal tests, HCE to have moderate acute toxicity from ingestion and low acute toxicity from dermal exposure.

Chronic Effects: No information is available on the chronic effects of HCE in humans. Animal studies have suggested that chronic inhalation exposure of animals to high concentrations resulted in neurobehavioral effects and chronically exposed to HCE by ingestion or gavage kidney effects have been observed.


No data are available on the carcinogenic effects of HCE in humans. Hepatocellular carcinomas were observed in mice following chronic oral exposure to HCE. An increased incidence of renal neoplasms was observed in orally-exposed male rats, but not in females.

Review of HCE in the Stage 2 assessment (Woodward-Clyde 1996) indicates that based on the available evidence, it does not appear reasonable to conclude that HCE is a human carcinogen.


The only evidence for carcinogenicity was observed in animals exposed to high doses of HCE and in nearly all cases these animals displayed evidence of nephrotoxicity. It is likely that the resulting tumours were due to cell proliferation that occurred in response to tissue injury rather than any mutagenic effect of HCE. No carcinogenic effects were seen at lower doses where tissue injury was not induced. Renal tumours observed in male rats were associated with hyaline droplet nephropathy which is thought to be specific to male rats, and is not observed in other species including humans. Hence the renal tumours observed have little relevance to humans. Available studies indicate HCE is not mutagenic or genotoxic. On this basis the most reasonable approach for evaluating HCE was considered to be to use the threshold approach.

Review of HCE by the USEPA (USEPA 2011b) considered that HCE is “likely to be carcinogenic to humans” by all routes of exposure based on data from oral cancer bioassays in F344/N rats and B6C3F1 mice (no new studies than considered in the Stage 2 assessment (Woodward-Clyde 1996)). No human data are available to assess the carcinogenic potential of HCE. The mechanistic data available for HCE is limited; and the mode(s) of carcinogenic action of HCE in the liver, kidney, and adrenal gland is unknown. However, there are data suggesting that the induction of kidney tumours in male rats involves the accumulation of α2u-globulin in the kidney and the induction of liver tumours in male and female mice may involve increased cytotoxicity, inflammation, and regenerative cell proliferation in the liver, respectively. HCE has been shown to be a promoter not an initiator. Review of genotoxicity by the USEPA (USEPA 2011b) indicated that the genetic activity profile for HCE is predominantly negative; however some positive findings have been reported.

While the USEPA review derived non-threshold values for HCE, there is insufficient supporting data to suggest that a non-threshold approach is appropriate. A non-genotoxic mechanism of carcinogenicity is also considered appropriate by Health Canada (Health Canada 2005b).


Variation in response among segments of the population may be due to age, genetics, and ethnicity, as well as to differences in lifestyle, nutrition, and disease status. These could be potential risk factors that play an important role in determining an individual’s susceptibility and sensitivity to chemical exposures. Limited information is available in relation to HCE, however the following is noted from the USEPA (USEPA 2011b):

No studies were located that addressed possible childhood susceptibility to HCE-induced toxicity or carcinogenicity.

Toxicity studies in rats indicate that male rats are more sensitive to HCE-induced nephrotoxicity than females. Evidence suggests that female rats are more sensitive to HCE-induced hepatotoxicity. The reasons for these sex-specific differences are unknown.

CYP450 enzymes are polymorphic in the human population. Polymorphisms result in CYP450 enzymes with variant catalytic activity for substrates such as HCE. This could potentially result in decreased HCE detoxification or increased HCE bioactivation.


Classification

HCE has been classified as “likely to be carcinogenic to humans” by all routes of exposure based on evidence of statistically significant increased incidences of multiple tumour types in male rats and both sexes of mice (USEPA 2011b)

IARC (IARC 1999c) has classified HCE in Group 2B (possibly carcinogenic to humans) based on inadequate evidence in humans and sufficient evidence in experimental animals for carcinogenicity.

The National Occupational Health and Safety Commission (NOHSC) classified HCE as Category 3 carcinogen (possibly carcinogenic to humans). NICNAS has not classified HCE.

B16.5 Quantitative Toxicity Values On the basis of the available information a threshold approach is considered relevant for the assessment of HCE. The following threshold values are available from relevant Australian and International sources:

Table B15 Summary of Published Toxicity Reference Values: Hexachloroethane

Source Value Basis/Comments Australian – No evaluations available International WHO NA No evaluations available Health Canada (Health Canada 2005b)

NA No TDI was derived, however a LOAEL for non-neoplastic effects (most critical) of 10 mg/kg/day associated with histopathological changes in the kidney in the male rat was identified. This was compared with background intakes of HCE where a 3200 fold margin of safety was identified.

ATSDR (ATSDR 1997a)


Chronic MRL based on atrophy and degeneration of renal tubes in rats and 100 fold uncertainty factor. An intermediate (0.01 mg/kg/day) and acute (1 mg/kg/day) MRL was also derived. Acute and intermediate inhalation MRLs were derived based on tremors in rats, both values being the same at 58 mg/m3.

USEPA (USEPA 2011b)


Oral RfD based on a BMDL10 of 0.728 mg/kg/day associated with kidney effects in rats and a 1000 fold uncertainty factor. RfC based on a NOAEL of 465 mg/m3 associated with effects on the nervous system (tremours and ruffled pelt) in rats, adjustment for a human equivalent dose and a 3000 fold uncertainty factor. The USEPA evaluation also derived a non-threshold slope factor.

On the basis of the available information the oral and inhalation values presented by the USEPA (USEPA 2011b) has been adopted in this assessment. It is noted that the oral RfD provides a more conservative assessment of health risks when compared with the use of the non-threshold slope factor.


B17 Mercury B17.1 General Mercury is a heavy metal which exists in three oxidation states: 0 (elemental), +1 (mercurous) and +2 (mercuric). As well as the common mercurous and mercuric inorganic salts, mercury can also bind covalently to at least one carbon atom. Thus the most commonly encountered exposures associated with mercury are with elemental mercury, inorganic mercuric compounds and methylmercury.

Mercury occurs naturally as a mineral is widely distributed by natural and anthropogenic processes. The most significant natural source of atmospheric mercury is the degassing of the Earth’s crust and oceans and emissions from volcanoes. Man-made sources such as mining, fossil fuel combustion and industrial emissions generally contribute less on a global scale, but more on a local scale. Wet and dry deposition to land and surface water result in mercury sorption to soil and sediments (ATSDR 1999; HSDB).

Uses of mercury include use in the electrical and chlor-alkali industry (lamps, batteries and as cathodes in the electrolysis of sodium chloride to produce caustic soda and chloride), industrial and domestic instruments, laboratory and medical instruments and dental amalgam (mixed in proportion of 1:1 with a silver-tin alloy).

B17.2 Properties Elemental mercury is a dense, silvery white metal which is liquid at room temperature, readily volatilises and is considered to be the predominant form of mercury in the atmosphere. Mercury compounds differ greatly in general properties and solubility. Due to the wide range in properties associated with the forms of mercury, key properties have not been listed here, however they are available in a number of published reviews (ATSDR 1999; WHO 2003a).

B17.3 Exposure Exposure of the general population to mercury may occur via inhalation, oral or dermal contact. Exposure to elemental mercury may occur in the workplace or home if mercury is spilled. Inorganic mercury compounds are found in some batteries, pharmaceuticals, ointments and herbal medicines. Exposure to inorganic mercury can occur via inhalation or ingestion. Methylmercury is most commonly found in fish, especially larger fish at the top of the food chain with exposure typically associated with ingestion.

Current literature indicates that mercury (Hg) in the environment, including groundwater, exhibits complex behaviour that affects both its mobility and potential toxicity. Mercury has a low solubility in water; however, it also has the potential to form multiple species in the environment, which can lead to increased total mercury concentrations in aqueous systems. The relative toxicity of mercury is also dependent on the form in which it occurs, which, in groundwater, is dependent on: biogeochemical processes; partitioning between solids, groundwater, and vapour; and complexation with dissolved organic and inorganic ligands. Redox, pH conditions, and groundwater composition are, consequently, all important components of determining the likely form, and therefore, potential fate of mercury in the environment.


On the basis of the potential for long-range transport, persistence in water, soil and sediment, bioaccumulation, toxicity and ecotoxicity, mercury is considered persistent and is addressed in the 1998 UN-ECE Convention on Long-Range Transboundary Air Pollution on Heavy Metals (UNECE 1998a). The United Nations Environment Programme (UNEP) Governing Council concluded, at its 22nd session in February 2003, after considering the key findings of the Global Mercury Assessment report, that there is sufficient evidence of significant global adverse impacts from mercury to warrant further international action to reduce the risks to humans and wildlife from the release of mercury to the environment. The UN Governing Council decided that national, regional and global actions should be initiated as soon as possible and urged all countries to adopt goals and take actions, as appropriate, to identify populations at risk and to reduce human-generated releases.

B17.4 Background Exposure/Intake

Elemental and Inorganic Mercury

Background intakes from food, water and air were listed in the documentation associated with the derivation of the current health investigation level (HIL) for soil (Imray & Neville 1996), with the total intake of mercury (derived from inorganic or elemental sources, both of which add to the body burden of mercury) estimated for a 2 year old child was 2.1 μg/day (50% of the adopted TI of 5 μg/day which was based on methylmercury rather than inorganic mercury). The most significant exposures were derived from dietary intakes and dental amalgams.

Review of current information from Australia indicates the following:

Mercury levels are reported in the 20th Australian Total Diet Survey (FSANZ 2003). Dietary intakes of total mercury (which includes organic mercury in seafood) ranged from 0.01 to 0.2 μg/kg/day for toddlers (aged 2 years). This is consistent with intakes reported in the more recent survey (FSANZ 2011).

Typical concentrations of mercury reported in drinking water in the ADWG (NHMRC 2011 Updated 2016) are less than 0.0001 mg/L, resulting in an intake (1 L/day and body weight of 15.5 kg) by toddlers of 0.0073 μg/kg/day.

Review (NHMRC 1999) of intakes associated with amalgam fillings in Australian children and adults (based on average number of fillings of 0.5 and 8 respectively) provides an reasonable estimate of daily mercury absorption per person of about 0.3 μg for children and 3.5 μg for adults. The estimate for children is expected to be conservative as the use of mercury dental amalgams is declining.

Based on the above, background intakes by young children may be up to 0.23 μg/kg/day from oral intakes (dietary, dental and water). This is slightly higher than estimated intakes of 0.1 μg/kg/day from the Netherlands (Baars et al. 2001) and 0.037 μg/kg/day from the UK (UK EA 2009) for a 20kg child. These intakes comprise approximately 40% of the recommended oral TRV.

Levels of inorganic mercury in air are not available for Australia with estimates from the WHO (2003) for mercury in air ranging from 10 to 20 ng/m3 from the US (no indication on


speciation between elemental an inorganic). These concentrations comprise up to 10% of the recommended inhalation TRV.

Methyl Mercury

Background intakes from food, water and air were listed in the documentation associated with the derivation of the current soil HIL (Imray & Neville 1996), with the total intake of methylmercury estimated to be 2.4 μg/day (approximately 50% of the adopted TI of 5 μg/day which was based on methylmercury). The most significant exposures were derived from dietary intakes of seafood.

Review of current information from Australia indicates the following:

Mercury levels are reported in the 20th Australian Total Diet Survey (FSANZ 2003). Dietary intakes of total mercury (which is dominated by organic mercury in seafood) ranged from 0.01 to 0.2 μg/kg/day for toddlers (aged 2 years). This is consistent with intakes reported in the more recent survey (FSANZ 2011).

The most recent review of methylmercury by JECFA (WHO 2004a) included a review of estimated dietary intakes from a number of countries. The review references previous total diet surveys (from 1992 and 1995) and indicates that the mean intake of methylmercury for the population is approximately 0.7 μg/kg/week. It is noted that the 95th percentile intake estimated exceeds the recommended PTWI adopted by JECFA (WHO 2004). This is a conservative estimate but it suggests intakes may be a significant proportion of the recommended PTWI.

Reviews of background intakes of methylmercury by the UK (UK EA 2009) and Netherlands (Baars et al. 2001) suggest intakes ranging from 8% to 20% of the adopted TDI (similar to the recommended TRV). Data from Australia suggests intakes may be higher and hence a value of 80% is recommended to address the potential for a significant proportion of the recommended oral TRV to be derived from background intakes.

B17.5 Health Effects The following information is available from UK (UK EA 2002, 2009) and ATSDR (1999).

Elemental Mercury (Hg0)

General Limited data is available concerning the absorption of elemental mercury. Inhaled mercury vapour by humans indicates approximately 80% of the vapour crosses the alveolar membranes into the blood. Ingested elemental mercury is poorly absorbed from the gastrointestinal tract (with approximately 0.01% absorbed, WHO 2003) unless there is an unusual delay in passage through the gastrointestinal tract or a gastrointestinal abnormality. This is partly due to the formation of sulfur laden compounds on the surface of the metal which prevents absorption. The processes of absorption in the gastrointestinal tract via sorption of mercury vapour (following partitioning in the GI tract to a vapour phase) have not been demonstrated in the available studies or case studies associated with accidental ingestion of elemental mercury. When evaluating exposures to elemental mercury, absorption following ingestion is too low to be of significance as the vapour inhalation pathway is of most importance.


Dermal absorption of mercury vapour is limited and may only contribute approximately 2.5% of absorbed mercury following inhalation exposures. No data are available concerning dermal absorption of liquid metallic mercury.

Absorbed mercury is lipophilic and rapidly distributed to all tissues and able to cross the blood-brain and foetal barriers easily. Mercury is oxidised in the red blood cells by catalase and hydrogen peroxide to divalent ionic mercury. Approximately 7-14% of inhaled mercury vapour is exhaled within a week after exposure. The rest of the elemental mercury is either excreted via sweat and saliva, or is excreted as mercuric mercury. Approximately 80% is excreted as mercuric mercury via faeces and urine. Half-life elimination is approximately 58 days.

Acute exposure to high concentrations of mercury vapour has been associated with chest pains, haemoptysis, breathlessness, cough and impaired lung function with the lung identified as the main target following acute exposure.

The central nervous system is generally the most sensitive indicator of toxicity of metallic mercury vapour. Data on neurotoxic effects are available from many occupation studies.

Chronic exposure to metallic mercury may result in kidney damage with occupational studies indicating an increased prevalence of proteinuria.

Carcinogenicity and Genotoxicity Both USEPA and IARC indicate that elemental mercury is not classifiable as to its human carcinogenicity. No adequate animal studies are available for elemental mercury and occupational studies have indicated conflicting results.

Inorganic Mercury Compounds

General Limited data is available concerning the absorption of inhaled mercury compounds; however it is expected to be determined by the size and solubility of the particles. Absorption of ingested inorganic mercury has been estimated to be approximately 5 to 10% with absorption be children greater than for adults.

Review of dermal absorption by New Zealand (MfE 2011) has noted that “Mercury reacts with skin proteins, and as a result penetration does not increase commensurably with increasing exposure concentration but rather approaches a plateau value. Mercury has a permeability coefficient in the order of 10–5 cm/h (Guy et al., 1999), which compares to permeability coefficients in the order of 10–4

cm/h for lead.” ATSDR (1999) note that absorption of mercurous salts in animals can occur through the skin, however no quantitative data are available, hence a default value of 0.1% has been adopted based on the lower end of the range for metals (USEPA 1995b).

The USEPA (USEPA 2004) has recommended the use of a gastrointestinal absorption factor (GAF) of 7% for inorganic mercury based on mercuric chloride and other soluble mercury salt studies used in the derivation of the oral RfD. The GAF is used to modify the oral toxicity reference value to a dermal value in accordance with the USEPA (2004) guidance provided.


Inorganic mercury compounds are rapidly distributed to all tissues following absorption. The fraction that crosses the blood-brain and foetal barriers is less than for elemental mercury due to poor lipid solubility. The major site of systemic deposition of inorganic mercury is the kidney. Most inorganic mercury is excreted in the urine or faeces.

Acute exposure to high concentrations of ingestion of inorganic mercury has been associated with gastrointestinal damage, cardiovascular damage, acute renal failure and shock.

The kidney is the critical organ associated with chronic exposure to inorganic mercury compounds. The mechanism for the end toxic effect on the kidney, namely autoimmune glomerulonephritis, is the same for inorganic mercury compounds and elemental mercury and results in a condition sometimes known as nephrotic syndrome.

There is some evidence that inorganic mercury may cause neurological effects, particularly associated with studies of mercuric chloride. Reproductive and developmental effects have been observed in rats given mercuric chloride.

Carcinogenicity and Genotoxicity IARC have considered inorganic mercury compounds not classifiable as to human carcinogenicity. The USEPA has classified mercuric chloride as a possible human carcinogen (Class C) based on increased incidence of squamous cell papillomas of the forestomach and marginally increased incidence of thyroid follicular cell adenomas and carcinomas from a long term oral studies in rats.

Carcinogenicity studies in experimental animals are available on mercuric chloride only where no carcinogenic effect was observed in mice or female rats, while marginal increases in the incidence of thyroid follicular adenomas and carcinomas and forestomach papillomas were observed in male rats exposed orally. Mercuric chloride binds to DNA and induces clastogenic effects in vitro; in vivo, where both positive and negative results have been reported, without a clear-cut explanation of the discrepancy. The overall weight of evidence is that mercuric chloride possesses weak genotoxic activity but does not cause point mutations (WHO 2011). The current US evaluation (USEPA) evaluation of mercuric chloride indicates that a linear low-dose extrapolation is not appropriate as kidney tumour seen in mice occurred at doses that were also nephrotoxic. On this basis, in accordance with Australian (enHealth 2012) guidance it is not considered appropriate that a non-threshold dose-response approach is adopted for the assessment of mercuric chloride.

Methylmercury

General Limited data are available concerning the absorption of inhaled methylmercury compounds, however studies on rats indicates rapid and almost complete absorption of inhaled methylmercury vapour. Ingested methylmercury is almost completely absorbed. ATSDR (ATSDR 1999) also noted no information was identified for absorption of methylmercury via dermal absorption. The UK (EA 2009) notes that dermal absorption of methylmercury is reported to be similar to that of inorganic mercury. Hence the value adopted for inorganic mercury has also been adopted for methylmercury.

Methylmercury is distributed via the blood to all tissues. It can cross into the brain and foetus. The major site of systemic deposition of methylmercury is the kidney. Hair levels are typically used as an


index of exposure to mercury and there is a proportional relationship between mercury intake, blood mercury and hair mercury. Methylmercury is converted to mercuric mercury in animals and humans, though less readily than for elemental mercury.

The key target of methylmercury in humans is the CNS, particularly the brain. Evidence from animal and human studies indicates that the embryo and foetus are more sensitive to methylmercury than adults.

Other effects associated with methylmercury include damage to other tissues and organs including the lung, cardiovascular system, liver and kidney. In animals, the most sensitive indicator of damage other than CNS effects, are renal effects.

Carcinogenicity and Genotoxicity USEPA and IARC have classified methylmercury as a possible human carcinogen (USEPA Class C and IARC Group 2B) on the basis of long term animal studies. Both agencies consider that the evidence for carcinogenicity of methylmercury in humans is inadequate. The USEPA have concluded that methylmercury is not a potent genotoxic agent. Methylmercury induced tumours in mice were considered likely to have a non-genotoxic origin and hence it, in accordance with Australian (enHealth 2012), it is not appropriate to evaluate potential health effects on the basis of a non-threshold dose-response approach.

B17.6 Quantitative Toxicity Values

General

Review of toxicological studies and risk assessments by several countries and international organisations have established levels of daily or weekly intakes of mercury that are estimated to be “safe” (refer to the WHO (UNEP 2008) review). That is, there is a threshold or reference level below which exposures/intakes are not associated with adverse effects. The WHO makes it clear in their assessment that these reference levels are not a clear dividing line between safe and unsafe. This is because they have incorporated a number of safety/uncertainty factors into their calculation of the reference level for mercury which means a slight exceedance of this value does not immediately result in adverse effects.

Elemental and Inorganic Mercury

On the basis of the available information in relation to elemental and inorganic mercury a threshold approach is consider appropriate based on the most sensitive effect associated with mercury exposure. The following threshold values are available from relevant Australian and International sources:


Table B16 Toxicity Reference Values for Inorganic and Elemental Mercury


NA Guideline established on the basis of methylmercury only

FSANZ (FSANZ 2011)

NA Value for total mercury referenced from JECFA 1989, based on methylmercury


TDI = 0.002 mg/kg/day The current WHO DWG (2011, consistent with the previous evaluation conducted in 2003) has derived a guideline of 0.006 mg/L based on a TDI of 0.002 mg/kg/day derived from a NOAEL of 0.23 mg/day associated with kidney effects in a 26-week study in rats and an uncertainty factor of 100. A similar TDI was derived on the basis of a LOAEL of 1.9 mg/kg/day associated with renal effects in a 2-year rat study and an uncertainty factor of 1000.

JECFA (JECFA 2011)

PTWI = 0.004 mg/kg (equivalent to PTDI = 0.0006 mg/kg/day)

Review of mercury by JECFA indicated that the predominant form of mercury indoors, other than fish and shellfish, is inorganic mercury and while data on speciation is limited the toxicological database on mercury (II) chloride was relevant for establishing a PTWI for foodborne inorganic mercury. A PTWI was established on the bases of a benchmark dose approach, where the BMDL10 of 0.06 mg/kg/day for relative kidney weight increases in male rates was considered as the point of departure. A 100 fold uncertainty factor was applied.

WHO (WHO 2000d)

TC = 0.001 mg/m3 TC or guideline value derived on the basis of a LOAEL derived from occupational studies on elemental vapour. The WHO note that “since cationic inorganic mercury is retained only half as much as the vapour, the guideline also protects against mild renal effects caused by cationic inorganic mercury”. “Present knowledge suggests, however, that effects of the immune system at lower exposures cannot be excluded”.

WHO (WHO 2003a)1


TDI derived for inorganic mercury as noted in the DWG above. A TC in air was also derived for elemental mercury in air (0.0002 mg/m3) associated with a LOAEL associated with CNS effects in workers exposed to elemental mercury. The evaluation provides a revision on the limited TC presented in the WHO (2000).

UK (UK EA 2009)


TDI referenced from the WHO (2003) and WHO DWG (2011). Inhalation value (covered to a does by the UK) based on the WHO (2003) value assumed to be relevant to inorganic mercury in air.



TDI for mercuric chloride derived on the same basis as WHO. TC derived on the same basis as ATSDR and WHO (2003).

ATSDR (ATSDR 1999)

Inh. MRL = 0.0002 mg/m3

No chronic duration MRLs have been derived for inorganic mercury. An intermediate duration (or sub-chronic) oral MRL of 0.002 mg/kg/day was derived. The chronic inhalation MRL for elemental mercury based on a LOAEL (HEC) of 0.0062 mg/m3 associated with CNS effects in workers and an uncertainty factor of 30.

USEPA (IRIS) RfD = 0.0003 mg/kg/day RfC = 0.0003 mg/m3

RfD (last reviewed in 1995) for inorganic mercury based on a LOAEL of 0.226 mg/kg/day associated with autoimmune effects in a subchronic rat feeding study and an uncertainty factor of 1000. RfC (last reviewed in 1995) for elemental mercury based on a LOAEL (HEC) of 0.009 mg/m3 associated with CNS effects in workers and an uncertainty factor of 30. A subchronic RfC is also available from HEAST (1995), which is equal to the chronic RfC.


Notes: 1 This document is an update of a former evaluation of inorganic mercury presented in the WHO EHC 118 (WHO 1991).

In this evaluation the WHO states that following review of a number of animal studies in relation to inorganic mercury, no “no-observed-adverse-effect-level” (NOAEL) could be determined. This is a reflection of the limitations in the available animal studies rather than because there is no safe dose. These studies typically only consider perhaps 3-4 different doses and depending on the spacing of the quantitative magnitude of these doses it may or may not be possible to ascertain a dose which could be a NOAEL as the lowest dose use in the study may have been too high resulting in some effects being observed at all the dose levels. Hence this is not a definitive statement in relation to the determination of whether or not there is a safe level of mercury exposure and certainly does not imply that the WHO evaluation has stated that the safe dose for mercury is zero. It is important to note that since the 1991 WHO evaluation there have been numerous more robust studies undertaken that have enabled a safe dose to be more reliably determined as outlined in this table.

The PTWI derived for inorganic mercury available from JECFA (2011) is considered to provide the most current review of the available studies in relation to exposure to inorganic mercury and has been adopted for the assessment of exposure to inorganic mercury, via all pathways of exposure.

Inhalation values for elemental mercury are derived from occupational studies associated with elemental mercury vapour. The more current review provided by WHO (2003), consistent with that adopted by UK (UK EA 2009), RIVM (Baars et al. 2001) and ATSDR (1999), has been adopted for the assessment of inhalation exposures to elemental mercury. Limited subchronic evaluations are available and hence the chronic TRV has been adopted for the assessment of sub-chronic exposures. As inhalation is the most significant pathway of exposure relevant to this form of mercury, no values have been adopted for oral and dermal exposures.

Methylmercury

Long-term exposure to methylmercury has induced renal tumours in mice, but only at doses at which significant nephropathy was also evident (WHO 2004a). Review by the USEPA (IRIS) concluded that methylmercury is not a potent genotoxic agent and that methylmercury induced tumours in mice were likely to have a non-genotoxic mode of action. On this basis a threshold approach is considered appropriate based on the most sensitive effect associated with methylmercury exposure. The following threshold values are available from relevant Australian and International sources:


Table B17 Toxicity Reference Values for Methylmercury


TDI = 0.00047 mg/kg/day

Current ADWG derived a guideline of 0.001 mg/L on the basis of a PTWI of 0.0033 mg/kg derived from the older JECFA evaluation (see below). The guideline was considered sufficient to be protective of pregnant women and nursing mothers.

FSANZ (FSANZ 2003, 2011)

PTWI = 0.003 mg/kg/week (PTDI = 0.00047 mg/kg/day)

Value for total mercury referenced from older JECFA 1989, based on methylmercury.


Not established for methylmercury

The current WHO DWG has derived a guideline for inorganic mercury in drinking water only.

JECFA (WHO 2004a)


The most current evaluation by JECFA derived a PTWI of 0.0016 mg/kg based on a steady state intake of 1.5 μg/kg/day (from review of mercury in hair and blood, a benchmark dose approach to assess the relationship between maternal hair concentrations and foetal neurotoxicity and a pharmacokinetic model) estimated to represent the exposure that would be expected to have no appreciable adverse effects on children and applying an uncertainty factor of 6.4. The PTWI was considered to be sufficient to protect developing foetuses, the most sensitive subpopulation identified. The previous evaluations by JECFA (WHO 2000b) identified a PTWI of 0.0033 mg/kg methylmercury based on review of oral intakes of mercury and hair and blood mercury levels. Subsequent review of the PTWI by JECFA identified that the value may not be adequately protective of foetuses and infants who are more sensitive than adults.

UK (UK EA 2009)


Value adopted is referenced from JECFA for all routes of exposure.


TDI = 0.0001 mg/kg/day

Derived on the basis of a NOAEL of 1.3 μg/kg/day for developmental effects in humans (and hair concentrations) and an uncertainty factor of 10.

ATSDR (1999) MRL = 0.0003 mg/kg/day

Chronic oral MRL derived on the basis of a NOAEL of 0.0013 mg/kg/day (adjusted) associated with CNS effects in humans (and hair concentrations) and an uncertainty factor of 4.5.

USEPA (IRIS) RfD = 0.0001 mg/kg/day

RfD (last reviewed in 2001) based on a BMD of 0.0009 to 0.0015 mg/kg/day (adjusted) based on CNS effects in humans (and blood concentrations) and an uncertainty factor of 10.

The PTWI derived for methylmercury from JECFA (WHO 2004) is considered to be based on the most recent detailed review of available studies in relation to exposure to methylmercury. The TRV established by JECFA is within the same range of values previously established by the US EPA and ATSDR and is recommended for use in the derivation of a soil HIL for methylmercury. No dermal or inhalation specific data are available and hence the PTWI is recommended to be adopted for all routes of exposure.

Limited subchronic evaluations are available and hence the chronic TRV has been adopted for the assessment of sub-chronic exposures.


B17.7 Summary The following provides a summary of the TRVs adopted for the assessment of potential exposure to mercury (elemental, inorganic and methyl) as well as relevant assumptions in relation to the proportion of the total intake that may occur from sources other than contamination (i.e. drinking water, diet, air, dental fillings).

Inorganic Mercury: Chronic Oral TRV (TRVO) = 0.0006 mg/kg/day (JECFA 2011) for all routes of exposure, Subchronic Oral TRV = 0.002 mg/kg/day (ATSDR 1999) for all routes of exposure Gastrointestinal absorption factor (GAF) = 0.07 (USEPA 2004) Dermal absorption factor, soil (DAF) = 0.001 (or 0.1%) (USEPA 1995b) Background intakes from other sources (as % of TRV):

BIO = 40% for oral and dermal intakes BIi = 10% for inhalation

Elemental Mercury: Inhalation TRV (TRVI) = 0.0002 mg/m3 (WHO 2003a) – relevant to the inhalation pathway only (other pathways not of significance for this form of mercury) for chronic and sub-chronic exposures. Background intakes from other sources (as % of TRV):

BIi = 10% for inhalation Methylmercury: Oral TRV (TRVO) = 0.00023 mg/kg/day (UK EA 2009; WHO 2004a) for all routes of exposure Dermal absorption factor, soil (DAF) = 0.001 (or 0.1%) (as for inorganic mercury) Gastrointestinal absorption factor (GAF) = 1 Background intakes from other sources (as % of TRV):

BIO = 20% for oral and dermal intakes BIi = 20% for inhalation


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Documents

Appendix B Toxicity Summaries - Orica