Warning This Web page has been archived on the Web.

Archived Content

Information identified as archived on the Web is for reference, research or recordkeeping purposes. It has not been altered or updated after the date of archiving. Web pages that are archived on the Web are not subject to the Government of Canada Web Standards. As per the Communications Policy of the Government of Canada, you can request alternate formats on the Contact Us page.

Help the Government of Canada organize its website!

Complete an anonymous 5-minute questionnaire. Start now.

Follow-up Report on Five PSL1 Substances for Which There Was Insufficient Information to Conclude Whether the Substances Constitute a Danger to the Environment

1,2-Dichlorobenzene
1,4-Dichlorobenzene
Trichlorobenzenes
Tetrachlorobenzenes
Pentachlorobenzene

December 2003


(PDF Version - 236 KB)

Table of Contents

List of Acronyms and Abbreviations

Table of Acronyms and Abbreviations
Acronym or abbreviationDefinition
ACRacute to chronic ratio
AETApparent Effect Threshold
BAFbioaccumulation factor
BCFbioconcentration factor
CBzchlorobenzene
CBzschlorobenzenes (1,2-dichlorobenzene, 1,4-dichlorobenzene, trichlorobenzenes, tetrachlorobenzenes, pentachlorobenzene)
CEPACanadian Environmental Protection Act
CEPA 1999Canadian Environmental Protection Act, 1999
CTVCritical Toxicity Value
CTVSEDCritical Toxicity Value for sediments
DCBdichlorobenzene
1,2-DCB1,2-dichlorobenzene
1,4-DCB1,4-dichlorobenzene
dwdry weight
EC50median effective concentration
EEVEstimated Exposure Value
ENEVEstimated No-Effects Value
ENEVSEDEstimated No-Effects Value for sediments
ENEVSOILEstimated No-Effects Value for soils
EqPequilibrium partitioning
HC5hazardous concentration for 5% of exposed organisms
Kowoctanol-water partition coefficient
LC50median lethal concentration
LC90lethal concentration to 90% of the test organisms
LOECLowest-Observed-Effect Concentration
MDLmethod detection limit
NOECNo-Observed-Effect Concentration
NPRINational Pollutant Release Inventory
OCorganic carbon
PCBspolychlorinated biphenyls
PSLPriority Substances List
PSL1first Priority Substances List
QCBpentachlorobenzene
STPsewage treatment plant
TCBtrichlorobenzene
TeCBstetrachlorobenzenes

Back to Top

Synopsis

1,2-Dichlorobenzene (1,2-DCB), 1,4-dichlorobenzene (1,4-DCB), trichlorobenzenes (TCBs), tetrachlorobenzenes (TeCBs) and pentachlorobenzene (QCB), which appeared on the first Priority Substances List (PSL1), were assessed to determine whether these substances should be considered “toxic” as defined under the Canadian Environmental Protection Act (CEPA). It was concluded in the PSL1 assessment that these compounds were not “toxic” under Paragraphs 11(b) or 11(c) of CEPA; however, there was insufficient information to conclude whether they could have immediate or long-term harmful effects on the environment, under Paragraph 11(a). Concentration data for these chlorobenzenes (CBzs) in freshwater and marine sediments and soil environments were lacking. Corresponding data reporting effects on benthic and soil-dwelling organisms were also needed to complete this assessment.

Subsequent to the completion of the PSL1 assessments, a revised CEPA, CEPA 1999, came into effect. Paragraph 64(a) of CEPA 1999 has a definition of “toxic” that is similar to that in Paragraph 11(a) under the original CEPA and addresses whether a substance has or may have an immediate or long-term harmful effect on the environment. However, in CEPA 1999, Paragraph 64(a) has been expanded to include effects on biodiversity. Research studies to address data gaps for the CBzs of interest were funded, and emphasis was placed on studies that examined effects on benthic organisms exposed to the CBzs of interest. Additionally, recent literature was reviewed for new data on concentrations in sediment and soil for each of the CBzs under consideration and for information on the effects on organisms resulting from exposure to these compounds.

Both 1,2-DCB and 1,4-DCB are produced in Canada, based on reports from the early 1990s. 1,4-DCB is used more extensively than 1,2-DCB, primarily as an air freshener/deodorizer. During the mid-1990s, 40–45 tonnes of TCBs were expected to be imported into Canada, although imports of TeCBs and QCB were not anticipated.

The primary route of entry for CBzs into Canadian surface waters and associated sediments is via effluents from industrial and sewage treatment plants. 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB have been identified in pulp and paper mill effluents. Effluents from iron and steel manufacturing contribute to loadings of TCBs, TeCBs and QCB, while petroleum refinery effluents have been reported to contain TeCBs and QCB. The more highly chlorinated benzenes, particularly hexachlorobenzene, are subject to reductive dechlorination, which may contribute to accumulation of the lower chlorinated homologues (e.g., DCBs and TCBs) in buried sediments. The main source of CBzs to Canadian soils is accidental spillage of industrial chemicals, although CBzs may be added to agricultural soils during amendment with sewage sludge. Industrial emissions to the atmosphere represent another route of entry into the Canadian environment.

Maximum Canadian concentrations of each of the CBzs under consideration in this report were observed in sediment samples collected from the St. Clair River in Ontario. 1,4-DCB was the only CBz detected in Canadian soil samples.  The CBzs of interest in this report are known to cause both chronic and acute effects in controlled tests on benthic and soil-dwelling organisms. In general, benthic organisms are more sensitive to the CBzs than soil-dwelling species, based on toxicity studies to date. 

Concentrations of the CBzs of interest in the highly contaminated sediments of the St. Clair River are elevated enough that sensitive benthic organisms could experience adverse effects.

Each of the CBzs under investigation in this report has been estimated to persist in sediment for longer than 2 years. The half-lives of 1,2-DCB, 1,4-DCB, TCBs and TeCBs in soil have been estimated to be approximately 8 months, while QCB’s half-life in soil has been estimated to be 2 years.  Additionally, TeCBs and QCB are subject to atmospheric transport from its source to remote areas and, therefore, are considered persistent in air.  All of the CBzs of interest in this report therefore meet the criteria for persistence as defined in the Persistence and Bioaccumulation Regulations of CEPA 1999(Government of Canada, 2000) due to the persistence of these compounds in sediment and soil.  The higher chlorinated products, TeCBs and QCB also are persistent in air.  The lower chlorinated benzenes (1,2-DCB, 1,4-DCB and TCBs) are not expected to be highly bioaccumulative. However, the TeCBs and QCB do have a high potential to bioaccumulate and meet the bioaccumulation criteria defined in the Persistence and Bioaccumulation Regulations of CEPA 1999(Government of Canada, 2000).

There are special concerns about persistent and bioaccumulative substances. Persistent substances can remain in the environment for long periods of time, increasing the probability and the duration of exposure.  In addition persistent substances are subject to long-range transport, which results in low-level, widespread contamination  Bioaccumulative substances have the potential to biomagnify, and consequently releases of extremely low concentrations of persistent and bioaccumulative substances may - either alone or in combination with other similar substances - cause severe adverse effects.

Based on the information available, it is concluded that 1,2-DCB, 1,4-DCB and TCBs are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity, but TeCBs and QCB are entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity. Therefore, 1,2-DCB, 1,4-DCB and TCBs are not considered “toxic” as defined in Paragraph 64(a) of CEPA 1999. It is, however, concluded that TeCBs and QCB be considered “toxic” as defined in Paragraph 64(a) of CEPA 1999.

QCB and TeCBs are persistent, bioaccumulative, predominantly anthropogenic and are considered “toxic” under Paragraph 64(a) of CEPA 1999, and as such, meet the criteria for Track 1 substances under the Toxic Substances Management Policy.  Therefore QCB and TeCB should be subject to virtual elimination of releases to the environment.  Since there is currently no commercial demand for QCB and TeCBs in Canada, options to prevent their reintroduction into the Canadian market should be explored.

Back to Top

1. Introduction

1,2-Dichlorobenzene (1,2-DCB), 1,4-dichlorobenzene (1,4-DCB), trichlorobenzenes (TCBs), tetrachlorobenzenes (TeCBs) and pentachlorobenzene (QCB) appeared on the first Priority Substances List (PSL1) of the Canadian Environmental Protection Act (CEPA), which was published in the Canada Gazette, Part I, on February 11, 1989. Assessments were performed to determine whether these chlorobenzenes (CBzs) should be considered “toxic” as defined under CEPA andwere completed in 1993 (Government of Canada, 1993a, 1993b, 1993c, 1993d, 1993e)1. It was concluded that these substances do not constitute a danger either to the environment on which human life depends or to human life or health, and, therefore, they were not found to be “toxic” under Paragraph 11(b) or 11(c) of CEPA. Additionally, during the period over which the original assessments were conducted, it was determined that concentrations of 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB present in Canadian air and surface waters were not likely to cause adverse effects on aquatic biota or wildlife. There was, however, a lack of acceptable data on the effects of these CBzs on benthic and soil-dwelling organisms and on concentrations of CBzs in Canadian soils. Therefore, it was not possible to determine whether environmental harm was occurring due to accumulations of these substances in sediment and soil. The lack of data led to the conclusion that there was insufficient information available on sediments and soils to determine whether these substances should be considered “toxic” under Paragraph 11(a) of CEPA.

A revised CEPA, CEPA 1999, came into effect on March 31, 2000. This new legislation includes Paragraph 64(a), which is similar to Paragraph 11(a) of the original CEPA and addresses whether a substance has or may have an immediate or long-term harmful effect on the environment, although it has been expanded to include effects on biodiversity. CEPA 1999 places more emphasis on pollution prevention, gives consideration to the precautionary principle and requires special treatment of persistent and bioaccumulative substances. Substances that are shown to be both persistent and bioaccumulative, therefore, will be assessed using a more conservative approach than is used for other substances.

As a result of the publication of the PSL Assessment Reports for the CBzs, additional studies were designed and funded. Day et al. (1995) and Doe et al. (1995) reported data on the toxicity of 1,2-DCB, 1,4-DCB, TCBs and TeCBs to freshwater and marine benthic organisms. Additionally, concentrations of 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB were determined in sediments near point sources (i.e.,outfalls from sewage treatment plants [STPs] and textile manufacturing plants) in Atlantic Canada. Laboratory studies included toxicity testing to determine effects of exposure to sediments from point source locations (Rutherford et al., 1995). Concentrations of 1,2-DCB and 1,4-DCB were determined in sediment collected by the Ontario Ministry of the Environment near outfalls from chemical manufacturing plants and an STP on the St. Clair River, near Sarnia (DeLuca and Fox, 1995).

A literature search for new data on the CBz substances of interest was performed in 1995 and repeated in 1999. The National Pollutant Release Inventory (NPRI) and Accelerated Reduction/Elimination of Toxics databases supported by Environment Canada were also reviewed for CBz data.

Federal and Ontario government scientists were also requested to provide unpublished data relevant to the assessment of 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB under Paragraph 11(a) of CEPA or Paragraph 64(a) of the revised CEPA, CEPA 1999.Footnote 1

Recent data confirmed the previous conclusion, which found that environmental harm is not likely to result from the presence of 1,2-DCB, 1,4-DCB, TCBs, TeCBs or QCB in Canadian air and water. The focus of this report, therefore, is the determination of whether accumulations of the specific CBzs in aquatic sediments or soils would harm exposed benthic or soil-dwelling organisms.

Back to Top

2. Entry Characterization

2.1  Anthropogenic Releases in Canada

There are no known natural sources of 1,2-DCB, 1,4-DCB, TCBs, TeCBs or QCB. 1,2-DCB and 1,4-DCB are the only CBzs under examination that are produced in Canada (Table 1). Based on reports from the mid-1990s, 1,4-DCB is used to the greatest extent in Canada (1000 tonnes), primarily as an air freshener/deodorizer (Kovrig, 1996). Survey results projected imports of TCBs to range from 40 to 45 tonnes during the mid-1990s (Camford Information Services, 1991). Imports of TeCBs and QCB into Canada were not anticipated based on survey results (Camford Information Services, 1991).

Both 1,2-DCB and 1,4-DCB releases were reported to the NPRI for 1994–1998. Releases of 1,2-DCB ranged from 0.4 to 0.5 tonnes, while 1,4-DCB releases ranged between 8.1 and 10.4 tonnes (NPRI, 1994, 1995, 1996, 1997, 1998). Releases to air were reported for both compounds. Disposal by incineration was reported to range between 8 and 23 tonnes for 1,2-DCB and between 0.4 and 0.5 tonnes for 1,4-DCB (NPRI, 1994, 1995, 1996, 1997, 1998). Total 1,4-DCB emissions were reported to be 55 tonnes by members of the Canadian Chemical Producers’ Association in 1997, down from 116 tonnes in 1993 (Canadian Chemical Producers’ Association, 1999). Releases of the other CBzs of interest in this report (TCBs, TeCBs and QCB) have not been reported to the NPRI.

CBzs enter Canadian surface waters and associated sediments primarily via effluents from industrial treatment plants and STPs. The major industrial sectors include chemical manufacturing and textile plants. CBzs have also been observed in effluents from pulp and paper mills (Government of Canada, 1993a, 1993b, 1993c, 1993d, 1993e). Effluents from iron and steel manufacturing contribute to loadings of the TCBs, TeCBs and QCB, while petroleum refinery effluents have been reported to contain TeCBs and QCB (Government of Canada, 1993c, 1993d, 1993e). Reductive dechlorination of the more highly chlorinated benzenes, particularly hexachlorobenzene, may lead to the accumulation of the lower chlorinated homologues (e.g., DCBs and TCBs) in buried sediments (Beurskens et al.,1993a, 1993b).

The main reported source of CBzs to Canadian soils is accidental spillage of industrial chemicals, including dielectric fluids containing polychlorinated biphenyls (PCBs) (Government of Canada, 1993a, 1993b, 1993c, 1993d). Other possible sources include industrial emissions to the atmosphere (Ding et al., 1992) and application of sewage sludge to agricultural soils (Webber and Nichols, 1995).

A few studies have been performed in which CBz concentrations have been reported in sewage sludge. In a study of sewage sludge samples collected from 12 municipalities across Canada, levels of 1,2-DCB and 1,4-DCB were reported (Table 2). Levels of TCBs were below detection limits in sludge from all municipalities. TeCBs and QCB were not included in this survey (Webber and Nichols, 1995). DCB concentrations observed in Canadian sludge samples were lower than concentrations reported in the United States during the 1980s and below levels currently reported in the United Kingdom (Table 2).

Back to Top

3. Exposure Characterization

3.1 Environmental Fate

3.1.1 Sediment

Under anaerobic laboratory conditions, dechlorination by both biotic and abiotic processes has been observed for all CBzs (Bosma et al., 1988; Peijnenburg et al., 1992; Beurskens et al., 1993b, 1994; Yonezawa et al., 1994). Reported dechlorination half-lives range from only a few days to over 1 year, depending upon the CBz studied and the nature of the sediment used. Mackay et al. (1992) estimated average half-lives in surface sediment of approximately 2 years for all of the CBzs considered in this report. CBzs entering the water column generally partition to particulate matter and accumulate in bottom sediments, based on results of fugacity modelling (Mackay et al.,1992)and empirical studies (Oliver and Carey, 1986). CBzs have been shown to persist in sediments for long periods. Oliver and Nicol (1982, 1983) compared the relative proportions of different CBz congeners in surface and subsurface sediments and found little evidence of either microbial oxidation or anaerobic dechlorination of higher chlorinated benzene congeners in Lake Ontario sediments. CBzs have been detected in sediment cores dating back to the early 1900s (Eisenreich et al., 1989; Muir et al., 1995, 1996; Rawn et al., 2000a, 2000b).

Although a large fraction of the di- through pentachlorinated benzenes partition to organic matter, there will be some fraction present in sediment pore waters, either complexed as colloids with xdissolved organic matter or as freely dissolved molecules (Di Toro et al.,1991). The uncomplexed molecules may pass through cell membranes, entering organisms during exposure to sediment pore waters. Additionally, direct ingestion of organic carbon (OC) contaminated with CBzs may be an important route of exposure for some benthic organisms.

Desorption studies have suggested that irreversible adsorption occurs in some sediments, and, therefore, equilibrium is not always achieved. Irreversibility in binding of organic compounds is expected to increase with exposure time, although this theory has been questioned by some authors (Kan et al., 1994).

3.1.2 Soil

CBzs may enter surface soil as a result of spills, from sewage sludge additions and via atmospheric deposition from both local and distant industrial sources. Similar to observations in sediments, CBzs partition between particulate and liquid phases. Due to the hydrophobic nature of the di- through penta-CBzs, they are considered to be relatively immobile in soils, particularly in soils with a high OC content. Volatilization and biodegradation are the main routes of loss for these compounds from soils. Mean half-lives in soil have been estimated to be approximately 8 months for 1,2-DCB, 1,4-DCB, TCBs and TeCBs and 2 years for QCB (Mackay et al., 1992).

Although the bioavailability of these compounds may be reduced in aged soils (Gas Research Institute, 1995), uptake by soil-dwelling organisms can also occur by exposure to freely dissolved forms in pore waters. Another route of exposure for soil-dwelling organisms is ingestion of soil organic matter. In plants, absorption of CBzs may occur via direct uptake by roots or through foliage after volatilization from the soil surface (Trapp et al., 1990; Scheunert et al., 1994; Wang and Jones, 1994).

3.1.3 Biota   

Bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) for CBzs have generally been reported on a whole-body basis. BCFs ranging between 270 and 560 were reported for 1,2-DCB in rainbow trout (Oncorhynchus mykiss) in laboratory studies (Government of Canada, 1993a). Reported BCFs for 1,4-DCB in rainbow trout ranged between 370 and 1400 (Government of Canada, 1993b). BCFs for TCBs were reported to be between 100 and 4000 in a variety of aquatic biota (Government of Canada, 1993c).

BAFs reported for TeCBs were between 1180 and 135 000 in fathead minnow (Pimephales promelas), rainbow trout, guppy(Poecilia reticulata) and earthworms (Eisenia andrei) (Government of Canada, 1993d). BAFs of 810 and 20 000 were reported for QCB in mussel (Mytilis edulis) and rainbow trout, respectively, but a much higher BAF for earthworms (E. andrei) (401 000) has also been reported (Government of Canada, 1993e). More recently, Burkhard et al. (1997) reported BAFs based on freely dissolved, lipid-normalized concentrations for TCBs, TeCBs and QCB in a number of species. When considered on a whole-body wet weight basis, the BAFs reported by Burkhard et al. (1997) were between 427 and 630, between 871 and 1905, and between 6310 and 12 883 for TCBs, TeCBs and QCB, respectively. Bioaccumulation of CBzs generally increases with degree of chlorination.

The estimated log octanol-water partition coefficient (log Kow) for both 1,2-DCB and 1,4-DCB is 3.4. The log Kow estimates for TCB, TeCBs and QCB are 3.85–4.30, 4.5 and 5.0, respectively (Mackay et al., 1992).

3.2 Environmental Concentrations

3.2.1 Sediment Near Point Sources

The CBzs, similar to other non-ionic hydrophobic compounds, partition into the organic matter in sediment. Additionally, the bioavailability of these compounds is inversely proportional to the OC content of the sediment. Therefore, CBz concentrations have been OC normalized in Table 3 using the relationship:

OC-normalized concentration (mg/kg dw) :

  • = CCBz / foc

where CCBz represents the CBz concentration in whole sediment (mg/kg dw) and foc represents the OC fraction in the sediment.

The highest reported CBz concentrations in Canadian sediment were observed near industrial sites on the St. Clair River at Sarnia, Ontario, during the 1980s ( S DCBs: <MDL–31 µg/g dw, or <MDL–2070 µg/g OC) (Oliver and Pugsley, 1986). In general, the highest concentrations of the CBzs of interest were measured in samples collected near the Dow Chemical Canada 1st Street Sewer. Sampling stations were located upstream and downstream of chemical manufacturing sites. Fox et al. (1983) reported high CBz concentrations (maximum 1,4-DCB concentration = 1.3 µg/g dw, or 37 µg/g OC) in surficial sediment in Lake Ontario near the Niagara River mouth (Table 3).

More recently, 1,4-DCB was detected in sediment samples collected near municipal wastewater treatment plant effluents (<10–90 ng/g dw, or <0.1–16 µg/g OC) in Nova Scotia and New Brunswick (Rutherford et al., 1995). In another recent study, elevated levels of 1,4-DCB (1.7 µg/g dw, or 40 µg/g OC) were reported in sediment near outfalls from municipal wastewater treatment plants near Victoria, British Columbia (EVS, 1992, 1996). The highest CBz concentrations in Canadian sediment remain in the St. Clair River, adjacent to organic chemical and petrochemical plants near Sarnia (DeLuca and Fox, 1995; Kauss, 1995). Median concentrations of the CBzs of interest in the most contaminated stretch of the river (1–2 km) fell by as much as an order of magnitude between 1984 and 1994 (Oliver and Pugsley, 1986; Bedard and Petro, 1992; Kauss, 1995). However, direct comparisons between the older sampling and more recent work are not possible due to differences in sampling locations and sample collection and analysis techniques.

No recent concentration data for CBzs in sediment from the Niagara River delta were identified. Results of ongoing monitoring of Niagara River water, however, indicate that concentrations of CBzs in river water and suspended sediments have decreased by as much as 10-fold since the early 1980s (Kuntz, 1993), similar to observations in sediments from the St. Clair River. It is, therefore, expected that current CBz concentrations in surface sediment in the Niagara River delta are lower than previously reported for this region (Table 3).

Although 1,4-DCB concentrations were elevated in sediment samples collected near the outfalls of primary STPs at Sarnia, Ontario (DeLuca and Fox, 1995; Kauss, 1995), Halifax, Nova Scotia (Rutherford et al., 1995), and Victoria, British Columbia (EVS, 1992; Chapman et al., 1996), levels were below detection (approximately 0.01 µg/g dw, or 2.2 µg/g OC) near a secondary STP at Fredericton, New Brunswick, and a lagoon STP at Berwick, Nova Scotia (Rutherford et al., 1995). These results indicate that 1,4-DCB levels are not enriched in sediments near all the Canadian STPs (Rutherford et al., 1995). The CBzs of interest were also below detection limits (approximately 0.01 µg/g dw, or 2.4 µg/g OC) in sediment collected near the outfalls from textile plants in Caraquet, New Brunswick, Bridgetown, Nova Scotia, and Magog, Quebec (Rutherford et al., 1995).

3.2.2 Long-range Transport

Some of the CBzs of interest (TeCBs and QCB) have been reported in lake sediments from both temperate regions and the Canadian Arctic (Eisenreich et al., 1989; Muir et al., 1995, 1996; Allen-Gil et al., 1997; Rawn et al., 2000a, 2000b). Movement of organic compounds to Arctic regions via long-range transport and deposition has been the focus of much study in recent years. Muir et al. (1996) reported that maximum CBz (represented by S [QCB + hexachlorobenzene]) concentrations were observed in lake sediments dated to the late 1970s and 1980s, approximately 5–10 years later than maximum concentrations in Lake Ontario. These results are consistent with the cold condensation hypothesis, which explains the movement of organics to remote northern regions (Wania and Mackay, 1993). Allen-Gil et al. (1997) reported TeCBs ( S [1,2,3,4-TeCB + 1,2,4,5-TeCB]) and QCB levels in surface slices of sediment cores collected in Arctic U.S. lakes (mean concentrations: 0.41 ng/g dw and 0.10 ng/g dw, respectively). Total TeCBs ( S [1,2,3,4-TeCB + 1,2,4,5-TeCB]) concentrations detected in Yukon lake sediments ranged from below the MDL (<0.03 ng/g dw) to 0.54 ng/g dw, and QCB levels ranged from below detection levels (<0.03 ng/g dw) to 1.55 ng/g dw (Rawn et al., 2000b).

3.2.3 Soil

CBzs are expected to partition to solid organic matter (Kenaga and Goring, 1980), although the most bioavailable fraction exists in soil pore waters. The freely dissolved, uncomplexed fraction is inversely proportional to the OC content of the soil (van Gestel and Ma, 1988; Trapp et al., 1990; Hulzebos et al., 1993; CEU, 1995). Therefore, concentrations have been summarized on an OC-normalized basis (Table 4).       

CBz concentrations in agricultural soils from 14 sites across Canada were generally below detection limits (approximately 0.05 µg/g dw, or 3.5 µg/g OC) (Webber, 1994). 1,4-DCB, the only CBz detected in the study, had a detection frequency of approximately 20%, and the maximum reported concentration was 0.14 µg/g dw (4.5 µg/g OC).

The CBzs of interest in this report (di- to pentachloro-) have not been measured in sludge-amended soils in Canada. 1,2-DCB and 1,4-DCB concentrations in sludge-amended soils can be estimated using the Ontario Ministry of Environment regulations (OMEE, 1994), in which 40 000 kg/ha was taken to be the maximum addition of sludge to soil (Webber and Nichols, 1995) and 2 × 106 kg/ha was taken to be the mass of soil in the plough layer. Mean concentrations of 1,2- and 1,4-DCB in Canadian sludge were observed to be 0.42 mg/kg dw and 0.87 mg/kg dw, respectively (Webber and Nichols, 1995). This indicates that soil concentrations following sludge treatment would be 0.008 µg/g dw (0.4 µg/g OC) and 0.017 µg/g dw (0.9 µg/g OC), respectively, assuming an OC content of 2%, based on the following relationship using 1,2-DCB as an example:

Soil concentration following sludge treatment :

  • = (0.42 mg/kg dw × 40 000 kg) / (2 × 106 kg)
  • = 0.008 mg/kg dw soil (or µg/g dw soil)

CBz contamination in Canadian soils as a result of atmospheric fallout from nearby industrial activity has not been studied. Elevated concentrations were, however, observed in soils downwind of a highly industrialized area near Niagara Falls, New York (Ding et al.,1992) (Table 4). The Niagara Falls soils have been considered representative of “worst-case” exposure conditions for industrial areas in Canada for the TCBs, TeCBs and QCB.

Back to Top

4. Effects Characterization

4.1 Benthic Organisms

Effects of 1,2,4-TCB on benthic organisms were reported in two studies during the 1980s. A significant effect on marine macrobenthic community structures was observed in a number of taxa exposed to 1,2,4-TCB at nominal concentrations of 100 and 1000 µg/g dw in sediments (Tagatz et al., 1985). Throughout the study, measured concentrations in sediments ranged between 2.1 and 97 µg/g dw and between 519 and 790 µg/g dw, respectively. Due to the extremely low (<0.02%) OC level in the sediment, conversion to OC-normalized results was not possible for these data (Di Toro et al., 1991).

Clark et al. (1987) conducted 10-day bioassays with two marine species using sediment containing approximately 0.3–0.6% OC, spiked with 1,2,4-TCB. No lethality was observed in grass shrimp (Palaemonetes pugio) or amphioxus (Branchiostoma caribaeum) at nominal concentrations of 10 µg/g dw (approximately 2000 µg/g OC) and 75 µg/g dw (approximately    15 000 µg/g OC).

More recently, Day et al. (1995) reported chronic whole-sediment toxicity of 1,2-DCB, 1,4-DCB, 1,2,3-TCB and 1,2,4,5-TeCB to two species of freshwater benthic invertebrates, mayfly (Hexagenia spp.) and oligochaete worms (Tubifex tubifex), in sediment systems under open static conditions over 21 and 28 days, respectively. Nominal concentrations of 0.5, 5, 50 and 500 µg/g dw were used for 1,2-DCB, 1,4-DCB and 1,2,3-TCB exposures, while 1,2,4,5-TeCB concentrations of 0.5, 5, 50 and 150 µg/g dw were used. Concentrations were measured at the beginning and termination of individual studies(Day et al., 1995). For these studies, a mixture of natural sediment, kaolin and fine silica sand, with an average OC content of 3.93% (3.38–4.45%), was used. Survival and biomass and survival and reproduction were taken to be the endpoints for Hexagenia and T. tubifex studies, respectively. Survival was not affected by exposure to CBzs in either species, although reduction in growth of Hexagenia spp. was observed (18–34%) in the 500 µg/g dw exposures with 1,4-DCB and 1,2,3-TCB and in the 150 µg/g dw exposure with 1,2,4,5-TeCB (Table 5). Reproduction of T. tubifex was impaired (64–72%) in the 500 µg/g dw exposure trials with 1,2-DCB, 1,4-DCB and 1,2,3-TCB (Table 5). Because effects on growth and reproduction were observed at only the highest concentration level of each exposure series tested (Table 5) -- with the exception of 1,2-DCB and 1,2,4,5-TeCB for Hexagenia spp. and Tubifex tubifex, respectively, where no effect was observed following exposure to spiked sediments -- estimates of traditional endpoint values (e.g., LC50, EC50, LOEC, etc.) were not possible.

Doe et al. (1995) reported acute toxicity of 1,2-DCB, 1,4-DCB, 1,2,3-TCB and 1,2,4,5-TeCB based on 10-day exposures of the infaunal amphipod Rhepoxynius abronius in marine sediments. The sediment used in this series was a mixture of two natural sediments, resulting in an OC content of 0.55% (Doe et al., 1995). Nominal concentrations of 4, 20, 100 and 500 µg/g dw were used in the 1,2-DCB, 1,4-DCB and 1,2,3-TCB exposure studies. For 1,2,4,5-TeCB, nominal concentrations were 1.2, 6.0, 30 and 150 µg/g dw. Concentrations of each exposure system were measured at the beginning of each study, although measurements at the end of the study were restricted to the highest treatment level only (Doe et al., 1995). Mortality of R. abronius was significant at the 100 µg/g dw nominal exposure for 1,2-DCB and 1,2,3-TCB, while significant mortality was reported in the 500 µg 1,4-DCB/g dw and 30 µg 1,2,4,5-TeCB/g dw nominal exposure systems (Table 6). Doe et al. (1995) reported LOEC and NOEC values on a ng/g dw basis for the DCBs, 1,2,3-TCB and 1,2,4,5-TeCB, which have been converted to OC-normalized values in Table 6.

The mode of action of the CBzs is considered non-specific or narcosis (van Wezel et al., 1996a, 1996b). Effect levels for individual CBzs are, therefore, expected to be approximately equal, for a given species, based on molar concentrations (McCarty et al., 1992). Because Day et al. (1995) and Doe et al. (1995) did not conduct toxicity tests for QCB, estimates of its effect levels were made based on the results of the four CBzs tested, on a molar basis. For example, for 1,2,4,5-TeCB, which has a molecular weight of 215.9 g/mol, the LOEC (8.7 µg/g dw, or 1582 µg/g OC) reported by Doe et al. (1995) on a molar basis would be:

LOEC :

  • = (1582 µg/g OC) / 215.9 g/mol
  • = 7.33 µg·mol/g2

The molar LOEC for QCB, therefore, was estimated to be:

LOEC :

  • = 7.33 µg·mol/g2 × 250.3 g/mol
  • = 1835 µg/g

This calculation was repeated for 1,2-DCB, 1,4-DCB and 1,2,3-TCB, and the results for each congener were used to determine a range where effects would be expected due to QCB exposure, for each organism. The range of QCB concentrations over which the lowest effect level would be expected, based on individual congener calculations for T. tubifex,was 2750–9010 µg/g OC. The range of lowest QCB concentrations expected to cause effects in Hexagenia spp. was estimated to be between 400 and 9510 µg/g OC. LOEC values for R. abronius were estimated to range between 1840 and 10 410 µg/g OC. Calculations of LOEC values for R. abronius were based solely on initial concentrations.

Additional techniques may be used as part of a weight-of-evidence approach to determine the toxicity of organic compounds to benthic organisms. Di Toro et al. (1991) proposed a method for estimating toxicity based on the assumption that equilibrium exists between non-ionic compounds bound to sediment OC and those freely dissolved in pore waters. Concentrations of the freely dissolved compounds in pore waters are considered to be proportional to OC-normalized values in sediment, and, therefore, effects of dissolved compounds on pelagic organisms may be used as a surrogate for effects on benthic organisms exposed to contaminated sediments (Di Toro et al., 1991). Estimates of OC-normalized concentrations in sediment resulting in effects on freshwater and marine organisms were made using the equilibrium partitioning (EqP) method (Table 7). Concentrations were calculated using the relationship:

Csed (µg/g OC) :

  • = [Cdiss (µg/L) × Kow (L/kg)] / 1000

where Csed represents the OC-normalized concentration in sediment that is likely to cause an effect in benthic biota, Kow is the octanol-water partition coefficient, Cdiss represents the corresponding concentration in the freely dissolved state and 1000 is the factor used to convert µg/kg OC to µg/g OC.

By using the lowest effect concentrations (e.g., LC50 and EC50 [16–28 days] data) for pelagic organisms reported in the literature, benthic effects estimates were determined using the EqP method (Table 7). Toxicity data for all the CBzs of interest were available in the literature, although a larger data set exists for freshwater organisms than for marine organisms (Government of Canada, 1993f, 1993g, 1993h, 1993i; Environment Canada, 1994).

4.2 Soil-dwelling Biota

CBzs have been shown to affect soil microbial populations. Marinucci and Bartha (1979) reported a 24-hour EC50 (respiration) for soil microorganisms at approximately 50 µg 1,2,4-TCB/g dw based on nominal concentrations. In another study, microbial respiration was depressed by addition of 1000 µg 1,2-DCB/g dw (67 114 µg/g OC) initially, although no effect on respiration was observed by the final day of the 6-day experiment (Walton et al., 1989). Fourteen-day LC50 values reported for several earthworm species exposed to 1,4-DCB, 1,2,3-TCB and 1,2,4-TCB ranged from 115 to 563 µg/g dw soil (2592 to 6500 µg/g OC) (Table 8) (Neuhauser et al.,1986;van Gestel and Ma, 1990; van Gestel et al.,1991).

Hulzebos et al. (1993) reported 7- to 14-day EC50 values (growth) for lettuce (Lactuca sativa) exposed to DCBs, TCBs and TeCBs. Measurements using 1,2-DCB were not performed in the study, and no similar data were found in the literature for 1,2-DCB; therefore, the effect concentration for 1,4-DCB was taken to represent that for 1,2-DCB (Table 8). Lettuce was found to be more sensitive to 1,2,3-TCB (5.8 µg/g dw) and 1,2,4,5-TeCB (4.2 µg/g dw) than 1,4-DCB, for which the EC50 was estimated to be 213 µg/g dw (19 722 µg/g OC). EC50 values for individual TeCBs and TCB isomers varied by one and two orders of magnitude, respectively (Table 8).

Back to Top

5. Assessment of “Toxic” Under CEPA 1999

5.1 CEPA 64(a): Environment

The environmental risk assessment of a PSL substance is based on the procedures outlined in Environment Canada (1997). 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 for ecological harm 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, however, special concerns about persistent and bioaccumulative substances. Persistent substances can remain in the environment for long periods of time, increasing the probability and the duration of potential exposure. Releases of extremely low concentrations of persistent and bioaccumulative substances may lead to accumulations in organisms which can eventually cause adverse effects. Substances that, because of their persistence are subject to long-range transport, are of particular concern because they can result in low-level, widespread contamination. Remote and cold regions, such as the Canadian Arctic, can act as a sink for these compounds. Bioaccumulative substances have the potential to biomagnify through the food chain. Although current science is unable to accurately predict the cumulative effects of exposure to low levels of persistent and bioaccumulative substances on the environment, the potential exists for extensive, irreversible impacts. Assessments of such substances must be performed using a proactive, preventative approach to ensure that widespread cumulative effects do not occur. Therefore, environmental assessments of persistent and bioaccumulative substances require a more conservative approach than that used for other substances, even in situations where a substance is released in a small area and effects appear to be localized.

Conservative methodologies are used for both the exposure and effects characterizations for persistent and bioaccumulative substances. If exposure monitoring data are available, the maximum reported/estimated field concentration is used as the EEV. An additional application factor of 10 was to be used in the effects characterization to calculate the ENEV for persistent and bioaccumulative substances.

5.2 Persistence and Bioaccumulation Criteria as Defined in the Persistence and Bioaccumulation Regulations of CEPA 1999

The persistence criteria as defined under the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada, 2000) are available in Appendix A.

5.2.1 Persistence

5.2.1.1 Sediment

Mackay et al. (1992) estimated average half-lives in surface sediment of approximately 2 years for all of the CBzs considered in this report. Additionally, the tetra- and pentachlorinated congeners have been identified in sediments from lakes in both temperate regions and northern Canadian environments (Eisenreich et al., 1989; Muir et al., 1995; Rawn et al., 2000b). The detection of the TeCBs and QCB in northern lake sediments in the absence of nearby sources indicates that these residues are a result of long-range transport, and these CBzs, therefore, meet the criteria for persistence in air. CBzs have been reported in sediments dated to the early 1900s, although maximum concentrations were reported to occur between the 1970s and 1980s. These data are consistent with half-life estimates exceeding 1 year in a variety of sediments.

5.2.1.2 Soil

Mean half-lives in soil have been estimated by Mackay et al. (1992) to be approximately 8 months for 1,2-DCB, 1,4-DCB, TCBs and TeCBs and 2 years for QCB. All of the CBzs considered in this report are, therefore, likely to persist in soils under aerobic conditions (Government of Canada, 1993a, 1993b, 1993c, 1993d, 1993e).

5.2.1.3 Air

TeCBs and QCB have been identified in sediments from lakes in both temperate regions and northern Canadian environments (Eisenreich et al., 1989; Muir et al., 1995; Rawn et al., 2000b). The detection of the TeCBs and QCB in northern lake sediments in the absence of nearby sources indicates that these residues are a result of long-range atmospheric transport, and these CBzs, therefore, meet the criteria for persistence in air.

On the basis of the available information, it can be concluded that all of the CBzs of interest are persistent in soil and sediment according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999, and TeCBs and QCB are also persistent in air.

5.2.2 Bioaccumulation

BCFs reported on a whole-body basis for 1,2-DCB in rainbow trout (Oncorhynchus mykiss) ranged between 270 and 560 in laboratory studies (Government of Canada, 1993a). BCFs for 1,4-DCB in rainbow trout ranged between 370 and 1400 (Government of Canada, 1993b), and BCFs between 100 and 4000 have been reported for TCBs. Reported BAFs for TeCBs range between 1180 and 135 000 (Government of Canada, 1993c, 1993d). BAFs of 810 and 20 000 were reported for QCB in mussel (Mytilis edulis) and rainbow trout, respectively, although the BAF determined for earthworms (Eisenia andrei) was much higher (401 000) (Government of Canada, 1993e). Both 1,2-DCB and 1,4-DCB have an estimated log Kow of 3.4. The log Kow estimates for TCBs, TeCBs and QCB were 3.85–4.30, 4.5 and 5.0, respectively (Mackay et al., 1992).

On the basis of the available information, it is concluded that the TeCBs and QCB are bioaccumulative substances according to the criteria stipulated in the Persistence and Bioaccumulation Regulations of CEPA 1999.

5.3 Assessment Endpoints

The CBzs under investigation in this report (1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB) were assessed during the PSL1. At that time, it was determined that concentrations of these compounds in Canadian air and surface waters were not likely to cause adverse effects on aquatic biota or wildlife. The assessment endpoints of interest in this report, therefore, are adverse effects on populations of benthic and soil-dwelling species.

5.4 Environmental Risk Characterization

5.4.1 Sediment

Effects on growth and reproduction were observed in two freshwater species, Hexagenia spp. and Tubifex tubifex, respectively, following exposure to sediment spiked with 1,2-DCB, 1,4-DCB, 1,2,3-TCB and 1,2,4,5-TeCB (Day et al., 1995). The TCB and TeCB isomers studied were assumed to represent all members of each homologue group. Effects on growth and reproduction were observed at the highest concentration level of each exposure series tested (Table 5), with the exception of 1,2-DCB and 1,2,4,5-TeCB for Hexagenia spp. and Tubifex tubifex, respectively. In these studies, no effect was observed following exposure to spiked sediments. Based on these data limitations, estimates of traditional endpoint values (e.g., LC50, EC50, LOEC, etc.) were not possible. For QCB, where no measurements were made, effect concentrations were extrapolated using the measured molar effect concentrations determined for each of the other CBzs. The EqP estimates for benthic organisms were within the range reported for Hexagenia spp. and Tubifex tubifex by Day et al. (1995) for 1,4-DCB and 1,2,4,5-TeCB (Table 5). Measured effects concentrations for 1,2-DCB and 1,2,3-TCB were higher than the EqP estimates. Based on these results, the EqP results were taken to be the conservative CTVSED for 1,2-DCB (1382 µg/g OC), 1,4-DCB (1005 µg/g OC), TCBs (1637 µg/g OC), TeCBs (2846 µg/g OC) and QCB (2500 µg/g OC) for freshwater benthos (Table 9).

Doe et al. (1995) reported LOECs for marine organisms exposed to sediment spiked with 1,2-DCB, 1,4-DCB, 1,2,3-TCB and 1,2,4,5-TeCB. The LOEC for QCB was estimated by extrapolating the molar-based values measured for the DCBs, 1,2,3-TCB and 1,2,4,5-TeCB. EqP values were also determined for marine benthic organisms, and comparisons between measured and calculated values were made. The more conservative estimate was taken to be the CTV for marine benthic organisms. The measured LOECs reported by Doe et al. (1995) for marine benthos were taken to be the CTVSEDs for 1,2-DCB (1127 µg/g OC) and TeCBs (1582 µg/g OC). The median LOEC for QCB, based on the molar estimates reported by Doe et al. (1995), was taken to be the CTVSED (3080 µg/g OC) (Table 9). The EqP value was used as the CTVSED for 1,4-DCB (4999 µg/g OC) and TCBs (504 µg/g OC) (Table 9) in marine systems.

5.4.1.1 Determination of Estimated No-Effects Values (ENEVSEDs)

The CTVSEDs determined for each of the CBzs under consideration in this report, in both freshwater and marine sediments, were divided by an application factor, resulting in an ENEVSED to convert chronic lowest-reported-effect levels to no-effect concentrations and to account for extrapolation from laboratory to field conditions and inter- and intraspecies variability (Environment Canada, 1997) (Table 10). An acute to chronic ratio (ACR) of 3:1 (Carlson and Kosian, 1987) was applied to the marine CTVSED for 1,2-DCB, 1,4-DCB, TeCBs and QCB because these CTVSEDs were based on acute studies. Although the freshwater CTVSED for TCB was based on an acute endpoint (LC90) (Lay et al., 1985), the test was performed over a 21-day period using daphnids; therefore, it may be considered a chronic study, and an ACR was thus not deemed necessary. An additional application factor of 10 was used for TeCBs and QCB since these have been shown to be persistent and bioaccumulative compounds according to the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada, 2000) and since there are special concerns about the long-term cumulative effects of these types of substances. Freshwater ENEVSEDs ranged from 25 µg/g OC for QCB to 164 µg/g OC for TCBs. Marine ENEVSEDs ranged from 5 µg/g OC for TeCBs to 167 µg/g OC for 1,4-DCB.

5.4.1.2 Determination of Risk Quotients for Sediments

The maximum reported CBz concentrations observed in Canadian sediments were from the St. Clair River (Table 3). These levels were taken to represent the EEVs (EEVSEDs) for freshwater benthic organisms because they are suitable representatives of conservative estimates for freshwater Canadian sediments. Risk quotients were calculated for each CBz of interest in this report using the following relationship, with 1,2-DCB as an example:

Quotient :

  • = EEV / ENEV
  • = 52 µg/g OC / 138 µg/g OC
  • = 0.4

1,2-DCB was found to have a risk quotient less than 1 (Table 11), indicating that 1,2-DCB concentrations are below a level for concern in Canadian freshwater sediments. Risk quotients of greater than 1 were found for 1,4-DCB in approximately 25% of the samples from St. Clair River, while risk quotients of greater than 1 for TCBs were found in 18% of the St. Clair River samples. Risk quotients for TeCBs and QCB exceeded a value of 1 in 28% (11 of 39) and 23% (9 of 39) of samples collected from the St. Clair River, respectively. In general, the highest risk quotients were determined for sediment samples collected within 0.5 km of the Dow Chemical Canada 1st Street Sewer.

Based on the results of the risk analysis for freshwater benthic organisms using conservative EEV data, 1,2-DCB is present in Canadian sediments at concentrations not expected to result in effects on freshwater benthic organisms. 1,4-DCB, TCBs, TeCBs and QCB are present at concentrations in Canadian sediments such that effects on freshwater benthic organisms are possible. The very high concentrations used as freshwater EEVSEDs have been reported at only one site in Canada, the St. Clair River, which indicates that these concentrations do not represent the majority of Canadian sediments.

The lower chlorinated CBzs (e.g., 1,2-DCB, 1,4-DCB and TCBs) have not been reported in sediments from remote areas and do not meet the criteria for both persistence and bioaccumulation in the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada, 2000). A less conservative approach to determining EEVSEDs for 1,4-DCB and TCBs in freshwater sediments was, therefore, employed, which takes into account the distribution of concentrations in Canada. Examination of data in Table 3, which summarizes the highest concentrations of 1,4-DCB and TCBs reported in Canadian sediment,Footnote 2 indicates that maximum concentrations in freshwater sediment from other locations in Canada are consistently below ENEVs (i.e., risk quotients <1.0). Furthermore, concentrations of 1,4-DCB and TCBs are less than the ENEVs in approximately 75% of the samples collected from the most highly contaminated 2-km stretch of the St. Clair River. For example, risk quotients calculated using median concentrations for 1,4-DCB and TCBs for this 2-km stretch are 0.4 and 0.5, respectively. This indicates that, in the vast majority of cases, the presence of 1,4-DCB and TCBs in Canadian freshwater sediments will not likely result in adverse effects on freshwater benthic organisms.

The higher chlorinated CBzs, TeCBs and QCB, however, have been reported in freshwater sediments from northern regions of Canada, indicating that they are subject to long-range transport and deposition. TeCBs and QCB also meet the persistence and bioaccumulation requirements of the Persistence and Bioaccumulation Regulationsof CEPA 1999 (Government of Canada, 2000). These factors necessitate the use of a conservative approach in the assessment of TeCBs and QCB.

Risk quotients for all CBzs of interest in marine systems were less than 1, and, therefore, effects on marine benthic organisms would not be anticipated based on current CBz concentrations in marine sediments.  

5.4.2 Soil

Limited effects data for soil-dwelling organisms exist, and no new data were produced during this review of the CBzs. The lowest concentrations causing an effect were, therefore, taken to be the CTVSOIL estimates for each CBz under examination in this report (Table 12).

5.4.2.1 Determination of Estimated No-Effects Values (ENEVSOILs)

The CTVSOILs determined for 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB in soils were divided by an application factor of 10 to convert chronic lowest-reported-effect levels to no-effect concentrations and to account for extrapolation from laboratory to field conditions and inter- and intraspecies variability (Environment Canada, 1997). An additional factor of 3 was used for all CBzs to account for the limited data on effects on terrestrial organisms. For the TeCBs and QCB, an additional application factor of 10 was used because these compounds are persistent and bioaccumulative. The resulting ENEVSOILs range from 0.6 µg/g OC for 1,2,4,5-TeCB to 157 µg/g OC for 1,2- and 1,4-DCB (Table 12).

5.4.2.2 Determination of Risk Quotients for Soils

The maximum 1,2-DCB and 1,4-DCB concentrations estimated for Canadian sludge-amended soil, based on sludge concentrations reported in Webber and Nichols (1995), were taken to be the EEV for Canadian soil. Individual isomer concentrations reported by Ding et al. (1992) (Table 4) were taken to be the EEVSOILs for TCBs and TeCBs. This enabled the calculation of ENEVSOIL values for individual isomers (Table 12). Risk quotients, therefore, were calculated using EEVSOIL and ENEVSOIL determinations for the individual isomers where data were available (Table 13).

The concentrations used to calculate risk quotients for TCBs, TeCBs and QCB are representative of a highly industrialized area in the United States and, therefore, are expected to overestimate the majority of Canadian soil concentrations. These data were selected to represent worst-case conditions that may be observed near highly industrialized areas in Canada. The risk quotients for all CBzs under consideration were below 1, despite these extreme conditions. This indicates that CBz concentrations would not result in effects on Canadian soil-dwelling organisms.

5.5 Sources of Uncertainty

During the freshwater sediment exposure studies with 1,2-DCB, 1,4-DCB, TCBs and TeCBs, effects were observed at the highest exposure level only; therefore, the EqP method was used as an additional line of evidence to estimate effect concentrations for freshwater benthic organisms. Effects on benthic organisms as a result of exposure to QCB in sediment, in both freshwater and marine systems, were not reported; therefore, QCB effect levels were estimated by relating effect concentrations of all the other CBzs tested to that of QCB, on a molar basis.

There are a number of sources of uncertainty in this environmental risk assessment. Soil concentration data for the Canadian environment were limited to one study in which highly industrialized areas were not emphasized. Representative data, therefore, were taken from a highly industrialized region in the United States, where such data were available. Additionally, there are limited data in the literature for effects of CBzs on soil-dwelling organisms. An additional application factor was applied to the effects data to account for the small data set in the literature for effects on soil-dwelling organisms.

Although current scientific methods are unable to accurately predict the effects of persistent and bioaccumulative substances on the environment, these substances have been dealt with in a conservative manner in this assessment. Persistent substances can remain in the environment for long periods of time, thereby increasing the probability and duration of exposure relative to compounds that do not persist in the environment. Additionally, substances that are subject to long-range transport are of particular concern, because remote and/or cold regions, such as the Canadian Arctic, can act as a sink for such contaminants. Bioaccumulative substances have the potential to biomagnify through the food chain. Even releases of extremely low concentrations of persistent and bioaccumulative substances can lead to accumulations in organisms having the potential – either alone or in combination with other similar substances - to cause adverse effects on organisms that are continually exposed to them over long periods; therefore, an additional application factor of 10 was applied to the TeCBs and QCB, which are persistent and bioaccumulative substances.

Non-polar halogenated organic compounds occur together in sediments near industrial effluents (Bedard and Petro, 1997). Narcosis is the mode of action of many of these compounds, including CBzs. As a result, a cumulative effect on exposed organisms would be anticipated (McCarty et al., 1992). Toxicity studies using St. Clair River sediments confirmed that exposure to multiple narcotic substances, including QCB, was correlated with increased lethality of mayfly (Hexagenia limbata) and midge (Chironomus tentans) (Bedard and Petro, 1997). Sediments near the Dow Chemical sewer outfall, where multiple non-polar chlorinated organic compounds were detected, are characterized by low abundance and poor benthic invertebrate diversity (Bedard and Petro, 1997).

Losses of the volatile compounds, such as TCBs, were observed in toxicity studies relative to field samples (Bedard and Petro, 1997), indicating that many toxicity studies may result in an underestimation of toxicity.

ENEVSEDs determined in this assessment were compared with hazard-based assessments and effect threshold studies. Quantitative structure–activity relationships and EqP models were used by van Leeuwen et al. (1992) to estimate concentrations of CBzs that would not be expected to affect 95% of the species in benthic communities (marine or freshwater). These concentrations, known asHC5s because they are expected to be hazardous concentrations for 5% of the exposed species, were estimated to be approximately 107 µg/g OC for both 1,2- and 1,4-DCB and 115 µg/g OC for TCBs. These values are within an order of magnitude of the freshwater and marine ENEVSEDs developed in this analysis. The HC5 for TCBs (115 µg/g OC) compares extremely well with the ENEVSEDs developed for TCBs in the present assessment (164 µg/g OC [freshwater] and 50 µg/g OC [marine]) (Table 10). The HC5s estimated for TeCBs and QCB (119 µg/g OC and 120 µg/g OC, respectively), however, were 1–2 orders of magnitude higher than the freshwater and marine ENEVSEDs determined in this analysis (Table 10). The lower ENEVSED values for the TeCBs and QCB determined in this assessment are a result of the conservative approach taken with persistent and bioaccumulative substances. Using another approach, marine Apparent Effect Thresholds (AETs) were developed in the Puget Sound area of Washington State, using data from paired sediment chemistry–effects measurements performed with field sediment samples (Barrick et al., 1988). AET estimates for 1,2-DCB, 1,4-DCB and TCBs (2 µg/g OC, 16 µg/g OC and 3 µg/g OC, respectively) were lower by an order of magnitude or more than marine ENEVSEDs developed in this assessment. The AET method has, however, been criticized because the results may be strongly influenced by the presence of unmeasured contaminants (Chapman, 1989).       

5.6 Conclusions

Concentrations of 1,2-DCB, 1,4-DCB, TCBs, TeCBs and QCB in Canadian soil are unlikely to be causing harm to populations of soil-dwelling organisms.  However, it is possible that concentrations of 1,4-DCB, TCBs, TeCBs and QCB in sediment from the St. Clair River near Sarnia may be harming benthic organisms.

Of the CBzs under consideration, only TeCBs and QCB meet the criteria for both persistence and bioaccumulation specified in the Persistence and Bioaccumulation Regulations of CEPA 1999 (Government of Canada, 2000).  There are special concerns about persistent and bioaccumulative substances. Persistent substances can remain in the environment for long periods of time, increasing the probability and the duration of exposure.  In addition persistent substances are subject to long-range transport, which results in low-level, widespread contamination  Bioaccumulative substances have the potential to biomagnify, and consequently releases of extremely low concentrations of persistent and bioaccumulative substances may - either alone or in combination with other similar substances - cause severe adverse effects.

CEPA 64(a): Based on available data, it is concluded that 1,2-dichlorobenzene, 1,4- dichlorobenzene and the trichlorobenzenes are not entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity. Therefore 1,2-dichlorobenzene, 1,4-dichlorobenzene and the trichlorobenzenes are not considered “toxic” as defined under Paragraph 64(a)of CEPA 1999.

Based on available data, tetrachlorobenzenes and pentachlorobenzene are entering the environment in a quantity or concentration or under conditions that have or may have an immediate or long-term harmful effect on the environment or its biological diversity and that the tetrachlorobenzenes and pentachlorobenzene are considered “toxic,” as defined under Paragraph 64(a)of CEPA 1999.

Back to Top

6. Considerations for Follow-up

It is recommended that both TeCBs and QCB be added to the List of Toxic Substances (Schedule I) of CEPA 1999.

QCB and TeCBs are persistent, bioaccumulative, predominantly anthropogenic and are considered “toxic” under Paragraph 64(a) of CEPA 1999, and as such, meet the criteria for Track 1 substances under the Toxic Substances Management Policy.  Therefore QCB and TeCB should be subject to virtual elimination of releases to the environment.  Since there is currently no commercial demand for QCB and TeCBs in Canada, options to prevent their reintroduction into the Canadian market should be explored.

Back to Top

7. References

Allen-Gil, S.M., C.P. Gubala, R. Wilson, D.H. Landers, T.L. Wade, J.L. Sericano and L.R. Curtis. 1997. Organochlorine pesticides and polychlorinated biphenyls (PCBs) in sediments and biota from four US Arctic lakes. Arch. Environ. Contam. Toxicol. 33: 378–387.

Barrick, R., S. Becker, L. Brown, H. Beller and R. Pastorok. 1988. Sediment quality values refinement: 1988 update and evaluation of Puget Sound. Vol. 1. Prepared by Tetra Tech Inc., Bellevue, Washington, for Puget Sound Estuary Program, U.S. Environmental Protection Agency, September 1988. 74 pp.

Bedard, D. and S. Petro. 1992. St. Clair River study, 1990 laboratory sediment bioassay report. Water Resources Branch, Ontario Ministry of Environment and Energy, Toronto, Ontario. 20 pp. (Internal Technical Report).

Bedard, D. and S. Petro. 1997. Laboratory sediment bioassay report on upper St. Clair River sediments in the vicinity of industrial point sources 1994 & 1995. Standards Development Branch, Ontario Ministry of Environment and Energy, Etobicoke, Ontario. 76 pp.

Beurskens, J.E., C.G. Dekker, J. Jonkhoff and L. Pompstra. 1993a. Microbial dechlorination of hexachlorobenzene in a sedimentation area of the Rhine River. Biogeochemistry 19: 61–81.

Beurskens, J.E., C.G. Dekker and H. van den Heuvel. 1993b. High levels of chlorinated aromatic compounds in deep Rhine sediments with special reference to microbial transformations. Land Degrad. Rehabil. 4: 367–371.

Beurskens, J.E., C.G. Dekker, H. van den Heuvel, M. Swart and J. de Wolf. 1994. Dechlorination of chlorinated benzenes by an anaerobic microbial consortium that selectively mediates the thermodynamic most favourable reactions. Environ. Sci. Technol. 28(4): 701–706.

Bosma, T.N., J.R. van der Meer, G. Schraa, M.E. Tros and A.J. Zehnder. 1988. Reductive dechlorination of all trichlorobenzene and dichlorobenzene isomers. FEMS Microbiol. Ecol. 53: 223–229.

Burkhard, L., B.R. Sheedy, D.J. McCauley and G.M. DeGraeve. 1997. Bioaccumulation factors for chlorinated benzenes, chlorinated butadienes and hexachloroethane. Environ. Toxicol. Chem. 16: 1677–1686.

Calamari, D.,S. Galassi and F. Setti. 1982. Evaluating the hazard of organic substances on aquatic life: the paradichlorobenzene example. Ecotoxicol. Environ. Saf. 6: 369–378.

Calamari, D.,S. Galassi, F. Setti and M. Vighi. 1983. Toxicity of selected chlorobenzenes to aquatic organisms. Chemosphere 12(2): 253–262.

Camford Information Services. 1991. Chlorobenzene. CPI Product Profile. Don Mills, Ontario. 4 pp.

Canadian Chemical Producers’ Association. 1999.Reducing emissions 7. 1998 emissions inventory and five year projections. A Responsible Care Initiative. Ottawa, Ontario.

Carlson, A.R. and P.A. Kosian. 1987. Toxicity of chlorinated benzenes to fathead minnows (Pimephales promelas). Arch. Environ. Toxicol. Chem. 16: 129–135.

CEU (Commission of the European Union). 1995. Technical guidance document on environmental risk assessment for existing substances in the context of Commission Regulation XX/94 in accordance with Council Regulation (EEC) No. 793/93 on the evaluation and control of existing substances. Chapter 3. 82 pp.

Chapman, P.M. 1989. Current approaches to developing sediment quality criteria. Environ. Toxicol. Chem. 8: 589–599.

Chapman, P.M., J. Downie, A. Maynard and L.A. Taylor. 1996. Coal and deodorizer residues in marine sediments -- contaminants or pollutants? Environ. Toxicol. Chem. 15(5): 638–642.

Clark, J.R., J.M. Patrick, Jr., J.C. Moore and E.M. Lores. 1987. Waterborne and sediment-source toxicities of six organic chemicals to grass shrimp (Palaemonetes pugio) and amphioxus (Branchiostoma caribaeum). Arch. Environ. Contam. Toxicol. 16: 401–407.

Day, K.E., D. Milani, S.M. Backus and M.E. Fox. 1995. The toxicity of 1,2-dichlorobenzene, 1,4-dichlorobenzene, 1,2,3-trichlorobenzene and 1,2,4,5-tetrachlorobenzene to two species of freshwater benthic invertebrates in spiked-sediment toxicity tests. Final report, November 1995. National Water Research Institute, Burlington, Ontario. 13 pp.

DeLuca, B. and M. Fox. 1995. Personal communication. Unpublished analytical data for di- and trichlorobenzenes in archived sediment samples provided by P. Kauss of the Ontario Ministry of Environment and Energy. National Water Research Institute, Burlington, Ontario.

DeWolf, W., J.H. Canton, J.W. Deneer, R.C. Wegman and J.L. Hermens. 1988. Quantitative structure–activity relationships and mixture toxicity studies of alcohols and chlorohydrocarbons: reproducibility of effects on growth and reproduction of Daphnia magna. Aquat. Toxicol. 12: 39–49.

Ding, W.-H., K.M. Aldous, R.G. Briggs, H. Valente, D.R. Hilker, S. Connor and G.A. Eadon. 1992. Application of multivariate statistical analysis to evaluate local sources of chlorobenzene congeners in soil samples. Chemosphere 25(5): 675–690.

Di Toro, D.M., C.S. Zarba, D.J. Hansen, R.C. Swartz, C.E. Cowan, S.P. Pavlou, H.E. Allen, N.A. Thomas and P.R. Paquin. 1991. Technical basis for establishing sediment quality criteria for nonionic organic chemicals using equilibrium partitioning. Environ. Toxicol. Chem. 10: 1541–1583.

Doe, K.G., A.L. Huybers, S.J. Wade and J.D. Vaughan. 1995. Toxicity of 1,2-dichlorobenzene, 1,4-dichlorobenzene, 1,2,3-trichlorobenzene and 1,2,4,5-tetrachlorobenzene to marine invertebrates and bacteria in spiked sediment toxicity tests. Final report, November 1995. Environmental Quality Laboratory, Environment Canada, Dartmouth, Nova Scotia. 16 pp.

Eisenreich, S.J., P.D. Capel, J.A. Robbins and R. Bourbonniere. 1989. Accumulation and diagenesis of chlorinated hydrocarbons in lacustrine sediments. Environ. Sci. Technol. 23: 1116–1126.

Environment Canada. 1994. Canadian Environmental Protection Act. Priority Substances List. Supporting documentation for Priority Substances List assessment report: Trichlorobenzenes. Commercial Chemicals Branch, Environment Canada, Hull, Quebec. 77 pp.

Environment Canada. 1997. Environmental assessments of Priority Substances under the Canadian Environmental Protection Act. Guidance manual version 1.0 -- March 1997. Chemicals Evaluation Division, Commercial Chemicals Evaluation Branch, Hull, Quebec (Environmental Protection Series EPS/2/CC/3E).

EVS. 1992. Sediment and related investigations off the Macaulay and Clover Point sewage outfalls. Final report, September 1992. EVS Consultants, North Vancouver, British Columbia. 193 pp.

EVS. 1996. Investigation of effects of 1,4-dichlorobenzene exposure of juvenile polychaete worms. Final report for Capital Regional District. EVS Consultants, Victoria, British Columbia (EVS Project No. 3/073-23.1).

Fox, M.E., J.H. Carey and B.G. Oliver. 1983. Compartmental distribution of organochlorine contaminants in the Niagara River and the western basin of Lake Ontario. J. Great Lakes Res. 9(2): 287–294.

Gas Research Institute. 1995. Environmentally acceptable endpoints in soil: Risk-based approach to contaminated site management based on availability of chemicals in soil. Draft report. Environmental and Safety Research Group, April 1995.

Government of Canada. 1993a. Canadian Environmental Protection Act Priority Substances List Assessment Report: 1,2-Dichlorobenzene. Environment Canada and Health Canada, Ottawa, Ontario. 27 pp.

Government of Canada. 1993b. Canadian Environmental Protection Act Priority Substances List Assessment Report: 1,4-Dichlorobenzene. Environment Canada and Health Canada, Ottawa, Ontario. 30 pp.

Government of Canada. 1993c. Canadian Environmental Protection Act Priority Substances List Assessment Report: Trichlorobenzenes. Environment Canada and Health Canada, Ottawa, Ontario. 39 pp.

Government of Canada. 1993d. Canadian Environmental Protection Act Priority Substances List Assessment Report: Tetrachlorobenzenes. Environment Canada and Health Canada, Ottawa, Ontario. 42 pp.

Government of Canada. 1993e. Canadian Environmental Protection Act Priority Substances List Assessment Report: Pentachlorobenzene. Environment Canada and Health Canada, Ottawa, Ontario. 32 pp.

Government of Canada. 1993f. Canadian Environmental Protection Act. Priority Substances List. Supporting document. Environmental sections. 1,2-Dichlorobenzene. Commercial Chemicals Branch, Environment Canada, Hull, Quebec. 71 pp.

Government of Canada. 1993g. Canadian Environmental Protection Act. Priority Substances List. Supporting document. Environmental sections. 1,4-Dichlorobenzene. Commercial Chemicals Branch, Environment Canada, Hull, Quebec. 75 pp.

Government of Canada. 1993h. Canadian Environmental Protection Act. Priority Substances List. Supporting document. Environmental sections. Tetrachlorobenzenes. Commercial Chemicals Branch, Environment Canada, Hull, Quebec. 40 pp.

Government of Canada. 1993i. Canadian Environmental Protection Act. Priority Substances List. Supporting document. Environmental sections. Pentachlorobenzene. Commercial Chemicals Branch, Environment Canada, Hull, Quebec. 35 pp.

Government of Canada. 2000. Persistence and Bioaccumulation Regulations. Canada Gazette, Part II, 134: 607–611.

Hermens, J., H. Canton, P. Janssen and R. de Jong. 1984. Quantitative structure–activity relationships and toxicity studies of mixtures of chemicals with anaesthetic potency: Acute lethal and sublethal toxicity to Daphnia magna. Aquat. Toxicol. 5: 143–154.

Hulzebos, E.M., D.M. Ademe, E.M. Dirven-van Breemen, L. Henzen, W.A. van Dis, H.A. Herbold, J.A. Hoekstra, R. Baerselman and C.A. van Gestel. 1993. Phytotoxicity studies with Lactuca sativa in soils and nutrient solution. Environ. Toxicol. Chem. 12: 1079–1094.

Jacobs, L.W., G.A. O’Connor, M.A. Overcash, M.J. Zabik and P. Rygiewicz. 1987. Effects of trace organics in sewage sludges on soil–plant systems and assessing their risk to humans. In: A.L. Page, T.J. Logan and J.A. Ryan (eds.), Land application of sludge -- food chain implications. Lewis Publishers, Chelsea, Michigan. pp. 101–143.

Kan, A.T., G. Fu and M.B. Tomson. 1994. Adsorption/desorption hysteresis in organic pollution and soil/sediment interaction. Environ. Sci. Technol. 28: 859–867.

Kauss, P.B. 1995. Personal communication, September 7, 1995. Ontario Ministry of Environment and Energy, Etobicoke, Ontario.

Kenaga, E.E. and C.A. Goring. 1980. Relationship between water solubility, soil sorption, octanol–water partitioning, and concentration of chemicals in biota. In: J.G. Eaton, P.R. Parrish and A.C. Hendricks (eds.), Aquatic toxicology. American Society for Testing and Materials, Philadelphia, Pennsylvania. pp. 78–115 (ASTM STP 707).

Kovrig, M. 1996. Letter “Re: Tier I and Tier II Profiles,” addressed to D. Hogg, Ontario Ministry of Environment and Energy, Toronto, Ontario, sent June 11, 1996, by M. Kovrig, Recochem Inc., Brampton, Ontario.

Kuntz, K.W. 1993. Trends in contaminant levels in the Niagara River. State of the Environment Reporting Program, Environment Canada, Ottawa, Ontario. 12 pp. (State of the Environment Fact Sheet No. 93-2).

Lay, J.P., W. Schauerte, A. Muller, W. Klien and F. Korte. 1985. Long-term effects of 1,2,4-trichlorobenzene on freshwater plankton in an outdoor-model-ecosystem. Bull. Environ. Contam. Toxicol. 34: 761–769.

Mackay, D., W.Y. Shiu and K.C. Ma. 1992. Illustrated handbook of physical-chemical properties and environmental fate for organic chemicals. Vol. I. Monoaromatic hydrocarbons, chlorobenzenes and PCBs. Lewis Publishers, Boca Raton, Florida. 697 pp.

Marinucci, A.C. and R. Bartha. 1979. Biodegradation of 1,2,3- and 1,2,4-trichlorobenzene in soil and in liquid enrichment culture. Appl. Environ. Microbiol. 38(5): 811–817.

McCarty, L.S., D. Mackay, A.D. Smith, G.W. Ozburn and D.G. Dixon. 1992. Residue-based interpretation of toxicity and bioconcentration QSARs from aquatic bioassays: Neutral narcotic organics. Environ. Toxicol. Chem. 11: 917–930.

Mudroch, A. 1983. Distribution of major elements and metals in sediment cores from the western basin of Lake Ontario. J. Great Lakes Res. 9(2): 125–133.

Muir, D.C.G., N.P. Grift, W.L. Lockhart, P. Wilkinson, B.N. Billeck and G.J. Brunskill. 1995. Spatial trends and historical profiles of organochlorine pesticides in Arctic lake sediments. Sci. Total Environ. 160/161: 447–457.

Muir, D.C.G., A. Omelchenko, N.P. Grift, D.A. Savoie, W.L. Lockhart and G.J. Brunskill. 1996. Spatial trends and historical deposition of polychlorinated biphenyls in Canadian midlatitude and Arctic lake sediments. Environ. Sci. Technol. 30: 3609–3617.

Neuhauser, E.F., P.R. Durkin, M.R. Malecki and M. Anatra. 1986. Comparative toxicity of ten organic chemicals to four earthworm species. Comp. Biochem. Physiol. 83C(1): 197–200.

NPRI (National Pollutant Release Inventory). 1994. Summary report 1994, National Pollutant Release Inventory. Canadian Environmental Protection Act. Environment Canada. Minister of Supply and Services Canada (Cat. No. EN40-495-1/1-1994E).

NPRI (National Pollutant Release Inventory). 1995. Summary report 1995, National Pollutant Release Inventory. Canadian Environmental Protection Act. Environment Canada. Minister of Supply and Services Canada (Cat. No. EN40-495-1/1-1995E).

NPRI (National Pollutant Release Inventory). 1996. Summary report 1996, National Pollutant Release Inventory. Canadian Environmental Protection Act. Environment Canada. Minister of Supply and Services Canada (Cat. No. EN40-495-1/1-1996E).

NPRI (National Pollutant Release Inventory). 1997. Summary report 1997, National Pollutant Release Inventory. Canadian Environmental Protection Act. Environment Canada. Minister of Supply and Services Canada (Cat. No. EN40-495-1/1-1997E).

NPRI (National Pollutant Release Inventory). 1998. Summary report 1998, National Pollutant Release Inventory. Canadian Environmental Protection Act. Environment Canada. Minister of Supply and Services Canada (Cat. No. EN40-495-1/1-1998E).

Oliver, B.G. and J.H. Carey. 1986. Photodegradation of wastes and pollutants in aquatic environment. In: E. Pelizzetti and N. Serpone (eds.), Homogeneous and heterogeneous photocatalysis. D. Reidel Publishing Co., Dordrecht, Netherlands. pp. 629–650.

Oliver, B.G. and K.D. Nicol. 1982. Chlorobenzenes in sediments, water and selected fish from Lakes Superior, Huron, Erie and Ontario. Environ. Sci. Technol. 16: 532–536.

Oliver, B.G. and K.D. Nicol. 1983. Response to comment on “Chlorobenzenes in sediments, water and selected fish from Lakes Superior, Huron, Erie and Ontario.” Environ. Sci. Technol. 17: 505.

Oliver, B.G. and C.W. Pugsley. 1986. Chlorinated contaminants in St. Clair River sediments. Water Pollut. Res. J. Can. 21(3): 368–379.

OMEE (Ontario Ministry of Environment and Energy). 1994. Proposed guidelines for the clean-up of contaminated sites in Ontario, July, 1994. Toronto, Ontario.

Peijnenburg, W.J.G.M., M.J.’t Hart, H.A. den Hollander, D. van de Meent, H.H. Verboom and N.L. Wolfe. 1992. Reductive transformations of halogenated aromatic hydrocarbons in anaerobic water–sediment systems: kinetics, mechanisms and products. Environ. Sci. Technol. 11: 289–300.

Rawn, D.F.K., D.C.G. Muir, D.A. Savoie, G.B. Rosenberg, W.L. Lockhart and P. Wilkinson. 2000a. Historical deposition of PCBs and organochlorine pesticides to Lake Winnipeg (Canada). J. Great Lakes Res. 26: 3–17.

Rawn, D.F.K., W.L. Lockhart, P. Wilkinson, D.A. Savoie, G.B. Rosenberg and D.C.G. Muir. 2000b. Historical contamination of Yukon lake sediments by persistent organic pollutants (POPs): Influence of local sources and watershed characteristics. Sci. Total Environ. (in press).

Rogers, H.R., J.A. Campbell, B. Crathorne and A.J. Dobbs. 1989. The occurrence of chlorobenzenes and permethrins in twelve U.K. sewage sludges. Water Res. 23: 913–921.

Rutherford, L.A., K.E. Day, K.G. Doe, A. Huybers, P.A. Hennigar, G.R. Julien, S.L. Matthews, D. Milani, D. Vaughan and S. Wade. 1995. Environmental occurrence and toxicity of chlorobenzenes in freshwater and marine sediments. Environmental Protection Branch, Environment Canada, Dartmouth, Nova Scotia. October 1995. 35 pp.

Scheunert, I., E. Topp, A. Attar and F. Korte. 1994. Uptake pathways of chlorobenzenes in plants and their correlation with n-octanol/water partition coefficients. Ecotoxicol. Environ. Saf. 27: 90–104.

Tagatz, M.E., G.R. Plaia and C.H. Deans. 1985. Effects of 1,2,4-trichlorobenzene on estuarine macrobenthic communities exposed via water and sediment. Ecotoxicol. Environ. Saf. 10: 351–360.

Trapp, S., M. Matthies, I. Scheunert and E.M. Topp. 1990. Modelling the bioaccumulation of organic chemicals in plants. Environ. Sci. Technol. 24(8): 1246–1252.

U.S. EPA (Environmental Protection Agency). 1980a. Ambient water quality criteria for dichlorobenzenes. Office of Water Regulations and Standards, Criteria and Standards Division (PB-81-117525; EPA-440/5-80-039).

U.S. EPA (Environmental Protection Agency). 1980b. Ambient water quality criteria for chlorinated benzenes. Office of Water Regulations and Standards, Criteria and Standards Division (EPA-560/13-80-001).

van Gestel, C.A. and W.-C. Ma. 1988. Toxicity and bioaccumulation of chlorophenols in earthworms in relation to bioavailability in soil. Ecotoxicol. Environ. Saf. 15: 287–297.

van Gestel, C.A and W.-C. Ma. 1990. An approach to quantitative structure–activity relationships (QSARs) in earthworm toxicity studies. Chemosphere 21(8): 1023–1033.

van Gestel, C.A., W.-C. Ma and C.E. Smit. 1991. Development of QSARs in terrestrial ecotoxicology: earthworm toxicity and soil sorption of chlorophenols, chlorobenzenes and dichloroaniline. Sci. Total Environ. 109/110: 589–604.

van Leeuwen, C.J., P.T. Van Der Zandt, T. Aldenberg, H.J. Verhaar and J.L. Hermens. 1992. Application of QSARs, extrapolation and equilibrium partitioning in aquatic effects assessments. I. Narcotic industrial chemicals. Environ. Toxicol. Chem. 11: 267–282.

van Wezel, A.P., D.A.M. de Vries, D.T.H.M. Sijm and A. Opperhuizen. 1996a. Use of the lethal body burden in the evaluation of mixture toxicity. Ecotoxicol. Environ. Saf. 35: 236–241.

van Wezel, A.P., G. Cornelissen, J.K. van Miltenburg and A. Opperhuizen. 1996b. Membrane burdens of chlorinated benzenes lower the main phase transition temperature in dipalmitoyl-phosphatidylcholine vesicles: implications for toxicity by narcotic chemicals. Environ. Toxicol. Chem. 15: 203–212.

Walton, B.T., T.A. Anderson, M.S. Hendricks and S.S. Talmage. 1989. Physiochemical properties as predictors of organic chemical effects on soil microbial respiration. Environ. Toxicol. Chem. 8: 53–63.

Wang, M.-J. and K.C. Jones. 1994. Uptake of chlorobenzenes by carrots from spiked and sewage sludge-amended soil. Environ. Sci. Technol. 28(7): 1260–1267.

Wania, F. and D. Mackay. 1993. Global fractionation and cold condensation of low volatility organochlorine compounds in polar regions. Ambio 22:10–18.

Webber, M. 1994. Industrial organic compounds in selected Canadian municipal sludges and agricultural soils. Wastewater Technology Centre (operated by Rockcliffe Research Management Inc.), Burlington, Ontario. October 1994. 100 pp.

Webber, M. and J.A. Nichols. 1995. Organic and metal contamination in Canadian municipal sludges and a sludge compost. Wastewater Technology Centre (operated by Rockcliffe Research Management Inc.), Burlington, Ontario. February 1995. 169 pp.

Yonezawa, Y.,M. Fukui, S. Masunaga and Y. Urushigawa. 1994. Dechlorination of 1,2,4-trichlorobenzene in the sediment of Ise Bay. Chemosphere 28(12): 2179–2184.

Back to Top

Tables

Table 1. Summary of Information on Production and Uses of CBzs in CanadaTable Footnote 1
CBzProduced in CanadaPrimary applications
1,2-DCByesIndustrial cleaning solvents
1,4-DCByesAir fresheners/deodorizers
Moth and bird repellents
TCBsnoSolvents in textile industry
Chemical manufacturing
Transformer maintenance
TeCBsnoTransformer maintenance
QCBnoLaboratory reagent
1 Based on data reported by Camford Information Services (1991); data applicable to 1995.
Table 2. Median Concentration of CBzs in Sewage Sludge (µg/kg dw)
Country Reference  1,2-DCB1,4-DCBSTCBsSTeCBsQCB
CanadaWebber and Nichols (1995)MDLTable Footnote 1– 451Table Footnote 21–810Table Footnote 2NDTable Footnote 3NATable Footnote 4NATable Footnote 4
United StatesJacobs et al. (1987)645Table Footnote 52020Table Footnote 5404Table Footnote 5NDTable Footnote 3Not reported
United KingdomRogers et al. (1989)7900980078080MDLTable Footnote 1
United KingdomWang and Jones (1994)231011205586747
1 Below method detection limit.
2 Range of median values was determined for each of 12 sludge treatment plants.
3 Not detected.
4 Not analysed.
5 Data reported as means rather than medians.
Table 3. Recent OC-normalized CBz Concentrations, Reported as Median Values, in Canadian Sediments Near Point Sources (Range) (µg/g)
 1,2-
DCB
1,4-
DCB
STCBsSTeCBsQCB Year collected Reference
Industrial, particularly chemical manufacturing
St. Clair River near Sarnia, OntarioTable Footnote 1

1.5

(0.2

52)

37

(2

522)

25

(1

539)

3.5

(0.1

320)

12

(0.3

601)

1994DeLuca and Fox (1995); Kauss (1995)
Niagara River delta, Lake OntarioTable Footnote 2

0.64

(0.51

1.5)

5.2

(<2.2

37)

3.4

(2.5

9.8)

2.9

(2.0

9.1)

1.1

(0.7

3.4)

1981Fox et al. (1983)
Textile mills

Atlantic Canada

(3 sites)Table Footnote 1

<2.4

(<2.4)

<2.4

(<2.4)

<2.4

(<2.4)

<2.4

(<2.4)

<2.4

(<2.4)

1994Rutherford et al. (1995)
STPs 
Victoria, B.C.Table Footnote 1

0.06

(0.01

0.30)

1.0

(0.1

40)

0.02

(0.01

0.3)

1991EVS (1992)
Sarnia, OntarioTable Footnote 1

0.34

(0.13

0.83)

2.0

(0.40

7.5)

2.0

(0.20

4.9)

0.09

(0.07

0.42)

0.09

(0.07

0.42)

1994Kauss (1995)
Halifax, Nova ScotiaTable Footnote 1

<2.2

(<2.2)

0.4

(<0.1

16)

<2.2

(<2.2)

<2.2

(<2.2)

<2.2

(<2.2)

1994Rutherford et al. (1995)
Other Atlantic sites, 1994Table Footnote 3

<2.2

(<2.2)

<2.2

(<2.2)

<2.2

(<2.2)

<2.2

(<2.2)

<2.2

(<2.2)

1994Rutherford et al. (1995)
1 OC content ranged between 0.2% and 10.1%.
2 OC content assumed to be 3.5%, based on analyses of delta samples (Mudroch, 1983).
3 Fredericton, New Brunswick, and Berwick, Nova Scotia.
Table 4. Median OC-normalized CBz Concentrations in Soils (Range) (µg/g)
 1,2
DCB
1,4
DCB
1,3,5
TCB
1,2,4
TCB
1,2,3
TCB
1,2,
3,5/
1,2,
4,5
Te
CB
1,2,
3,4
Te
CB
QCBYear col-
lec-
ted
Refe-
rence
Near agricultural source (Canada)1<3.5 (<3.5)<3.5
(<3.5

4.5)
Not measured<3.5 (<3.5)Not
measured
Not
measured
Not
measured
Not measuredearly
1990s
Web-
ber (1994)
Near industrial source (Niagara Falls, NY)2Not measuredNot measuredNot measured0.127
(0.060

0.255)
0.027
(0.012

0.050)
0.231
(0.120

0.400)
0.156
(0.065

0.325)
0.052
(0.024

0.085)
late
1980s
Ding et al. (1992)
1 OC-normalized (OC content ranged between 0.1% and 3.8%).
2 Normalized assuming 2% OC content.
Table 5. Percentage of Freshwater Test Populations (Hexagenia spp. and Tubifex tubifex) Affected by CBz Exposure (µg/g OC normalized1) After 21-day and 28-day Exposures, Respectively (Day et al., 1995)
CBz  Hexagenia spp.Tubifex tubifex
Reduction in growth observed
(% of affected organisms)2 
ConcentrationReduction in number of young produced
(% of affected organisms)2 
Concentration
Nominal
(µg/g dw)
Initial
(µg/g OC)
Final
(µg/g OC)
Nominal
(µg/g dw)
Initial
(µg/g OC)
Final
(µg/g OC)
1,2-
DCB3
none50037892186750044481871
1,4-
DCB
2550036772346450031871573
1,2,3-
TCB
34500684514447250065565947
1,2,4,5-
TeCB4
1815031282012none15039813266
1 % OC = 3.93% ± 0.56%.
2 Statistical significance (P < 0.05) of effects, relative to solvent controls, demonstrated using Dunnett’s test.
3 No effect on growth of Hexagenia spp. was observed at the highest exposure concentration of 1,2-DCB, initial = 3789 µg/gOC, final = 218 µg/gOC.
4 No effect on growth of T. tubifex was observed at the highest exposure concentration of 1,2,4,5-TeCB, initial = 3981 µg/g OC, final = 3266 µg/g OC.
Table 6. Lowest Concentrations of CBzs Causing Effects on the Marine Amphipod Rhepoxynius Abronius (Doe et al., 1995)
CBzObserved % mortality1Nominal
(µg/g dw )
Initial
(µg/g OC2)
 Final
(µg/g OC2)
NOEC3
(µg/g OC2)
LOEC3
(µg/g OC2)
1,2-DCB231001127Not measured2891127
1,4-DCB315006121927313456121
1,2,3-TCB601003455Not measured8983455
1,2,4,5-TeCB25301582Not measured2541582
1 The lowest exposure concentration associated with mortality that is significantly different from acetone controls at 95% confidence level.
2 % OC = 0.55%.
3 NOEC = No-Observed-Effect Concentration; LOEC = Lowest-Observed-Effect Concentration.
Table 7. Lowest Effect Concentration Estimates in Sediment (µg/g OC) Based on Effects Data for Water Column Organisms, Calculated Using the EqP Method
CBz Kow1 FreshwaterMarine
EndpointDissolved concen-tration
(µg/L)
ReferenceEstimated Csed
(µg/g OC)
EndpointDissolved concen-tration
(µg/L)
ReferenceEstimated Csed (µg/g OC)
1,2-
DCB
2512

14-
day EC50

reproduction (Daph
nia)

550Calamari
et al.
(1983)
1382

96-
hour LC502

(mysid shrimp)

1970U.S. EPA (1980a)4949
1,4-
DCB
251228-
day LOEC reproduction (Daph
nia)
400Calamari
et al.
(1982)
1005

96-
hour LC502

(mysid shrimp)

1990U.S. EPA (1980a)4999
TCBs1258921-
day LC90(Daph
nia)
130Lay
et al.
(1985)
1637reduced colonization of sediment (molluscs)40Tagatz et al. (1985)504
Te
CBs
3162316-
day EC50 reproduction
(Daph
nia)
90De
Wolf
et al.
(1988)
2846

96-
hour LC502

(mysid
shrimp)

340U.S. EPA (1980b)10752
QCB10000016-
day EC502 reproduction
(Daph
nia)
25Her-
mens
et al.
(1984)
2500

96-
hour LC502

(mysid
shrimp)

160US EPA (1980b)16000
1 Mackay et al. (1992).
2 Nominal concentration.
Table 8. Lowest Effect Concentrations (µg/gOC) for Lettuce and Earthworms
CBzSpeciesEndpointConcentration
(µg/gOC)
Source
1,2-DCBno data47122Assumed equal to 1,4-DCB
1,4-DCB

earthworm

(E. andrei)

14-day LC5047122van Gestel et al. (1991)
1,2,3-TCB

lettuce

(Lactuca sativa)

14-day EC501 (growth)1193Hulzebos et al. (1993)
1,2,4-TCB

earthworm

(E. eugeniae)

14-day LC50125922Neuhauser et al. (1986)
1,3,5-TCBlettuce

7-day EC501

(growth)

10 6484Hulzebos et al. (1993)
1,2,3,4-TeCBlettuce14-day EC501 (growth)29634Hulzebos et al. (1993)
1,2,4,5-TeCBlettuce14-day EC501 (growth)1854Hulzebos et al. (1993)
QCB

earthworm

(L. rubellus)

14-day LC5041362van Gestel et al. (1991)
1 Determined using nominal concentrations.
2 4.9% OC.
3 0.84% OC.
4 1.08% OC.
Table 9. CTVSEDs Selected for Benthic Freshwater and Marine Organisms
CBz FreshwaterMarine
CTVSED
(µg/g OC)
Data typeCTVSED
(µg/g OC)
Data type
1,2-DCB1382EqP-based chronic EC501127Measured LOEC (R. abronius)
1,4-DCB1005EqP-based chronic LOEC4999EqP-based (acute effect)
TCBs1637EqP-basedLC90504EqP-based (chronic effect)
TeCBs2846EqP-based chronic EC501582Measured LOEC (R. abronius)
QCB2500EqP-based chronic EC503080Median of extrapolated LOEC (R. abronius)
Table 10. Application Factors and Derived ENEVSEDs for Benthic Organisms (Freshwater and Marine)
CBz  FreshwaterMarine
CTVSED
(µg/g OC
Factors appliedENE
VSED
(µg/gOC
CTVSED
(µg/g OC
Factors appliedENE
VSED
(µg/gOC
chronic to no effectsP & B1ACRchronic to no effectsP & B1
1,2-
DCB
138210138112731038
1,4-
DCB
1005101014999310167
TCBs1637101645041050
TeCBs28461010291582310105
QCB250010102530803101010
1 P&B = persistence and bioaccumulation.
Table 11. Risk Quotients for Benthic Organisms, Based on Maximum EEVs (EEVSEDs) for Canadian Sediments
CBz FreshwaterMarine
Maximum EEV (µg/g OC)ENEV (µg/g OC)QuotientMaximum EEV (µg/g OC)ENEV (µg/g OC)Quotient
1,2-DCB521380.4<2.238<0.06
1,4-DCB5221015.2401670.24
TCBs5391643.3<2.250<0.05
TeCBs3202911<2.25<0.44
QCB6012524<2.210<0.22
Table 12. Determination of ENEVSOILs for Soil-dwelling Organisms
CBz CTVSOIL
(µg/g OC
Factor appliedENEVSOIL
(µg/g OC
Limited dataChronic to no effectsPersistence and bioaccumulation
1,2-DCB4712310157
1,4-DCB4712310157
1,2,3-TCB1193104.0
1,2,4-TCB259231086
1,2,3,4-TeCB2963310109.9
1,2,4,5-TeCB185310100.62
QCB41363101014
Table 13. Risk Quotients for Terrestrial Organisms, Based on Maximum EEVs (EEVSOILs) for Canadian Soils
CBzMaximum EEV
(µg/g OC)
Data source for maximum EEVENEVSOIL
(µg/g OC)
Quotient
1,2-DCB0.42Calculated from Webber and Nicols (1995)1570.003
1,4-DCB0.87Calculated from Webber and Nicols (1995)1570.006
1,2,3-TCB0.05Ding et al. (1992)4.00.01
1,2,4-TCB0.25Ding et al. (1992)860.003
1,2,3,4-TeCB0.33Ding et al. (1992)9.90.03
1,2,4,5-/1,2,3,5-TeCB0.40Ding et al. (1992)0.620.65
QCB0.09Ding et al. (1992)140.006

Back to Top

Appendix A

Persistence and Bioaccumulation Criteria as Defined in the Persistence and Bioaccumulation Regulations of CEPA 1999
PersistenceaBioaccumulationb 
MediumHalf-life

Air

Water

Sediment

Soil

≥ 2 days  or is subject to atmospheric transport from its source to a remote area

≥ 182 days

≥ 365 days

≥ 182 days

BAFc ≥ 5000;

BCFd ≥ 5000;

log Kowe ≥ 5

a  A substance is persistent when at least one criterion is met in any one medium.
b When the BAF of a substance cannot be determined in accordance with generally recognized methods, then the BCF of a substance will be considered, however, if neither its BAF nor its BCF can be determined with recognized methods, then the log Kow will be considered.
c Bioaccumulation factor means the ratio of the concentration of a substance in an organism to the concentration in water, based on uptake directly from the surrounding medium and food.
d Bioconcentration factor means the ratio of the concentration of a substance in an organism to the concentration in water, based only on uptake directly from the surrounding medium.
e octanol-water partition coefficient means the ratio of the concentration of a substance in an octanol phase to the concentration of the substance in the water phase of an octanol-water mixture

Back to Top

Appendix B: Search Strategy - New Information for the Assessment of "Toxic" to the Environment Under Paragraph 64 (a) of CEPA 1999

To identify relevant information on Canadian production, importation, use, and environmental release, searches of the NPRI (National Pollutant Release Inventory, Environment Canada), the ARET (Accelerated Reduction/Elimination of Toxics, Environment Canada) were performed.

Data relevant to the assessment of whether 1,2-dichlrobenzene, 1,4-dichlorobenzene, trichlorobenzenes, tetrachlorobenzenes or pentachlorobenzene are “toxic” to the environment under paragraph 64 (a) of CEPA 1999 were identified from existing review documents, published reference checks and on-line searches of the following databases up to December, 1999.  A search was conducted by name or CAS registry number in the following databases: Aquire, Registry of toxic effects of chemical substances (RTECS), Environment Abstracts, CAB abstracts, Current Contents, Poltox, Capulus Bib Abstracts, UnCover.


Footnotes

[1] The PSL 1 Assessment Report for the chlorobenzenes is available on the following websites: http://www.hc-sc.gc.ca/ewh-semt/pubs/contaminants/psl1-lsp1/chlorobenzene/index-eng.php
[2] Additional data on concentrations in sediment from less contaminated sites elsewhere in Canada are presented in Government of Canada (1993a, 1993b, 1993c, 1993d, 1993e).

Back to Top