Priority Substances List Assessment Report for Hexachlorobutadiene

3.0 Assessment of "Toxic" under CEPA 1999

3.1 CEPA 1999 64(a): Environment

The environmental risk assessment of a PSL substance is based on the procedures outlined in Environment Canada (1997a). Analysis of exposure pathways and subsequent identification of sensitive receptors are used to select environmental assessment endpoints (e.g., adverse reproductive effects on sensitive fish species in a community). For each endpoint, a conservative Estimated Exposure Value (EEV) is selected and an Estimated No-Effects Value (ENEV) is determined by dividing a Critical Toxicity Value (CTV) by an application factor. A conservative (or hyperconservative) quotient (EEV/ENEV) is calculated for each of the assessment endpoints in order to determine whether there is potential ecological risk in Canada. If these quotients are less than one, it can be concluded that the substance poses no significant risk to the environment, and the risk assessment is completed. If, however, the quotient is greater than one for a particular assessment endpoint, then the risk assessment for that endpoint proceeds to an analysis where more realistic assumptions are used and the probability and magnitude of effects are considered. This latter approach involves a more thorough consideration of sources of variability and uncertainty in the risk analysis.

There are special concerns about persistent and bioaccumulative substances. Persistent substances can remain bioavailable for long periods of time, increasing the probability and the duration of potential exposure. Even extremely low concentrations of persistent and bioaccumulative substances can have adverse effects on organisms that are continually exposed to them over long periods of time. Substances that are subject to long-range transport are of particular concern because cold regions, such as the Canadian Arctic, can act as a sink for such contaminants. Because of these concerns, environmental assessments of persistent and bioaccumulative substances are more conservative than those for other substances. Persistent and bioaccumulative substances may be determined to be toxic if they have the potential to harm the environment or its biological diversity, even if this is known to occur only within limited geographical areas within Canada.

3.1.1 Assessment endpoints

Current Canadian sources of HCBD are minor but potentially numerous. They include possible releases in landfill leachates, releases during refuse combustion and releases as a by-product in the production of other chlorinated chemicals. The most significant point source of HCBD in Canada appears to have been the Cole Drain, which discharges into the St. Clair River at Sarnia, Ontario. Recent remediation activities have practically eliminated discharges from this source, but benthic organisms are still exposed to HCBD from prior emissions from the drain. There is no indication that biota in Canadian marine systems are exposed to HCBD. Concentrations of HCBD in air and soil in Canada are generally low. The assessment endpoints for the environmental assessment of HCBD are normal growth and reproduction in populations of freshwater pelagic and benthic organisms in Canada.

3.1.2 Environmental risk characterization

3.1.2.1 Pelagic organisms

Concentrations of HCBD in St. Clair River water have declined considerably since the mid-1980s. The conservative EEV for pelagic organisms is 0.0027 µg/L, the highest reported concentration of HCBD in the St. Clair River in 1994.

The most sensitive freshwater species reported is the fathead minnow, with a 28-day LOEC of 13 µg/L, based on survival and growth.

This value, 13 µg/L, is the conservative CTV for pelagic organisms. Dividing this CTV by a factor of 100 to account for uncertainty surrounding laboratory to field extrapolation and inter- and intraspecies differences in sensitivity gives an ENEV of 0.13 µg/L.

The conservative quotient is calculated by dividing the EEV of 0.0027 µg/L by the ENEV, as follows:

Because the conservative quotient is less than 1, this substance is unlikely to cause a harmful effect on populations of pelagic organisms in the ambient aquatic environment.

This quotient would be lower for freshwater invertebrates, since they appear to be somewhat less sensitive than fish to HCBD. The application factor of 100 used for deriving the ENEV is conservative, as the CTV was based on a 28-day LOEC, rather than a 96-hour LC₅₀.

The risk quotient for pelagic organisms is presented in Table 1.

Table 1 Risk quotient for pelagic organisms

Parameter	Value
EEV	0.0027 µg/L
CTV	13 µg/L
Application factor	100
ENEV	0.13 µg/L
Quotient (EEV/ENEV)	0.02

3.1.2.2 Benthic organisms

The conservative EEV for benthic organisms is 243 µg/g dry weight, the highest reported concentration of HCBD in the top 5 cm of sediment in a 2-km stretch of the St. Clair River in an industrialized zone near Sarnia, Ontario, in 1994.

The CTV for benthic organisms is 20.8 µg/g dry weight, estimated using the Equilibrium Partitioning approach as presented in Section 2.4.1.2 Dividing this CTV by a factor of 100 to account for the uncertainty surrounding the extrapolation from laboratory to field conditions and interspecies and intraspecies variations in sensitivity gives an ENEV of 0.21 µg/g dry weight.

The conservative quotient is calculated by dividing the EEV of 243 µg/g by the ENEV, as follows:

Since the conservative quotient is more than 1, it is necessary to consider further the exposure of benthic biota to HCBD in the St. Clair River.

Figure 2 The cumulative density function for HCBD in St. Clair River sediments (0-5cm)

The cumulative density function for HCBD in St. Clair River sediments, at a depth of 0-5 cm, is shown in Figure 2. As stated in Section 2.3.2.4, the 99th-, 95th- and 90th-percentile values are 194, 60.9 and 18.7 µg/g dry weight, respectively, while the median is 0.9 µg/g dry weight.

Risk quotients for benthic organisms at various exposure levels in St. Clair River sediments are presented in Table 2. The ENEV in this table is the same as that used in the conservative risk assessment, 0.21 µg/g dry weight.

As indicated in Table 2, a quotient exceeding 1 occurs frequently in the sediments in the St. Clair River near Sarnia, Ontario. In fact, the concentration of HCBD in sediments in this area equalled or exceeded the ENEV of 0.21 µg/g dry weight at 29 of 39 sample stations. Benthic organisms in highly contaminated locations within this 2-km stretch of the St. Clair River could experience adverse effects because of their inability to move to less contaminated areas.

The sediment in this section of the St. Clair River contains a wide variety of organic and inorganic contaminants, including mercury, polychlorinated biphenyls, polychlorinated aromatic hydrocarbons, petroleum hydrocarbons and hexachlorobenzene, along with HCBD (Bedard and Petro, 1997). Whole-sediment toxicity tests were conducted on three species - the mayfly, Hexagenia limbata (21-day mortality and growth), the midge, Chironomus tentans (10-day mortality and growth), and the fathead minnow (21-day mortality) - using sediment samples taken from the most contaminated area. Significant correlations were found between lethality and HCBD concentration. HCBD bulk sediment concentrations explained 94% of the variation in midge mortality and 54% of the variation in mayfly mortality (Bedard and Petro, 1997). These results support the conclusion that benthic organisms in the most contaminated part of the St. Clair River can be harmed by HCBD in the sediments.

Table 2 Summary of risk quotients for freshwater benthic organisms

EEV (µg/g dry weight)	Descriptor	CTV (µg/g)	Application factor	ENEV (µg/g)	Quotient (EEV/ENEV)
243	Maximum reported concentration, 1994	20.8	100	0.21	1157
194	99th percentile, 1994	20.8	100	0.21	924
60.9	95th percentile, 1994	20.8	100	0.21	290
18.7	90th percentile, 1994	20.8	100	0.21	89
0.9	Median, 1994	20.8	100	0.21	4.3

Because HCBD is persistent, with a half-life in air ranging from 60 days to 3 years and with a potential for long-range transport, as supported by measurements in Great Slave Lake sediments, and because the substance bioaccumulates, with a BCF ranging up to 19 000, a probabilistic risk assessment will not be performed. HCBD is still released to the environment in many sites, with concentrations in effluents up to 0.9 µg/L, compared with a pelagic ENEV of 0.13 µg/L.

3.1.2.3 Sources of uncertainty

There are several sources of uncertainty associated with the environmental assessment of HCBD. There were no acute or chronic toxicity studies using benthic organisms identified for HCBD. Effects on benthic organisms were therefore estimated using the Equilibrium Partitioning approach. This approach is based on the assumption that sediment interstitial water is the primary route of exposure of benthic organisms to HCBD, that continuous equilibrium exchange between sediment solids and interstitial water occurs, and that distribution of HCBD between these two phases can be estimated using the organic carbon/water partition coefficient of the substance and the organic carbon content of the sediment. Benthic organisms in highly contaminated areas of the St. Clair River at Sarnia, Ontario, may be adversely affected by HCBD, but the exact spatial extent of this area cannot be determined from existing data, because concentrations of the substance above the ENEV of 0.21 µg/g dry weight occurred at the sampling sites located farthest downstream. Concentrations of HCBD in sediments downstream from the source of contamination have been slowly declining since the mid-1980s.

3.1.2.4 Conclusion

The available information therefore indicates that HCBD poses little or no risk to pelagic aquatic organisms in Canada. HCBD poses a risk to benthic organisms in the most contaminated portions of the St. Clair River.

3.2 CEPA 1999 64(b): Environment upon which life depends

Worst-case calculations were made to determine if HCBD has the potential to contribute to depletion of stratospheric ozone, ground-level ozone formation or climate change. The Ozone Depletion Potential (ODP) was calculated to be 0.07, the POCP was estimated to be 0.01 and the GWP was calculated to be 0.037. These figures imply that HCBD is not likely to contribute significantly to ground-level ozone formation, but it does have the potential to contribute to depletion of stratospheric ozone and to climate change. Some substances currently subject to the Montreal Protocol have ODP values similar to the one calculated for HCBD; however, there is general agreement that at these ODP values, substances should not be automatically subject to controls. Other criteria, such as quantities emitted, also have to be taken into consideration. The concentration of HCBD in the Canadian atmosphere is low; estimates of its half-life in air based on photochemical degradation through reactions with hydroxyl radicals and ozone range from 60 days to three years.

Canadian sources of HCBD should not contribute significantly to depletion of stratospheric ozone or to climate change. HCBD is not produced or imported in Canada. Main Canadian sources are from combustion and as a byproduct in the production of some chlorinated chemicals. Under the Montreal Protocol, these sources (incidentally produced substances) are not subject to controls.

According to the U.S. Toxic Release Inventory, 2 tonnes of HCBD were released to the environment in the United States in 1995; 75 % of this total was to the air (Toxic Release Inventory, 1997). The load to the atmosphere, however, does not include all possible releases from eve ry type of industrial facility (ATSDR, 1994). HCBD is also on the high production volume list of the Organisation for Economic Co-operation and Development (OECD), which means that it is produced in excess of 10 000 tonnes per year in at least one OECD country (SIDS Manual, 1994). Limited information on quantities, concentrations or conditions of foreign sources of HCBD prevents us from reaching an overall conclusion on the danger to the environment on which life depends.

3.3 CEPA 1999 64(c): Human health

3.3.1 Estimated population exposure

Available data on levels of HCBD in environmental media in Canada upon which estimates of population exposure may be based are quite limited. Point estimates of average daily HCBD intake (on a body weight basis), based on the data on levels of HCBD in ambient air, drinking water and food summarized in Section 2.3.2 and reference values for body weight, inhalation volume and amounts of drinking water and food consumed daily, are presented for five age groups in Table 3. ¹ Based on these estimates, intake may range from 0.01 to 0.2 µg/kg-bw per day. However, it should be noted that these estimates are based on very few samples of only a small number of foodstuffs in early studies in other countries or primarily on limits of detection (or one-half the detection limit for air) in monitoring surveys in which HCBD was only rarely detected in other media. They are presented primarily, therefore, for the purpose of identifying the potential relative contribution of these media to overall population exposure.

If estimates were based on the limit of detection for food and beverages in the limited pilot multimedia study in Toronto in which HCBD was not detected, estimated intakes would be similar to values at the upper end of the range presented in Table 3.

Based on the values derived by either approach, food (or food and beverages) is likely the principal source of exposure, although ambient air may also contribute significant amounts in some areas; drinking water contributes negligibly to overall intake of HCBD. This is consistent with apportionment predicted on the basis of physical/chemical properties or fugacity modelling, although the latter was not helpful in further refinement of estimation of exposure because of a lack of quantitative data on emissions of HCBD into the Canadian environment.

In order to examine the distribution of population exposure to HCBD in Canada, probabilistic estimates were also derived for each of the five age groups, based on information on the distribution of body weights and inhalation volumes, as well as data from the national survey of concentrations of HCBD in ambient air. Data were inadequate to derive probabilistic exposure estimates for other media (i.e., drinking water or food). Estimates of mean, median and 95th-percentile intakes are included in Table 4, along with the point estimates derived in Table 3 for comparison. For example, the 95th-percentile estimates of intake from air range from 0.03 to 0.09 µg/kg-bw per day, compared with point estimates of 0.01-0.05 µg/kg-bw per day.

Table 3 Estimated exposure of the general population to HCBD

Medium	Estimated intake (µg/kg-bw per day)
	0-0.5 years 1	0.5-4 years 2	5-11 years 3	12-19 years 4	20-70 years 5
Air6	<0.02-0.02	0.04-0.05	0.03-0.04	0.01-0.02	0.01-0.02
Drinking water 7	<0.0001	<0.000 06	<0.000 03	<0.000 02	<0.000 02
Food 8	0.03-0.07 9	0.004-0.1	0.001-0.05	0.0009-0.03	0.001-0.03
Total	0.05-0.09	0.04-0.2	0.03-0.09	0.01-0.05	0.01-0.05

Note: Insufficient data were available with which to estimate intake from soil.

3.3.2 Hazard characterization

Because of the inadequacy of data in humans, hazard characterization and dose-response analysis for HCBD are based on studies in experimental animals.

In acute, short-term, subchronic and chronic studies in rats and mice exposed to HCBD via ingestion or inhalation, effects in the pars recta of the proximal tubules of the kidneys (including increased organ weights and biochemical and histopathological evidence of degeneration) consistently occur at the lowest dose or concentration that caused effects (K_ociba et al., 1971, 1977a; Schwetz et al., 1977; Harleman and Seinen, 1979; Stott et al., 1981; Yang et al., 1989; NTP, 1991; Jonker et al., 1993; Birner et al., 1995).

Table 4 Point versus probabilistic estimates of exposure to HCBD via inhalation

Approach	Parameter estimated	Estimated intake by inhalation 1(µg/kg-bw per day)
		0-0.5 years 2	0.5-4 years 3	5-11 years 4	12-19 years5	20-70 years 6
Point estimate	Average daily intake	0.02	0.04-0.05	0.03-0.04	0.01-0.02	0.01-0.02
Probabilistic	Median intake	0.01	0.03	0.03	0.01	0.01
Probabilistic	Mean intake	0.02	0.04	0.03	0.02	0.01
Probabilistic	95th-percentile intake	0.04	0.09	0.06	0.03	0.03

There was also an increased incidence of renal tubular tumours in male and female Sprague-Dawley rats administered the highest dose of HCBD in the diet for two years; nephrotoxicity in the form of hyperplasia and adenomatous proliferation in the renal tubular epithelium was also observed at this as well as a lower dose (K_ociba et al., 1977a). Unlike the mechanism of action associated with other halogenated hydrocarbons, accumulation of a2µ-globulin and hyaline droplet formation are not involved in the formation of renal tumours induced by HCBD.

The weight of available evidence indicates that HCBD is genotoxic in the presence of appropriate metabolic activation systems (Reichert et al., 1984; Vamvakas et al., 1988). This is consistent with the increased incidence of renal tumours observed in rats in vivo, binding of HCBD metabolites to kidney mitochondrial DNA in mice and small amounts of DNA alkylation in the kidney of rats (Stott et al., 1981; Schrenk and Dekant, 1989).

Both genotoxic and non-genotoxic steps may be involved in the induction of tumours by HCBD, although the critical rate-limiting step has not been identified. However, based on observations in the single adequate carcinogenesis bioassay, 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. The renal toxicity of HCBD is closely correlated with the site specificity of accumulation of active metabolites, and there is some (albeit limited) evidence that extent of activation may be less in humans than in rats (e.g., cleavage of the cysteine conjugate by renal b -lyase) (Lock, 1994).

Table 5 Critical studies and effect levels for renal toxicity in experimental animals exposed to HCBD via ingestion

Species	Protocol	Effects at LO(A)EL	Effect levels	Comments	Reference
Wistar rats (5 males and 5 females per group)	Rats were exposed to doses of 0, 1.25, 5 or 20 mg/kg-bw per day in the diet for 4 weeks	Decreased body weight and food consumption; increased relative kidney weight; decreased relative weight of adrenals; effects on urinary and biochemical parameters; histopathological effects in kidney	NOAEL (females) = 1.25 mg/kg-bw per day LOAEL (females) = 5 mg/kg-bw per day NOAEL (males) = 1.25 mg/kg-bw per day LOEL (males) = 5 mg/kg-bw per day	Small number of animals per group	Jonker et al., 1993
Wistar rats (10 males and 10 females per group)	Rats were exposed to doses of 0, 0.4, 1.0, 2.5, 6.3 or 15.6 mg/kg-bw per day by gavage for 13 weeks	Effects on urinary parameters; histopathological effects in kidney	NOEL (females) = 1.0 mg /kg-bw per d ay LOAEL (females) = 2.5 mg/kg-bw per day NOEL (males) = 2.5 mg/kg-bw per day LOAEL (males) = 6.3 mg/kg-bw per day	Small number of animals per group; large number of dose groups with goo d spac ing between dose levels	Harleman and Seinen, 1979
Sprague-Dawley rats (10-12 males and 20-24 females per group; 17 male and 34 female controls)	Rats were exposed to doses of 0, 0.2, 2.0 or 20 mg/kg-bw per day in the diet for about 5 months	Gross and histopathological changes in kidney	NOEL = 0.2 mg/kg-bw per day LOEL = 2.0 mg/kg-bw per day	Small number of animals per group	Schwetz et al., 1977
Sprague-Dawley rats (39-49 males and 40 females per group; 90 male and 90 female controls)	Rats were exposed to doses of 0, 0.2, 2.0 or 20 mg/kg-bw per day in the diet for 2 years	Effects on urinary biochemical parameters; histopathological effects in kidney	NOEL = 0.2 mg/kg-bw per day LO(A)EL = 2.0 mg/kg-bw per day	Good study protocol, except for dose spacing; description of non- neoplastic effects incomplete	Kociba et al., 1977a
B6C3F1 mice (10 males and 10 females per group)	Mice were exposed to doses of 0, 0.1, 0.4, 1.5, 4.9 or 16.8 (males) or 0, 0.2, 0.5, 1.8, 4.5 or 19.2 (females) mg/kg-bw per day in the diet for 13 weeks	Histopathological effects in kidney	LOEL (females) = 0.2 mg/kg-bw per day NOAEL (males) = 1.5 mg/kg-bw per day	Small number of animals per group; large number of exposure groups with good dose spacing	Yang et al., 1989; NTP, 1991

Based on limited data, reproductive and developmental effects and neurotoxicity are not considered to be critical endpoints for HCBD, since effects have been observed only at doses greater than those associated with renal toxicity. Data on effects of HCBD on immunological function have not been identified.

3.3.3 Dose-response analyses

Since non-neoplastic renal effects observed in experimental animals are considered critical and since available data are sufficient, a Tolerable Intake (TI) is derived on the basis of a benchmark dose (BMD) divided by an uncertainty factor. This value is compared with that which might be based on a No-Observed-(Adverse)-Effect-Level (NO[A]EL) for this endpoint, which draws on data from additional studies.

In the available short-term, subchronic and chronic studies, the kidney has consistently been observed to be the most sensitive target organ, with similar effect levels noted in the critical studies (Table 5). In the only identified long-term study in which animals were exposed via ingestion (K_ociba et al., 1977a), an increased incidence of renal tubular hyperplasia/proliferation and an increase in levels of renal coproporphyrin were observed in Sprague-Dawley rats administered 2.0 mg HCBD/kg-bw per day (considered to be the LO[A]EL) or more; renal tubular neoplasms were observed at the highest dose of 20 mg/kg-bw per day. The NOEL was considered to be 0.2 mg/kg-bw per day. Similarly, the LOEL and NOEL for renal toxicity (renal tubular dilation and hypertrophy with foci of renal tubular epithelial degeneration and regeneration) in a subchronic study in the same strain of rats (i.e., Sprague-Dawley) were also 2.0 and 0.2 mg/kg-bw per day, respectively (Schwetz et al., 1977). Renal tubular regeneration (of a severity greater than would be expected, based on comparison with data for the next dose group) also occurred in 1 of 10 mice at the lowest dose tested in a subchronic study in B6C3F₁ mice, 0.2 mg/kg-bw per day (Yang et al., 1989; NTP, 1991), which is considered to be the LOEL. In two of these studies (Harleman and Seinen, 1979; Jonker et al., 1993), decreases in body weight (generally associated with reduced food consumption) were also observed at the LOAEL for renal toxicity.

Sufficient information to permit modelling of the dose-response curve for development of a BMD for renal toxicity was presented in few of these studies. The endpoint that is most amenable to derivation of a BMD is the renal tubular regeneration observed in the 13-week study in B6C3F₁ mice (Yang et al., 1989; NTP, 1991), in which the incidence of this lesion is presented for each dose group. Using the THRESH program, which fits a polynomial model to the data, the BMD₀₅ (the dose associated with a 5% increase in the incidence of renal tubular regeneration) for female mice (which were observed to be more sensitive than males) was 160 µg/kg-bw per day ( χ ² = 0, df = 0, p = 1.0). The 95% lower confidence limit on this value (BMDL₀₅) is 34 µg/kg-bw per day. BMDs calculated for other endpoints in the available subchronic and chronic studies, although based on very limited data in some cases, were greater than those for renal tubular regeneration in female mice presented here.

A TI has been developed on the basis of the BMDL₀₅ for renal tubular regeneration in mice as follows:

where:

• (34) µg/kg-bw per day is the 95% lower confidence limit of the dose estimated to be associated with a 5% increase in renal tubular regeneration in mice administered HCBD for 13 weeks (Yang et al., 1989; NTP, 1991), and
• 100² is the uncertainty factor (x10 for interspecies variation and ¥10 for intraspecies variation; the default values are applied since limited available data on pharmacokinetics and pharmacodynamics in the experimental species and humans are considered insufficient to derive more appropriate values, although the 10-fold factor for interspecies variation is slightly less than a value that would be developed on the basis of the surface area to body weight correction for this species).

This TI is protective, based on consideration of the NOEL for renal toxicity of 0.2 mg/kg-bw per day observed in the chronic study in rats (K_ociba et al., 1977a) and supported by the results of the subchronic studies in rats and mice in which a NOEL and LOEL, respectively, of the same value were observed (Schwetz et al., 1977; Yang et al., 1989; NTP, 1991). Although the variation between doses in the study by K_ociba et al. (1977a) was large (i.e., 10-fold), it was less in the investigation in mice (i.e., 3-fold). Based on application of the same uncertainty factor applied in the derivation of the TI above (i.e., 100) to the NOEL of 0.2 mg/kg-bw per day, the resulting value is greater than 0.34 µg/kg-bw per day (i.e., 2 µg/kg-bw per day).

Available data on the effects associated with inhalation of HCBD are much more limited than those for ingestion. The only relevant studies identified include a short-term study in which renal toxicity was observed in rats exposed to concentrations of 25 ppm (267 mg/m³) HCBD and above for up to 15 days (NOEL = 5 ppm or 53 mg/m³) (Gage, 1970) and a developmental study in which reductions in maternal weight gain were observed in rats exposed to 5 ppm (53 mg/m³) and above (Saillenfait et al., 1989). (Interpretation of this latter observation is complicated by the absence of an exposure-response relationship and the lack of presentation of data on food consumption.) Both of these studies are considered to be inadequate to serve as a basis for derivation of a Tolerable Concentration (TC) in air. If derived on the basi s of the limited existing data, however, such values would, in any case, be greater than that developed above for ingestion, derivation of a TC although, it is noteworthy that renal toxicity was the critical effect in the limited short-term inhal ation st udy in rats.

3.3.4 Human health risk characterization

Based on the point estimates of exposure for the various age groups derived from limited available monitoring data, highly uncertain average total daily intakes of HCBD from air, food and drinking water range from 0.01 to 0.2 µg/kg-bw per day. "Reasonable worst-case" estimates also fall within this range. These estimates are based, for the likely principal medium of exposure, primarily on monitoring data for a small number of foodstuffs for which there is considerable uncertainty about the extent of representation of current exposure of the Canadian public. Some of these data were obtained from industrial areas of other countries at a time when releases of HCBD into the ambient environment were likely much greater than current releases. This is offset to some extent by assumed zero exposure from foodstuffs for which data on concentrations were not available. Moreover, although levels of HCBD in ambient air in Canada have been well characterized in a national survey, it should be noted that estimated intake in this medium is based on half detection limits in the vast majority of samples (>98%) in which HCBD was not detected. In view of these limitations, it is reassuring that the maximum value for estimated average total daily intake and reasonable worst-case estimates (i.e., estimates based on the pilot multimedia study in Toronto in which HCBD was not detected in any medium) of 0.2 µg/kg-bw per day, although also uncertain, is still less than the TI of 0.34 µg/kg-bw per day calculated from the 95% lower confidence limit of the BMD for effects in the kidney in subchronically exposed mice. It should be further noted that this TI is considered conservative, based on a value that might be derived on the basis of a NOEL for renal toxicity in rats exposed to HCBD for two years.

Therefore, on the basis of comparison of estimates of exposure and the TI (i.e., intakes to which it is believed that a person may be exposed daily over a lifetime without deleterious effects), it has been concluded that HCBD is not present in the environment in quantities or under conditions that may constitute a danger in Canada to human life or health.

3.3.5 Uncertainties and degree of confidence in human health risk characterization

There is a high degree of uncertainty inherent in the estimates of intake of HCBD in food, the likely principal medium of exposure, because of the limited number of foodstuffs for which monitoring data are available and the fact that those data that are available were often acquired in early surveys in other countries. There is also considerable uncertainty in the reasonable worst-case estimates for food due to the lack of determination of analytical recovery in the multimedia study.

Although confidence in the estimates of intake in air is greater, since levels of HCBD in ambient air in Canada have been well characterized in a national survey, a degree of uncertainty is introduced by the assumption of half detection limits in the vast majority of samples in which HCBD was not detected. This degree of uncertainty has been characterized quantitatively by calculating intakes also on the basis of the assumption of zero or detection limit for measurements below the detection limit in the national survey. Maximum values for estimated average intakes from air would be approximately one-third of those presented based on an assumption of zero for non-detectable concentrations and twice those presented based on the assumption of detection limit for these samples.

However, there is a high degree of certainty that drinking water contributes only negligible amounts of HCBD to overall exposure, based on the number of large, sensitive investigations.

The only route for which probabilistic estimates of exposure could be derived was inhalation via ambient air. Based on these estimates, intake of HCBD by 95% of the age group with the greatest intake per unit of body weight (i.e., 0.5-4 year olds) is about twice the (uncertain) point estimate for intake via inhalation (i.e., 0.09 µg/kg-bw per day versus 0.04-0.05 µg/kg-bw per day).

In addition, fugacity modelling was not helpful in refinement of estimation of exposure due to the lack of quantitative data on emissions of HCBD into the Canadian environment.

The overall degree of confidence in the population exposure estimates is, therefore, low, primarily as a result of the paucity of current, representative monitoring data for the likely principal medium of exposure of the general population in Canada (food).

The degree of confidence in the database on toxicity that serves as the basis for development of the TI is moderate to high. Although epidemiological data in humans are inadequate, there is consistent evidence from a wide range of acute, short-term, subchronic and chronic studies in rats and mice that critical effects are those that occur in the pars recta of the renal proximal tubules, although data on reproductive effects are somewhat limited, and information on effects on immunological function has not been identified. Moreover, the range of lowest-effect levels at which degenerative renal changes have been observed in long-term studies (subchronic and chronic) is small, and available data are sufficient to develop a BMD and associated lower 95% confidence interval for such effects. Although there is some uncertainty about the mode of induction of tumours by HCBD observed in a single study, there is reasonable assurance that tumours occur only in the presence of degenerative renal changes.

3.4 Conclusions

CEPA 1999 64(a): Based on available data, it has been concluded that HCBD is entering the environment in a quantity or concentration or under conditions that have an immediate or long-term harmful effect on the environment or its biological diversity. Therefore, HCBD is considered to be "toxic" as defined under Paragraph 64(a) of CEPA 1999.

CEPA 1999 64(b): Based on available data, it has been concluded that HCBD is not entering the environment, in Canada in a quantity or concentration or under conditions that constitute a danger to the environment on which life depends. Therefore, HCBD is not considered to be "toxic" as defined under Paragraph 64(b) of CEPA 1999.

CEPA 1999 64(c): Based on available data, it has been concluded that HCBD is not entering the environment in a quantity or concentration or under conditions that constitute or may constitute a danger in Canada to human life or health. Therefore, HCBD is not considered to be "toxic" as defined under Paragraph 64(c) of CEPA 1999.

Overall conclusion: Based on critical assessment of relevant information, HCBD is considered to be "toxic" as defined in Section 64 of CEPA 1999.

3.5 Considerations for follow-up (further action)

Pursuant to Subsection 77(4), because HCBD is considered to be toxic under the Act and meets the criteria for persistence and bioaccumulation in accordance with the Persistence and Bioaccumulation Regulations, is present in the environment primarily as a result of human activity, and is not a naturally occurring radionuclide or a naturally occurring inorganic substance, implementation of virtual elimination of HCBD under Subsection 65(3) is being proposed.

It is recommended that releases of HCBD as a by-product in the production of other chlorinated chemicals, such as vinyl chloride, allyl chloride and epichlorohydrin, be identified and that measures to reduce these releases be investigated.

HCBD releases during refuse combustion were identified. Preliminary information indicates that sources of HCBD from combustion are similar to those of dioxins, furans and hexachlorobenzene. It is recommended that measures to reduce emissions of HCBD from combustion sources complement initiatives currently under way to address dioxins, furans and hexachlorobenzene.

Since HCBD is persistent, bioaccumulative has the potential to harm, benthic species at low levels of exposure and not currently used in commerce in Canada, options to prevent its reintroduction into the Canadian market should be explored.

One potential source of HCBD in Canada identified in the current assessment is transboundary movement from foreign sources. It is recommended, therefore, that the significance of this source be considered in the context of international programs addressing long-range transport of transboundary pollutants.