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ARCHIVED - Polycyclic Aromatic Hydrocarbons - PSL1

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Environment Canada
Health Canada
1994
ISBN: 0-662-22209-1
Cat. No.: En40-215/42E

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Canadian Environmental Protection Act

Table of Contents

List of Tables

  • Table 1 Physical Properties of Polycyclic Aromatic Hydrocarbons
  • Table 2 Annual Atmospheric Emissions of Polycyclic Aromatic Hydrocarbons in Canada During 1990
  • Table 3 Concentrations of Polycyclic Aromatic Hydrocarbons in Ambient Air (ng/m3) in Canada
  • Table 4 Concentrations of Polycyclic Aromatic Hydrocarbons in Fresh Water (µg/L) in Canada
  • Table 5 Concentrations of Polycyclic Aromatic Hydrocarbons in Sediments (µg/g dry weight) in Canada
  • Table 6 Concentrations of Polycyclic Aromatic Hydrocarbons in Biota (µg/g) in Canada
  • Table 7 Most Sensitive Toxicity Endpoints Reported for Polycyclic Aromatic Hydrocarbons for Freshwater Organisms
  • Table 8 Field Population Studies in Fish
  • Table 9 Levels of Polycyclic Aromatic Hydrocarbons in B[a]P Equivalents/m in Canadian Ambient Air

List of Figures

  • Figure 1 Chemical Structures of Polycyclic Aromatic Hydrocarbons

Synopsis

Polycyclic aromatic hydrocarbons (PAHs) are emitted into the Canadian environment from both natural and anthropogenic sources. Forest fires, which release approximately 2000 tonnes of PAHs per year, are the single most important natural source of PAHs in Canada. However, since releases from that source are generally widely separated in time and space across the country, they do not result in continuous exposure in any specific area. Anthropogenic sources are numerous and result in emissions of PAHs into all environmental compartments. The greatest anthropogenic sources of PAHs released to the atmosphere are aluminum smelters [925 tonnes/year (t/yr)], with most of these emissions being released from smelters that use the Horizontal Stud Söderberg process. Major sources of PAHs to the aquatic and soil environments include creosote-treated products (up to 2000 t/yr), spills of petroleum products (76 t/yr), metallurgical and coking plants (4 t/yr), and deposition of atmospheric PAHs (amount unknown).

With the exception of some of the lighter compounds, which volatilize from water or soil, PAHs are relatively non-volatile and of low solubility in water. In air, Soil, and water, PAHs are mostly adsorbed to particulate matter, on which they are transported and can be degraded slowly. Since degradation is very slow in sediments, this medium is the major environmental sink for PAHs.

Polycyclic aromatic hydrocarbons have been measured in water from railway and utility ditches at concentrations higher than those recognized to cause harmful effects to biota under laboratory conditions. Concentrations of PAHs in surface waters elsewhere in the country are orders of magnitude lower, and are below concentrations that cause harmful effects to aquatic biota.

At several sites in Canada, including the tidal flats of Muggagh Creek in the vicinity of the tar ponds in Sydney Harbour, Nova Scotia, and near a former oil gasification plant at Kettle Creek, Port Stanley, Ontario, populations of aquatic organisms have been adversely affected by contamination by PAHs. Moreover, PAHs in sediments from Hamilton Harbour, Ontario, have been associated with mortality of sensitive aquatic invertebrates under laboratory conditions.

Under laboratory conditions, neoplastic effects such as liver tumours in aquatic organisms have been associated with exposure to PAHs. Field evidence also supports this association. In Vancouver harbour, neoplastic liver lesions were observed in up to 75% of the English sole caught in areas where sediments are highly contaminated by PAHs.

Polycyclic aromatic hydrocarbons generally do not absorb light of wavelengths critical to global warming. Unlike substances associated with depletion of stratospheric ozone, they are non-halogenated compounds of low to moderate persistence in the atmosphere. Given these properties and the low steady-state concentrations of PAHs in the atmosphere, they are not considered to contribute significantly to stratospheric ozone depletion, global warming, or ground-level ozone formation.

Owing principally to the limitations of the available data base, it was necessary to focus the human health assessment primarily on a small number of PAHs and consider exposure from air only.

Based primarily on the results of carcinogenicity bioassays in which PAHs have been administered to experimental animals by inhalation (benzo[a]pyrene only) and dermal application, and on supporting data, the five PAHs considered in the human health assessment (benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd]pyrene) have been classified as "Probably Carcinogenic to Humans" (i.e., substances for which there is believed to be some chance of adverse effects at any level of exposure). For such substances, estimated exposure is compared to quantitative estimates of cancer potency to characterize risk and provide guidance for further action, i.e., analysis of options to reduce exposure. Based on consideration solely of the potential effects of PAHs on human health by two limited approaches, the priority for analysis of options to reduce exposure would be moderate to high.

Based on these considerations, it has been concluded that polycyclic aromatic hydrocarbons are entering the environment in a quantity or concentration or under conditions that may have harmful effects on the environment. Polycyclic aromatic hydrocarbons are not considered to constitute a danger to the environment on which human life depends. The PAHs benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd]pyrene may constitute a danger in Canada to human life or health.

1.0 Introduction

The Canadian Environmental Protection Act (CEPA) requires the Minister of the Environment and the Minister of Health to prepare and publish a Priority Substances List that identifies substances, including chemicals, groups of chemicals, effluents, and wastes, that may be harmful to the environment or constitute a danger to human health. The Act also requires both Ministers to assess these substances and determine whether they are "toxic" as defined under Section 11 of the Act which states:

"...a substance is toxic if it is entering or may enter the environment in a quantity or concentration or under conditions

  1. having or that may have an immediate or long-term harmful effect on the environment;
  2. constituting or that may constitute a danger to the environment on which human life depends; or
  3. constituting or that may constitute a danger in Canada to human life or health."

Substances that are assessed as "toxic" as defined under Section 11 may be placed on Schedule I of CEPA. Consideration can then be given to developing regulations, guidelines, or codes of practice to control any aspect of these substances' life cycle, from the research and development stage through manufacture, use, storage, transport, and ultimate disposal.

The assessment of whether polycyclic aromatic hydrocarbons (PAHs) are "toxic", as defined under CEPA, was based on the determination of whether they enter or are likely to enter the Canadian environment in concentrations or quantities or under conditions that could lead to exposure of humans or other biota at levels that could cause adverse effects.

The term polycyclic aromatic hydrocarbons refers to the compounds made up of carbon and hydrogen atoms grouped into rings containing five or six carbon atoms. They are called "PAH derivatives" when an alkyl or other radical is introduced to the ring, and "heterocyclic aromatic compounds" (HACs) when any one carbon atom in a ring is replaced by a nitrogen, oxygen, or sulphur atom. Heterocyclic aromatic compounds and PAH derivatives were not considered in this report since little is known about their presence in the Canadian environment.

Polycyclic aromatic hydrocarbons constitute a class of chemical products that include about 100 individual compounds. Based on available data (NRCC, 1983; Bjørseth, 1983; Bjørseth and Ramdhal, 1985; LEI, 1990; EAG, 1990), a preliminary list was developed that included individual compounds representative of the diversity of chemical, environmental, and toxicological properties of PAHs. From this list, 13 compounds were selected for assessment. Sufficient data were available to assess presence in the Canadian environment and the effects on biota for 9 compounds: acenaphthene, anthracene, benz[a]anthracene, benzo[a]pyrene (B[a]P), fluoranthene, fluorene, naphthalene, phenanthrene, and pyrene.

Owing principally to limitations of available data, it was necessary to focus the human health assessment primarily on a small number of compounds for which the available data base would be adequate for evaluation of effects. In addition, carcinogenicity was considered to be the critical effect, since the doses of PAHs that induce non-neoplastic effects in experimental animals are considerably greater than those associated with increases in tumour incidence. The PAHs selected for human health assessment were benzo[a]pyrene (B[a]P), benzo[b]fluoranthene (B[b]F), benzo[j]fluoranthene ([B [j] F), benzo[k]fluoranthene (B[k]F), and indeno[1,2,3-cd] pyrene (IND).

These compounds were selected for the following reasons. (a) They have been classified in Groups 1 or 2 ("Carcinogenic to Humans" or "Probably Carcinogenic to Humans") of the classification scheme for this endpoint developed for the assessment of "toxic" under PARagraph 11(c) of CEPA (EHD, 1994) on the basis of data reviewed by the International Agency for Research on Cancer (IARC) (IARC, 1983). (b) Available data were considered sufficient to develop a preliminary estimate of potency for the most sensitive endpoint (i.e., carcinogenicity) in a target organ (i.e., the lung) relevant to one of the principal media of exposure in the general environment for which sufficient information was available to estimate exposure (i.e., air).

Available data were inadequate to develop estimates of exposure for, or carcinogenic potency of, a broad range of PAHs in the gastrointestinal tract following ingestion.

This focused approach to assessment of potential effects on human health was necessary principally because of the limited available data. Because of its limited scope, it should be considered as an interim approach only. For example, the five PAHs addressed here represent less than 25% (0.1 to 23.5%) of the 13 PAHs determined to be present at various locations in a recent survey of ambient air in Canada (Dann, 1992a). In addition, owing to the limitations of the data base, exposure from sources other than air has not been considered. Upon acquisition of additional data as suggested in Section 4.0, "Recommendations for Research", of this report, it may be possible to develop a more comprehensive approach to assessment.

Polycyclic aromatic hydrocarbons are a major component of creosote and, to a lesser extent, of waste crankcase oils. Creosote-impregnated waste materials and waste crankcase oils have been assessed separately as part of the CEPA Priority Substances List assessment program. These substances are therefore not considered directly in this assessment.

Creosote as a pesticide was not assessed in this report. The pesticide creosote is subject to the provisions of the Pest Control Products Act, and its regulatory status as a heavy-duty wood preservative is currently being re-evaluated (Agriculture Canada, 1992). As part of the pesticide re-evaluation process, Environment Canada and Health Canada will assess the potential effects on the environment and on human health resulting from these pesticidal uses of creosote.

Information relevant to the entry and environmental exposure and effects of PAHs was identified from reviews (Slooff et al., 1989; EAG, 1990; NRCC, 1983; NRC, 1983; Bjørseth, 1983; Bjøseth and Ramdhal, 1985; U.S. EPA, 1987; Ortech International, 1990; CCME, 1989; 1991), searches of data bases [Analytical Abstracts, EMBASE, Enviroline, POLLUTION ABSTRACT, International Register of Potentially Toxic Chemicals (IRPTC)], and from research reports appearing in scientific journals, and university and government publications available at the end of 1992. Works published since 1992 were considered when appropriate. In addition, to generate data essential for this assessment, studies were done by Environment Canada to characterize atmospheric concentrations of PAHs. Although much of the research on PAHs has been conducted outside of Canada, Canadian data on sources, use patterns, fate, and effects of PAHs on the environment were emphasized where available.

Data relevant to the assessment of whether PAHs are "toxic" to the environment obtained after the completion of the environmental sections of this report (i.e., May 1993) were not considered for inclusion. Similarly, data relevant to the assessment of whether PAHs are "toxic" to human health obtained after the completion of these sections of the report (i.e., July 1993) have not been incorporated.

Although review articles were consulted where considered appropriate, original studies that form the basis for the determination of "toxic" under CEPA (entry, and environmental exposure and effects) were critically evaluated by the following staff of Environment Canada. Original studies relevant to the estimation of carcinogenic potency for PAHs considered in this assessment have been critically evaluated by the following staff of Health Canada.

Staff of Environment Canada and Health Canada
Environment Canada Health Canada
R. Chénier
C. Fortin
A. Germain
C. Gonthier
B. Nadon
F. Perron
S. Ringuette
R. van Coillie
P.K.L. Chan
M.E. Meek

Quantitative estimates of carcinogenic potency were provided by S. Bartlett and M. Walker of Health Canada.

In this report, a synopsis that will appear in the Canada Gazette is presented. An extended summary of the technical information that is critical to the assessment is presented in Section 2.0. This information is presented in greater detail in supporting documentation that is available upon request. The assessment of whether PAHs are "toxic" as defined in CEPA is presented in Section 3.0.

As part of the review and approvals process established by Environment Canada for its contributions to Priority Substances List (PSL) assessments, sections of the supporting documentation dealing with presence in the Canadian environment were reviewed externally by J.F. Payne (Fisheries and Oceans Canada); sections dealing with sources and releases were reviewed by M. Lalonde (Alcan Ltd), G. Éthier and N. Sherwin (Canadian Petroleum Products Institute), and by representatives of provincial ministries of the environment, including D. Bezak (Manitoba), C. Bisco (Prince Edward Island), P. Blagden (Newfoundland), G. Chow (Saskatchewan), G. Houle (Quebec), L. Johnston (Northwest Territories), J. Knight (New Brunswick), C.S. Liu (Alberta), C. Mackinnon (Nova Scotia), T. Wakelin (British Columbia), and D. Yap (Ontario). Environmental components of the Assessment Report were reviewed by P. Hodson (National Water Research Institute), L.S. McCarty (L.S. McCarty Scientific Research and Consulting), and A.J. Nantel (Centre de toxicologie du Québec). The draft health-related sections of the supporting documentation and Assessment Report were approved by the Standards and Guidelines Rulings Committee of the Bureau of Chemical Hazards of Health Canada. The final Assessment Report was reviewed and approved by the Environment Canada/Health Canada CEPA Management Committee.

Copies of this Assessment Report and the unpublished supporting documentation are available upon request from:

Commercial Chemicals Branch
Environment Canada
14th Floor, Place Vincent Massey
351 St. Joseph Boulevard
Hull, Quebec
K1A 0H3

Environmental Health Centre
Room 104
Health Canada
Tunney's Pasture
Ottawa, Ontario
K1A 0L2

2.0 Summary of Information Critical to the Assessment of "Toxic"

2.1 Identity, Properties, Production, and Uses

Polycyclic aromatic hydrocarbons (PAHs) are organic substances made up of carbon and hydrogen atoms grouped into at least two condensed aromatic ring structures. These are divided into two categories: low molecular weight compounds composed of fewer than four rings and high molecular weight compounds of four or more rings. "PAH derivatives" include PAHs having an alkyl or other radical attached to a ring; "heterocyclic aromatic compounds" (HACs) include PAHs having any one carbon atom in a ring replaced by a nitrogen, oxygen, or sulphur atom. This report addresses only unsubstituted PAHs, containing only carbon and hydrogen atoms (Figure 1).

Pure PAHs are usually coloured, crystalline solids at ambient temperature. The physical properties of PAHs vary with their molecular weight and structure (Table 1). Except for naphthalene, they have very low to low water solubilities, and low to moderately high vapour pressures. Their octanol-water partition coefficients (Kow) are relatively high, indicating a relatively high potential for adsorption to suspended particulates in the air and in water, and for bioconcentration in organisms (NRCC, 1983; Slooff et al., 1989). Generally, PAHs only weakly absorb light of infrared wavelengths between 7 and 14 µm (Sadtler Research, 1982), the wavelengths usually absorbed by chemicals involved in global warming (Ramanathan, 1985).

Polycyclic aromatic hydrocarbons are present in the environment as complex mixtures that are difficult to characterize and measure. They are generally analyzed using gas chromatography coupled with mass spectrometry (GC-MS), or by using high pressure liquid chromatography (HPLC) with ultraviolet (UV) and fluorescence detectors (U.S. EPA, 1982; Bjørseth and Ramdhal, 1985; Slooff et al., 1989; Tardif and Chiu, 1992).

Polycyclic aromatic hydrocarbons are largely produced through the combustion or pyrolysis of organic matter either naturally or through human activity; very small amounts may be produced by diagenesis or biosynthesis (NRCC, 1983). Factors such as the type and quantity of fuel, the temperature and duration of combustion, and the availability of oxygen determine the nature and extent of PAH formation (NRCC, 1983).

Some PAHs are used commercially, but few data are available on the volumes used. Based on limited voluntary reporting of commercial activities for the Domestic Substances List, the following PAHs were reported as being produced or manufactured in Canada in 1986: naphthalene [more than 1000 tonnes (t) produced, 10 to 100 t imported]; anthracene (0.1 to 1 t produced, 1 to l0 t imported); fluorene (0.1 to 1 t produced, <1 t imported); and acenaphthene, benzo[a]pyrene, chrysene, and pyrene (<0.1 t imported for each) (Environment Canada, 1992). According to SRI (1985), roughly 70% of the naphthalene produced in the United States and Japan is used in the production of phthalic acid, and this is the expected major use of naphthalene in Canada.

Figure 1 Chemical Structures of Polycyclic Aromatic Hydrocarbons

Figure 1 Chemical Structures of Polycyclic Aromatic Hydrocarbons

Table 1 Physical Properties of Polycyclic Aromatic Hydrocarbons

Compound (C.A.S.N°)

Molecu-
lar weight

log
Kow

Water
solubility
at 25°C
(mg/L)

Melting
point (°C)

Vapour
pressure
at 25 °C
(mPa)

naphthalene (91-20-3)1

128.16

3.5

31.7

80.5

11 960

acenaphthene (83-32-9)1

154.21

4.33

3.42

95

594

fluorene (86-73-7)1

166

4.18

1.98

116.5

94.7

phenanthrene (85-01-8)1

178.24

4.5

1.29

101

90.7

anthracene (120-12-7)1

178.24

4.5

0.045

216

25

pyrene (129-00-0)1

202.26

4.9

0.135

156

91.3 x 10-6

fluoranthene (206-44-0)1

202.26

5.1

0.26

111

1328

benz[a]anthracene (56-66-3)1

228

5.6

0.0057

162

14.7 x 10-3

benz[a]pyrene (50-32-8)1,2

252.32

6.0

0.0038

179

0.37 x 10-6

benzo[b]fluoranthene (205-99-2)2

252.32

6.06

0.014

168

0.13 x 10-5 to 0.133 at 20°C

benzo[j]fluoranthene (205-82-3)2

252.32

 

 

166

 

benzo[k]fluoranthene (207-08-9)2

252.32

6.06

0.0043

217

2.8 x 10-9

indeno[1,2,3-cd]pyrene (193-39-5)2

276

6.4

0.00053

164

1.3 x 10-5

  1. compounds addressed in the assessment of effects on the environment
  2. compounds addressed in the assessment of effects on human health

References: ATSDR, 1990b; Merck Index, 1989; Slooff et al., 1989; CRC, 1988; NRCC, 1983; SRI, 1980; Santodonato et al., 1979, in Smith, 1984; Mackay and Shiu, 1977

Creosote is a complex and variable mixture of compounds distilled from coal tar. Seventeen PAHs, including the 13 compounds addressed in this report, account for about 63% of the creosote mixture (Environment Canada, 1988). Although Canadian production of creosote has been as high as about 45 x l06 L/yr at the end of the 1 940s, production has dropped to 12 to 14 x106 L (20 000 t) in 1990 (Envirochem, 1991). Creosote is used mainly as a wood preservative for treating railway ties and marine and freshwater pilings. Transmission and telephone poles have not been treated with creosote in Canada since the mid-1970s.

Coal derivatives and other petroleum products contain varying concentrations of PAHs. The concentration of PAHs (14 compounds, including 6 of the 13 considered in this report) in unleaded gasoline was determined to be 0.0054% (Westerholm et al., 1988). On the basis of reported use of gasoline in Canada in 1989 [35 000 ML (megalitres); (Statistics Canada, 1989)], it is estimated that approximately 1500 t of PAHs are present in the gasoline supply, annually.

2.2 Entry into the Environment

Polycyclic aromatic hydrocarbons may be released into the environment through natural phenomena such as forest fires, volcanic eruptions, diagenesis, and biosynthesis. Although PAHs are naturally present in coal derivatives and petroleum (NRCC, 1983), human activities are considered to be a major source of release of PAHs to the environment (Neff, 1979; NRC, 1983; NRCC, 1983; Bjørseth and Ramdhal, 1985; Slooff et al., 1989). The distribution and magnitude of certain emissions of PAHs are related to human population density (residential heating, transportation); however, others depend on the availability of power (aluminum smelters) or on the presence of natural resources [open air fires and agricultural burning, sawmill residue incinerators (teepee burners)].

Estimates for atmospheric emissions of PAHs in Canada were based on point and diffuse sources from 45 source-sectors (LGL, 1993). Estimated emissions of PAHs from Canadian sources for 1990 are outlined in Table 2. Forest fires represented the single largest source of PAHs to the environment, releasing about 2010 t of PAHs into the atmosphere, or 47% of the total atmospheric emissions inventoried. The aluminum smelting industry was the second largest source of atmospheric emissions of PAHs, accounting for 21% (925 t). In Canada, there are three processes used in aluminum smelting, all using coal tar pitch in the reduction of alumina: the Horizontal Stud Söderberg (HSS) process, which is the oldest one, the Vertical Stud Söderberg (VSS) process, and the pre-baked anode (PB) process. Through plant modernization programs, the Canadian aluminum smelting industry reduced RSS-related PAH emissions by 53% from 1983 to 1990. Nevertheless, although the HSS process accounted for only 20% of the total Canadian production of aluminum in 1990, it accounted for 75% of the PAHs emitted by the aluminum industry (LGL, 1993).

Other important sources of PAHs to ambient air include emissions from: residential wood heating (474 t/yr); agricultural burning and open air fires (358 t/yr); the incineration of wood residues by saw mills in teepee burners (249 t/yr); and transportation (201 t/yr).

The sources of PAHs that enter water and soil are varied and include: dispersion from creosoted materials (Wan, 1991; 1993); accidental oil spills; precipitation and atmospheric deposition; industrial processes (creosote, coal tar, asphalt, land-farming) (AMAI, 1986a;b; RDRC, 1987; Tecsult, 1989; Vandermeulen, 1989); municipal effluents; and disposal (burial) of wastes containing PAHs (Jackson et al., 1985; van Coillie et al., 1990). Polycyclic aromatic hydrocarbons can also reach groundwater and fresh and marine surface water by leaching through soil and by surface run-off (Wakeham et al., 1980; Slooff et al., 1989; Wan, 1991; LGL, 1993).

An estimate of the quantity of PAHs discharged to water and soil from creosote-treated wood products has been attempted based on the PAH content in creosote, the volume of treated wood in use, the retention rates of the compounds for different species of wood, and an estimated 20% release or loss of compounds during the time the treated wood was in service (40 years for pilings, 50 years for railway ties). According to these calculations, creosote-related PAH releases to soil and water could be up to 2000 t/yr (LGL, 1993).

Spills of petroleum hydrocarbons result in 76 t/yr of PAHs being released into the Canadian environment. About 88% of the total number of spills occur on land and 12% on water (LGL, 1993).

Table 2 Annual Atmospheric Emissions of Polycyclic Aromatic Hydrocarbons in Canada During 1990 (LGL, 1993)

Sources

PAH releases

tonnes

%

Anthropogenic Sources

Industrial Processes

Aluminum plants

925

21

Metallurgical (including ferro-alloy)

19.5

0.4

Coke production

12.8

0.3

Petroleum refineries

2.5

0.1

0.1

<0.1

Combustion Sources

Residential Heating Wood

474

11.0

Others

29

0.7

Open air fires/agricultural burning

358

8.3

Incineration Teepee burners

249

5.8

Municipal (with sludges)

1.3

<0.1

Industrial

1.1

<0.1

Transportation Diesel

155

3.6

Gasoline

45

1

Other

1.2

<0.1

Thermal Power Plants

11.3

0.3

Industrial Combustion Wood

5.7

<0.1

Other

10.2

0.2

Commercial and Institutional Heating

2.7

0.1

Cigarettes

0.2

<0.1

Natural Sources

Forest Fires

2010

47

Total

4314

100

The metallurgical sector (metals and coking plants) released about 3.9 t/yr of PAHs into water in 1990 (LGL, 1993). This total does not include run-off from a tar pond at a steel plant in Sydney, Nova Scotia, that released close to 0.8 t/yr of PAHs into the aquatic environment, based on estimated emissions in 1989 (Lane et al., 1990).

Atmospheric deposition has been estimated to be the main source of PAHs to soils and sediments (Slooff et al., 1989; Christensen and Zhang, 1993). Deposition rates of PAHs to terrestrial and aquatic surfaces in the north-eastern United States have been estimated at 0.8 to 3 ng/cm2 per year in non-urban areas and up to 35 ng/cm2 per year in sites closer to urban areas (Gschwend and Rites, 1981; Rites and Gschwend, 1982). The magnitude of the atmospheric deposition of PAHs in Canada resulting from the long-range transport from foreign sources cannot be assessed owing to a lack of information. A total of 484 t/yr of PAHs was estimated to enter the Great Lakes as a result of dry deposition of PAHs from sources in Canada and the United States (Eisenreich et al., 1981, cited in NRCC, 1983).

2.3 Exposure-related Information

2.3.1 Fate

With the exception of some of the lighter compounds that volatilize from water or soil, PAHs are relatively non-volatile and of low solubility in water. In the atmosphere, they are mostly found adsorbed to particulate matter that can be removed by wet or dry deposition onto water or soil. Polycyclic aromatic hydrocarbons released to soil will adsorb to particulate matter and will slowly be degraded by microbial activity or transported adsorbed to soil particles by surface runoff. In aquatic systems, PAHs generally adsorb to suspended matter or sediments where they persist. Thus, although most PAHs are emitted to the atmosphere, sediments are the major environmental sink for these compounds.

The atmosphere is the main medium of transport for PAHs (Suess, 1976; Bjørseth et al., 1979). Polycyclic aromatic hydrocarbons in the atmosphere are primarily associated with suspended particulates, although they are also present in the vapour phase (Pupp et al., 1974; Pierce and Katz, 1975; Miguel and Friedlander, 1978; van Vaeck and van Cauwenberghe, 1978; Neff, 1979; NRCC, 1983; Bjørseth and Ramdhal, 1985; Slooff et al., 1989). Atmospheric particulates and adsorbed PAHs can be transported over brig distances (Lunde and Bjørseth, 1977; Neff, 1979; NRCC, 1983) before they are removed from the atmosphere either through dry or wet deposition to soil or water, or through chemical transformation (Neff, 1979; NRCC, 1983; Van Noort and Wondergem, 1985; Ligocki, et al., 1985). In addition to direct deposition to soils, PAHs can be deposited onto or absorbed by plants, from which they can be washed by rain, oxidized, or deposited into soil as a result of plant decay (Eisler, 1987).

Ozone-induced oxidation and hydroxylation are the two most important mechanisms by which PAHs are transformed in the atmosphere; both of these reactions are activated by sunlight (Lyman et al., 1982; NRCC, 1983; Slooff et al., 1989). The photo-oxidation half-lives in air for different PAHs vary from 0.4 to 68.1 hours; photolysis half-lives vary from 0.37 to 25 hours, excluding the long half-life for naphthalene (1704 to 13 200 hours) (U.S. EPA, 1990; Slooff et al., 1989). These chemical transformations are affected by several factors, including the nature of the particles to which the atmospheric PAHs are adsorbed (Korfmacher et al., 1980; NRCC, 1983; Behymer and Rites, 1988) and the quantity of PAHs adsorbed to the particulate matter (Kamens et al., 1988; Slooff et al., 1989). Polycyclic aromatic hydrocarbons are more persistent when they are bound to particulates with a high organic carbon content and when present in large quantities on the particulates. Minor transformation pathways for PAHs include reactions with nitrogen oxides (NOx) and sulphur dioxide (S02). Such reactions, however, occur mostly at the points of emission where temperatures vary between 100 and 2000C (Atkinson et al., 1979).

As in the atmosphere, PAHs in the water column are generally associated with particulates (Harrison et al., 1975; Wakeham et al., 1980; Germain and Langlois, 1988). Volatilization, photolysis, hydrolysis, biodegradation, and adsorption to particulate matter followed by sedimentation are the main processes governing the fate of PAHs in water (NRCC, 1983; Eisler, 1987, Slooff et al., 1989). The rate of volatilization depends on weather conditions, movement of water, and the molecular weight of the compounds (NRCC, 1983; Slooff et al., 1989). Polycyclic aromatic hydrocarbons of low molecular weight may volatilize from water, as indicated by the volatilization half-lives of naphthalene (0.4 to 3.2 hours; Slooff et al., 1989; Southworth, 1979) and anthracene (17 hours; Southworth, 1979). A high molecular weight PAH such as pyrene, however, has a volatilization half-life ranging from 115 hours to 3.2 years (Southworth, 1979; Lyman et al., 1982). Many of the PAHs in oil spilled on water volatilize (NRCC, 1983). Polycyclic aromatic hydrocarbons can be biodegraded in water. Half-lives have been estimated to range from 0.5 to 20 days for naphthalene and from 0.6 to 5.2 years for pyrene under aerobic conditions (U.S. EPA, 1990). Photo-oxidation in water also occurs, with estimated half-lives of 8.6 days to 1.2 years for B[a]P (Smith et al., 1978) and 0.1 to 4.4 years for anthracene (Radding et al., 1976). For most PAHs in the water column, sedimentation constitutes the primary removal mechanism (NRCC, 1983).

Sediments are the final environmental sink for PAHs (Payne et al., 1988; Vandermeulen, 1989) where they persist and transform very slowly. Biodegradation half-lives of sediment-bound PAHs range from 0.3 to 129 days for naphthalene and from 0.3 to 58 years for B[a]P (Herbes and Schwall, 1978). Polycyclic aromatic hydrocarbons in sediments are relatively stationary. Near Seattle, Washington, 63% of particle-bound PAHs were found at the bottom of Puget Sound less than 100 m away from their point of entry (Murphy et al., 1988). Nonetheless, sediments may be partially resuspended and are then subject to transport processes (Windsor and Rites, 1979; Larsen et al., 1986).

Polycyclic aromatic hydrocarbons are removed from soils principally by volatilization and microbial activity, the extent of which varies, depending on several factors such as temperature, soil type, presence of other contaminants, and previous contamination (Beak, 1981; Bulman et al., 1985; PACE, 1988; ASTDR, 1990b; Cooper, 1991; Wild et al., 1991). Low molecular weight PAHs volatilize more rapidly than high molecular weight PAHs (Slooff et al., 1989; Wild and Jones, 1993). Biodegradation half-lives in soil have been estimated for various PAHs, including anthracene (170 days [Herbes and Schwall, 1978] to 8 years [Wild et al., 1991)]); phenanthrene (from 2.5 to 210 days [Sims and Overcash, 1983]) to 5.7 years [Wild et al., 1991]); and B[a]P (8.2 years [Wild et al., 1991] and 0.3 to 58 years [Herbes and Schwall, 1978]). In farming operations in which refinery wastes and sewage sludge are applied to the soil, low molecular weight PAHs (two and three rings) are expected to volatilize or biodegrade within three to four month; higher molecular weight PAHs (more than three rings) can be substantially biodegraded in a four-month period, but repeated applications of oily sludges containing PAHs may result in accumulation of these compounds in soil (PACE, 1988).

Contamination of groundwater by PAHs can occur as a result of leaching through soils, especially when PAHs are accompanied by mobile organic solvents or when channels are present in the soil (Bedient et al., 1984; Slooff et al., 1989). Naphthalene was the most mobile PAH reported below a creosote-contaminated site in the United States; concentrations of naphthalene at a depth of 3 m were 5% of those at a depth of 0.2 to 0.5 m (Wang et al., 1983). Contamination of groundwater has been observed following application of oily sludges to soil (PACE, 1988).

Polycyclic aromatic hydrocarbons can accumulate in a variety of organisms. Bioconcentration factors (BCFs) ranging from 4 to 7800 have been reported for various PAHs in unicellular algae (Lu et al., 1977; Dobrowsky and Epifanio, 1980; Casserly et al., 1983; Mailhot, 1987). Polycyclic aromatic hydrocarbons can be oxidized by these organisms, although only 10 to 37% of accumulated B[a]P was oxidized by the green alga, Selenastrum capricornutum (Warshawsky et al., 1983).

In aquatic invertebrates, bioconcentration from water has been measured in the cladoceran Daphnia pulex after one hour of exposure to five PAHs; the BCFs were correlated with the Kows for individual PAHs, and ranged from 131 for naphthalene to 10 109 for benz[a]anthracene (Southworth et al., 1978). For several invertebrates, bioconcentration of PAHs from sediments has also been observed, especially when PAH loadings in the sediments were high (Eadie et al., 1983; Pruell et al., 1986). For example, the BCF of anthracene increased from 1800 to 9096 in the marine amphipod Hyalella azteca when sediment was added to water containing 8 µg/L of anthracene (Landrum and Scavia, 1983). Biotransfer through food can also occur; 44% of the B[a]P content of the diatom Thalassiosira pseudomona was transferred to larvae of the clam Mercenaria mercenaria in 24 hours (Dobrowsky and Epifanio, 1980). The rate of elimination of PAHs in aquatic invertebrates is much lower than the rate of uptake (NRCC, 1983), resulting in a long half-life for PAHs in invertebrates. For example, in the mussel Mytilus edulis, depuration half-lives for B[a]P, benz[a]anthracene, and fluoranthene were 15, 18, and 30 days, respectively (Pruell et al., 1986).

Based on assays conducted in laboratories with radio-labelled PAHs, bioconcentration factors for fish range from 23 for 14C-naphthalene to 675 for 14C-anthracene (Spacie et al., 1983; Linder and Bergman, 1984; Solbakken et al., 1984); these values may be overestimated since radioactive counting detects metabolites as well as the original compounds. In addition to direct bioconcentration from water, uptake from food may also occur, as observed in the field in English sole (Parophrys vetulus) (Malins et al., 1984). However, the biomagnification of PAHs up aquatic food chains is not expected to occur since the elimination of PAHs and their metabolites is relatively rapid in fish (EAG, 1990). Following exposure for 5 hours to PAHs, the depuration half-lives in the bluegill sunfish (Lepomis macrochirus) were 17 hours for 14C-anthracene and 67 hours for 14C-B [a]P (Spacie et al., 1983). In alevin of the Atlantic cod (Gadus morhua), the half-lives were less than 24 hours for 14C-naphthalene and 14C-B[a]P (Solbakken et al., 1984).

Plants take in more PAHs from the atmosphere than from the soil; plants with large leaves, therefore, generally contain more PAHs than those with narrow leaves (Edwards, 1983). The outer surfaces of plants, such as fruit peel, contain more PAHs than the internal structures. After uptake, there is little transfer or translocation of PAHs within the plant (Edwards, 1986).

2.3.2 Concentrations

Polycyclic aromatic hydrocarbons are widely distributed in the environment and have been detected in numerous media to which humans and biota are exposed including air, water, food, soil, sediment, and tobacco smoke. Most available data pertain to concentrations near known sources of PAHs, rather than from uncontaminated sites.

Air. Concentrations of PAHs in ambient air have been measured in British Columbia, the Yukon Territory, Manitoba, Ontario, Quebec, and Nova Scotia. Concentrations for the 13 PAHs considered in this report are presented in Table 3. It is not always possible to separate the [b] and [k] isomers or to isolate the [j] isomer of benzofluoranthene with most current analytical methods. Tardif and Chiu (1992) successfully analyzed these compounds in a limited number of samples; the three isomers were present in consistent proportions at the seven sites sampled. The concentrations of the [b], [j] and [k] isomers of benzofluoranthene presented in Table 3 were calculated based on these proportions.

Data on atmospheric concentrations of PAHs are available for one rural site located at Walpole Island, Ontario. The median concentration of total PAHs was 7.2 ng/m3 (mean of 10.0 ng/m3), with a maximum of 40.4 ng/m3 (Dann, 1990). The highest concentrations for individual PAHs were recorded for phenanthrene and fluoranthene, while anthracene, B[a]P, and indeno[1,2,3-cd]pyrene were detected in fewer than half the samples.

The highest concentrations of PAHs in ambient air in Canada were measured at stations located about 1 km from aluminum smelters using the Horizontal Stud Söderberg process in Jonquière and Shawinigan, Quebec. The median concentrations of total PAHs (the sum of 26 compounds) were 693 ng/m3 (mean of 1687 ng/m3) at Jonquière and 435 ng/m3 (mean of 1519 ng/m3) at Shawinigan, with maxima of 10400 and 16 390 ng/m3, respectively. In Jonquière, the highest concentrations of individual PAHs were recorded for phenanthrene and fluoranthene. The maximum concentrations for B[a]P were 305 ng/m3 in Jonquière and 460 ng/m3 in Shawinigan (Ringuette et al., 1993).

Table 3 Concentration of Polycyclic Aromatic Hydrocarbons in Ambient Air (ng/m3) in Canada
Sector Aluminum plant* Wood heating Rural
Location Kitimat4 Jonquière5 Shawinigan6 Whitehorse5 Sept-Îles6 Walpole7
Number of samples 24 42 31 5 24 62
Sampling period 1990 1989 to 1991 1989 to 1991 1982 to 1983 1990 to 1991 1988 to 1989
Statistical Value Mean Median Mean Median Mean Median Mean Median Mean Median Mean Median

naphthalene

3.10

2.00

26.00**

3.26**

12.45

4.33

na

na

43.74

1.29

na

na

acenaphthene

15.10

5.40

121.72

11.99

55.00

9.85

Na

Na

4.52

0.90

0.32

0.23

fluorene

15.63

9.85

41.45

8.34

46.24

10.97

26.47

22.15

15.48

6.15

0.66

0.39

phenanthrene

57.81

45.50

371.26

197.78

389.65

93.87

271.43

170.77

50.35

22.11

4.18

2.93

anthracene

4.07

2.35

42.92

9.47

29.90

8.47

22.81

14.81

8.65

4.02

<0.05

<0.05

pyrene

14.36

10.00

195.89

90.42

206.78

62.07

49.41

43.78

12.28

5.19

0.73

0.51

fluoranthene

22.14

16.00

272.27

132.13

261.00

74.41

50.89

45.68

15.54

6.69

1.32

0.95

benz[a]anthracene

2.52

1.65

57.71

8.69

35.14

4.89

9.81

6.97

3.09

1.14

0.07

0.04

benzo[a]pyrene

2.06

1.15

36.37

7.15

28.47

4.21

7.66

5.39

1.93

0.68

0.08

<0.05

benzo[b,j,k]fluoranthene1

7.90

6.05

166.60

62.91

149.06

29.24

12.75

9.55

5.65

2.38

0.31

0.16

benzo[b]fluoranthene 2

4.42

3.39

93.3

35.23

83.47

16.37

7.14

5.35

3.16

1.33

0.17

0.09

benzo[j]fluoranthene2

1.96

1.51

41.65

15.73

37.26

7.31

3.19

2.39

1.41

0.60

0.08

0.04

benzo[k]fluoranthene 2

1.50

1.15

31.65

11.95

28.32

5.56

2.42

1.81

1.07

0.45

0.06

0.03

indeno[1,2,3-cd]pyrene

1.88

0.92

30.57

8.38

24.81

5.66

5.86

3.60

1.62

0.49

0.10

<0.05

Total PAHs3

155.50

110.03

1687.04

692.88

1519.13

434.79

506.50

498.11

202.09

60.52

9.97

7.19

Table 3 Concentration of Polycyclic Aromatic Hydrocarbons in Ambient Air (ng/m3) in Canada (continued)
Sector Urban: Road Transportation and combustion
Location Vancouver8 Winnipeg9 Windsor7 Toronto10 Montreal10 Sydney11
Number of samples 11 10 55 42 23 39
Sampling period 1985 1989 1988-1989 1984-1986 1989-1990 1981-1982
Statistical Value Mean Median Mean Median Mean Median Mean Median Mean Median Mean Median

naphthalene

na

na

na

na

na

na

na

na

na

na

na

na

acenaphthene

na

na

0.06

0.00

5.21

2.40

2.24

1.66

1.93

1.30

na

na

fluorene

na

na

0.94

0.96

4.65

2.80

5.51

4.50

3.12

2.30

na

na

phenanthrene

na

na

5.26

4.62

34.79

22.67

15.64

14.47

19.59

18.98

2.20

0.34

anthracene

na

na

0.45

0.28

3.77

0.63

1.60

0.94

1.89

1.28

Na

Na

pyrene

4.92

5.30

1.44

1.13

6.91

4.56

3.87

3.36

6.65

4.10

3.82

0.49

fuoranthene

6.76

8.50

1.55

1.13

11.08

7.03

4.91

4.82

9.60

5.45

3.52

0.36

benz[a]anthracene

0.81

0.70

0.19

0.12

0.74

0.23

0.40

0.17

1.16

0.28

4.40

0.32

benzo[a]pyrene

0.35

0.30

0.09

<0.05

0.82

0.37

0.30

0.16

0.56

0.29

1.74

0.29

benzo[b,j,k]fluoranthene1

1.32

1.20

0.37

0.16

2.71

1.22

1.26

0.83

5.48

2.11

3.64

0.74

benzo[b]fluoranthene 2

0.74

0.67

0.21

0.09

1.52

0.68

0.71

0.46

2.51

1.18

2.04

0.41

benzo[j]fluoranthene2

0.33

0.30

0.09

0.04

0.68

0.31

0.31

0.21

1.12

0.53

0.91

0.19

benzo[k]fluoranthene2

0.25

0.23

0.07

0.03

0.51

0.23

0.24

0.16

1.85

0.40

0.69

0.14

indeno[1,2,3-cd]pyrene

1.89

1.14

0.19

0.12

0.87

0.53

0.46

0.29

0.96

0.51

1.95

0.31

Total PAHs3

20.41

19.10

73.58

10.47

89.28

62.16

47.58

44.09

69.01

57.50

31.16

3.95

na - not analyzed

1 Total for [b], [j], and [k] isomers of benzofluoranthene

2 Values for [b], [j], and [k] isomers of benzofluoranthene calculated using proportions for isomers reported by Tardif and Chiu (1992)

3 Total of all PAHs analyzed at each location

4 Johnson, 1991

5 Ringuette et al., 1993

6 LEI, 1992a; b

7 Dann, 1992b; 1991

8 Faulkner, 1985

9 Bezak, 1990

10 Dann, 1989

11 Atwell et al., 1984

* Kitimat: Vertical Stud Söderberg
Shawinigan and Jonquière: Horizontal Stud Söderberg

High concentrations were also measured in urban areas where heating by wood combustion is prevalent. In Whitehorse, Yukon, the median concentration of total PAHs was 498 ng/m3 in winter (mean of 507 ng/m3), with a maximum of 1000 ng/m3. Phenanthrene, fluoranthene, and fluorene were present at the greatest concentrations (Ringuette et al., 1993). In Sept-Îles, Quebec, the concentrations of PAHs in winter were ten times higher than those measured in summer (Germain and Bisson, 1992).

High concentrations of PAHs have been reported in the vicinity of transportation sources. In downtown Montreal, the median concentration of PAHs was 57 ng/m3 (mean of 69 ng/m3) and the maximum measured was 289 ng/m3 (Dann, 1991).

Tobacco Smoke. The levels of PAHs in tobacco smoke have been identified in mainstream and sidestream smoke from cigarettes, cigars, and pipes. The concentrations of PAHs in the mainstream smoke, expressed as µg/100 cigarettes, ranged from 0.4 to 2.2 for B[b]F, 0.6 to 2.1 for B[j]F, 0.6 to 1.2 for B[k]F, 0.5 to 7.8 for B[a]P, and 0.4 to 2.0 for IND (IARC, 1983).

Surface Water. Concentrations of PAHs in fresh water at selected sites in Canada are presented in Table 4. Concentrations may include both dissolved PAHs and extractable PAHs adsorbed to suspended particles in the water column. Concentrations of PAHs in the Atlantic provinces were low, with a median concentration of B[a]P below the detection limit (0.8 to 1 ng/L) and a maximum concentration of 9 ng/L (Wong and Bailey, 1990). The highest PAH concentration in the St. Lawrence River was for phenanthrene, for which the median concentration was 17.6 ng/L and the maximum value was 119 ng/L (Envirodat, 1993). Concentrations of PAHs in the Niagara River ranged from 2.1 to 63.9 ng/L, with a median of 13.4 ng/L (Kuntz, 1990), while they ranged from below the detection limit (0.4 ng/L) to 6 ng/L in the Detroit River near Windsor, Ontario (Kaiser et al., 1985). In the Mackenzie River (Northwest Territories), water samples collected under the ice cover in March 1988 had lower concentrations of total PAHs (from 10.7 to 546 ng/L) than those recorded in June 1986 (54 to 1824 ng/L) at the same sites, indicating that run-off following ice breakup carried PAHs into the river (Nagy et al., 1987; 1989). The highest concentrations of PAHs in water in Canada were reported for water samples from ditches beside utility and railway lines near Vancouver, B.C. (Wan, 1991; 1993). The highest mean concentrations were measured near utility poles treated with creosote, with values of 488 µg/L for naphthalene, 1642 µg/L for phenanthrene (Wan, 1991), 2035 µg/L for fluoranthene, and 5356 µg/L for total PAHs (Wan 1993).

In the Atlantic provinces, concentrations of B[a]P in marine waters ranged from below the detection limit (0.8 to 1 ng/L) to 16 ng/L, while concentrations of fluoranthene ranged up to 113 ng/L. The detection frequency was low for B[a]P (2%), but higher for fluorene (25%) (Wong and Bailey, 1990). The highest concentration of total PAHs measured in harbours in Nova Scotia was 880 ng/L (O'Neill and Kieley, 1992).

Table 4 Concentrations of Polycyclic Aromatic Hydrocarbons in Fresh Water (µg/L) in Canada

Location

Railway Ditches Poles1

Railway Ditches with Poles1

Utility Right-of-way Ditches1

Farm Ditches1

Niagara River2

Quebec3

Nova Scotia3

New Brunswick3

Newfoundland3

Number of samples

3

2

4

2

51

 

 

 

 

Sampling period

1990 to 1991

1990 to 1991

1990 to 1991

1991

1988 to 1989

1987 to 1993

1987 to 1993

1987 to1993

1987 to1993

Statistical value

mean

median

mean

median

median

median

median

median

median

naphthalene

0.19

8.5

0.17

0.35

0.0036

*

*

*

*

acenaphthene

0.57

206

0.63

1.04

*

0.01
(n=44)

*

*

*

fluorene

0.22

116

1.72

0.3

0.0007

0.015 (n=44)

0.007
(n=45)

*

*

phenanthrene

0.44

1027

7.67

0.4

0.0021

0.0176 (n=53)

*

*

*

anthracene

0.13

81

1.65

0.16

0.0002

*

*

*

*

pyrene

0.19

1233

3.41

0.19

0.0003

0.015 (n=57)

0.005
(n=52)

*

*

fluoranthene

0.26

2035

3.08

0.3

0.0008

*

*

*

*

benz[a]anthracene

0.12

195

0.69

0.1

0.0003

*

*

*

*

benzo[a]pyrene

0.10

43

0.48

0.1

0.0002

0.00046 (n=57)

0.001
(n=168)

0.001
(n=78)

0.001
(n=79)

benzo[b+k]fluoranthene4

0.10

144

1.1

0.19

*

*

*

*

*

indeno[1,2,3-cd]pyrene

0.15

17.6

0.4

0.15

*

*

*

*

*

Total PAHs5

3.97

5356.3

23.2

5.56

0.0036

*

*

*

*

  1. Wan, 1993
  2. Kuntz, 1990
  3. Envirodat, 1993
  4. Total for [b] and [k] isomers of benzofluoranthene reported by the authors; data not available for [j] isomer
  5. Total of all PAHs analyzed at each location

* less than limit of detection or not reported

Groundwater. Few studies are available concerning the presence of PAHs in groundwater in Canada. Near a former refinery at Pincher Creek, Alberta, the concentrations of pyrene ranged from below the detection limit to 300 µg/L, with a median concentration of about 30 µg/L (ETL, 1984). Concentrations of fluorene at this site ranged from below the detection limit to 230 µg/L; the median was 40 µg/L (ETL, 1984). At Newcastle, New Brunswick, naphthalene was detected at concentrations as high as 2.8 µg/L and B[a]P as high as 0.32 µg/L in groundwater near a wood preserving plant (WMS, 1989).

Drinking Water. Few data are available on the levels of PAHs in drinking water, Particularly for those compounds being considered principally with respect to human health in this assessment. B[a]P was not detected in a survey of seven water treatment plants in the area of Niagara Falls (detection limit 1.0 µg/L) (MOE, 1984). The only PAHs detected in 2006 analyses of treated drinking water conducted in the Ontario Drinking Water Surveillance Program in 1987 were B[k]F (twice at 1 ng/L), fluoranthene (20 and 30 ng/L), and pyrene (twice at 40 ng/L). The range of compounds examined was not specified (MOE, 1989). In Quebec, the mean concentrations of PAHs in treated water determined recently ranged from <5 to 623 ng/L for fluoranthene, <5 to 40 ng/L for B[k]F, to 40 ng/L for B[b]F, and were less than 5 ng/L for B[a]P, IND, and benzo[ghi]perylene (Ayotte and Larue, 1990).

In a survey conducted in the Atlantic Region from 1985 to 1988 (Environment Canada, 1989a;b;c;d), the concentrations of fluoranthene, B[a]P, B[b]F, B[k]F,IND, and benzo[ghi]perylene (B[ghi]P) were determined (detection limits of 0.001,0.001,0.001, 0.001, 0.005, 0.005 µg/L, respectively). In Newfoundland, only fluoranthene was detected at all sites. Concentrations were near the detection limit, except for the water supply at Baie Verte for which a maximum value of 0.054 µg/L was reported. At this particular site, B[b]F (0.001 to 0.005 µg/L), B[k]F (0.001 to 0.003 µg/L), and B[a]P (0.001 to 0.003 µg/L) were also detected. In Prince Edward Island, fluoranthene was detected at low levels (range, 0.001 to 0.012 µg/L) in every supply. In one sample collected in 1986 from a well at St. Eleanors, P.E.I., B[b]F (0.001 to 0.003 µg/L), B[a]P (0.001 to 0.003 µg/L), and B[ghi]P (0.005 to 0.021 µg/L) were also detected. In Nova Scotia, only fluoranthene was detected in raw sources, at levels ranging from the detection limit to 0.008 µg/L. In New Brunswick, several samples contained fluoranthene at levels ranging from 0.001 to 0.005 µg/L. Concentrations of B[b]F, B[k]F, and B[a)P ranged up to 0.002, 0.001, and 0.003 µg/L, respectively, in samples collected at a well in Fredericton.

Sediments. Data on concentrations in sediments have been identified for all provinces, except Manitoba and Saskatchewan. Representative concentrations are summarized in Table 5. The highest concentrations were reported in sediments collected in or around industrialized harbours close to known sources of PAHs. The concentrations presented in this section are expressed on a dry weight (d.w.) basis, unless otherwise noted.

Table 5 Concentrations of Polycyclic Aromatic Hydrocarbons in Sediments (µg/g dry weight) in Canada

Location

Vancouver
Harbour
B.C. 1

Kitimat
Arm
B.C. 2

Hamilton
Harbour,
Ont. 3

Kettle
Creek,
Ont. 4

Montreal
Harbour
Que. 5

Muggah
Creek,
Sydney
Harbour6

Sydney
Harbour,
N.S.7

Luxton and Mountain Lakes, N.S.8

Lake ELA375,
Ont. 9

Number of samples

23

6

25

17

42

1

38

8

9

Sampling period

1987

1990

1988 to 1989

1987

1989

1980

1986

1980s

1990

Statistical value

median

median

median

median

median

Single result

median

median

median

naphthalene

0.40

0.14

19.8

*

0.2

*

0.05

0.24

0.014

acenaphthene

0.05

0.06

3.8

*

0.2

*

<0.05

<0.01

0.001

fluorene

0.10

0.11

7.0

*

0.4

*

<0.02

0.04

0.007

phenanthrene

0.51

0.34

25.8

*

.28

655

0.54

0.07

0.059

anthracene

0.12

0.26

10.9

*

0.6

*

<0.05

<0.01

0.004

pyrene

0.74

2.20

22.8

*

2.8

413

0.61

<0.01

0.004

fluoranthene

0.64

2.90

34.2

*

2.8

607

0.715

<0.01

0.063

benz[a]anthracene

0.24

2.45

8.8

*

1.1

414

0.37

<0.01

0.017

benzo[a]pyrene

0.34

4.35

8.9

1.0

1.2

109

0.515

<0.01

0.025

benzo[b+k]fluoranthene1

0.41

6.4

15.3

*

2.3

184.6

0.57

<0.01

0.098

indeno[1,2,3-cd]pyrene

0.16

3.3

15.7

*

1.0

71.5

0.29

<0.01

0.087

Total PAHs11

5.36

31.24

282.5

28

14.2

2830

5.20

0.51

0.56

  1. Boyd et al., 1989
  2. Goyette, 1991
  3. Murphy, et al., 1993
  4. AEL, 1988
  5. Environnement Illimitée Inc., 1990
  6. Sirota et al., 1983
  7. Kieley et al., 1988
  8. Keizer, 1990
  9. Lockhard, 1990
  10. Total for [b] and [k] isomers of benzofluoranthene; data not available for [j] isomer
  11. Total of all PAHs analyzed at each location

* less than detection limit or not reported

Sydney Harbour, Nova Scotia, is a major site of PAH contamination in Canada. The highest levels of PAHs were measured in the South Arm near the mouth of Muggah Creek, where a steel production complex is located. The lowest concentrations were measured in the Northern Arm (Sirota et al., 1983). At the mouth of Muggah Creek, the concentrations of phenanthrene, pyrene, and B[a]P were 655, 413, and 109 µg/g, respectively, with a total PAH concentration (12 compounds) of 2830 µg/g in 1981. A decrease in the steel plant production followed by the closure of the coking facility in 1988 led to a drop in PAH concentrations in sediments. In 1986, close to the same locations, the maximum concentration of total PAHs (18 compounds) was reported to be 310 µg/g near the effluent outfall of the steel complex; the 13 PAHs considered here accounted for 251 µg/g (Kieley et al., 1988). The concentration of total PAHs was 0.0029 µg/g at the least contaminated station of Sydney Harbour, located 6.5 km from Muggah Creek. The median concentration of total PAHs for all the stations was 5.2 µg/g. Other reported median and maximum values were 0.72 and 60.0 µg/g for fluoranthene; 0.61 and 33.0 µg/g for pyrene; and 0.515 and 28 µg/g for B[a]P.

In comparison, median concentrations in sediments at sites remote from pollution sources in Luxton and Mountain lakes in Kejimkujik National Park, Nova Scotia, were 0.51 µg/g for total PAHs and less than 0.01 µg/g for B[a]P; maximum concentrations were 0.86 and 0.05 µg/g, respectively. For phenanthrene, the median was 0.07 µg/g and the maximum, 0.10 µg/g. Pyrene was not detected in more than half of the samples; the maximum value was 0.05 µg/g (Keizer, 1990). Lockhart (1990) observed similar results in two experimental lakes in Northern Ontario (median and maximum total PAH concentrations: 0.56 and 0.85 µg/g; median and maximum B[a]P concentrations: 0.03 and 0.04 µg/g).

In the sediments of Montreal Harbour, Quebec, the median and maximum concentrations were 5.7 and 66.0 µg/g for phenanthrene, and 2.8 and 39.0 µg/g for pyrene. The median concentration of B[a]P was 1.2 µg/g, with a maximum of 29.0 µg/g. The median concentration of total PAHs was 14.2 µg/g and the maximum value was 278 µg/g (Environnement Illimitée Inc., 1990).

In Hamilton Harbour, Ontario, where the two largest steel mills in Canada are located, the reported median for total PAHs was 285 µg/g, with a range from 1.6 to 1470 µg/g. The median concentrations for pyrene and fluoranthene were 17.3 and 20.1 µg/g, respectively, and their maxima were 280 µg/g (pyrene) and 189 µg/g (fluoranthene). Median and maximum concentrations of B[a]P were 9.5 µg/g and 69.2 µg/g (Murphy et al., 1993).

The median concentration of total PAHs in the sediments of Kettle Creek, downstream from an old oil-gasification complex at Port Stanley, Ontario, was 28 µg/g; the maximum value was 499 µg/g. For B[a]P, median and maximum concentrations were 1.0 and 40.2 µg/g, respectively (Canviro, 1988).

In Vancouver Harbour, British Columbia, the median concentration for total PAHs was 5.4 µg/g and the maximum was 36.8 µg/g (Boyd et al., 1989). Median and maximum concentrations were 0.74 µg/g and 5.55 µg/g for pyrene; 0.64 and 5.63 µg/g for fluoranthene; and 2.42 µg/g and 0.134 µg/g for B[a]P. In Kitimat, B.C., the median for total PAHs was 31.2 µg/g, ranging from 0.14 to 258 µg/g (Goyette, 1991). Median and maximum concentrations were 2.9 and 27.0 µg/g for fluoranthene, 2.2 and 23.1 µg/g for pyrene, and 4.35 and 39 µg/g for B[a]P.

Soils. The highest concentrations of PAHs in Soils are found near former coking plants (CCME, 1989). The Resources Development Research Centre identified 144 potential sites 9f former coking plants in Canada, 80 of which are confirmed sites (RDRC, 1987). Although the majority (70) of these sites are located in Ontario, the RDRC identified sites in all provinces except Prince Edward Island. The site of a former coking facility in Lasalle, Quebec, was investigated after the facility closed in 1976. Concentrations of B[a]P, the only compound investigated in 1985, varied from the detection limit (unspecified) to 1300 µg/g (d.w.). The site has now been restored and monitoring studies entailing more than 800 soil samples have shown that the B [a]P concentration is less than 10 µg/g (d.w.) (ARGUS Groupe Conseil Inc., 1991). At the former coking facility in Sorel, Quebec, the maximum concentration of total PAHs (16 compounds) was reported to be 11 473 µg/g before site restoration, with a median concentration of 18 µg/g (Tecsult, 1989).

Kieley et al. (1986) reported median and maximum concentrations of 965 and 16 000 µg/g (d.w.) for total PAH (12 PAH compounds) in a soil sample collected at the site of a wood preserving plant in Newcastle, New Brunswick. In Ottawa, Ontario, decontamination measures were implemented at a site that was contaminating the Rideau River. The B[a]P concentration in soil was reported to be 2.4 µg/g before the restoration (RDRC, 1987). The site of a former refinery in Pincher Creek, Alberta, was investigated after the refinery was dismantled. Before site restoration, high PAH concentrations (primarily alkylated PAHs) were observed in the soil. Maximum concentrations for fluoranthene and pyrene were both 260 µg/g and their median concentrations were 0.75 and 0.50 µg/g, respectively. B[a]P was not detected, however. In the vicinity of the plant, total concentrations of PAHs in soil were reported to be 9810 µg/g (ETL, 1984).

Biota. Concentrations of PAHs in biota in Canada are presented in Table 6.

In marine organisms on the Atlantic coast, 12 PAHs were detected in lobster (Homarus americanus) from Sydney Harbour, Nova Scotia, in 1981. A mean concentration of B[a]P of 0.031 µg/g (d.w.) was recorded in the tails of the lobster and 0.84 µg/g (w.w.) in the hepatopancreas. Total concentrations of PAHs were 2.29 µg/g (d.w.) and 72.7 µg/g (w.w.), respectively (Sirota et al., 1983). Other studies have revealed high PAH concentrations in other organisms in the area (Matheson et al., 1983; Uthe and Musial, 1986; Kieley et al., 1988). For example, mussels (Mytilus edulis and Modiolus modiolus) collected in Sydney Harbour contained up to 4.2 µg/g (w.w.) of PAHs (O'Neill and Kieley, 1992).

Table 6 Concentrations of Polycyclic Aromatic Hydrocarbons in Biota (µg/L) in Canada

Location

Vancouver
Harbour B.C. 1

Fraser River, B.C. 2

St. Lawrence River 3,4

Sydney Harbour
(South Arm) N.S. 5

Sydney Harbour
N.S. 6

Species

Dungeness crab (Cancer magister) livver (d.w.)

Prickly sculpin (Cottus asper) liver (d.w.)*

English sole (Parophris vetulus) lier (w.w.)*

Mussels (Mytilus edulis) (w.w.)

Northern pike (Esox lucius) (w.w.)

Lobster (Homarus americanus) tail muscle (d.w.)

Lobster (Homarus americanus) hepatopancreas (w.w.)

Mussels (Modiolus modiolus, Mytilus edulis) (w.w.)

Number of samples

3

4

4

25

7

10

10

9

Sampling period

1986

9186

1981

1989

1989

1980

1980

1986

Statistical value

median

median

median

median

median

median

median

median

naphthalene

**

**

0.040

0.026

0.210

**

**

0.061

acenaphthene

**

**

<0.160

0.002

**

**

**

<0.007

fluorene

<0.06

<0.06

<0.10

0.009

0.030

**

**

<0.006

phenanthrene

0.037

0.016

0.020

0.024

0.040

0.43

3.18

0.064

anthracene

<0.001

<0.001

**

0.002

0.001

**

**

<0.05

pyrene

<0.010

<0.010

<0.040

0.009

0.004

0.20

4.82

0.042

fluoranthene

0.013

0.013

**

0.026

0.020

0.49

11.55

0.025

benz(a)anthracene

<0.001

<0.001

**

0.007

**

0.478

18.4

0.035

benzo(a)pyrene

<0.005

<0.005

**

0.002

**

0.03

0.84

0.004

benzo(b+k)fluoranthene7

0.008

<0.005

**

**

**

0.84

2.49

0.015

indeno(1,2,3-cd)pyrene

<0.060

<0.060

**

**

**

0.04

0.82

0.027

Total PAHs8

**

**

0.078

0.153

0.405

2.29

72.7

0.680

  1. Goyette and Boyd, 1989
  2. Chapman et al., 1981
  3. Langlois, 1989
  4. Lapierre, 1989
  5. Sirota et al., 1983
  6. Kieley et al., 1988
  7. Total for [b] and [k] isomers of benzofluoranthene; data not available for [j] isomer
  8. Total of all PAHs analyzed at each location

* d.w.- dry weight; w.w.- wet weight

** less than limit of detection (variable for each compound or study) or not reported

On the Pacific coast, studies have been conducted in Vancouver Harbour, British Columbia. Total concentrations of PAHs in mussels collected near creosoted barricades and wharfs were between 54 and 215 µg/g (w.w.) (Goyette and Boyd, 1989). In 1986, appreciable concentrations of phenanthrene [0.830 µg/g (d.w.)] and pyrene [0.36 µg/g (d.w.)] were measured in the livers of Dungeness crab (Cancer magister). These compounds were also detected in the livers of English sole (Parophris vetulus), but in lower concentrations [0.037 µg/g for phenanthrene and 0.021 µg/g for pyrene, (d.w.)] (Goyette and Boyd, 1989).

In freshwater organisms, PAHs have been detected in Quebec in numerous species of fish from Lakes St. Francis, St. Louis, and St. Pierre and from the St. Lawrence Estuary. The median concentrations for total PAHs in the livers of northern pike (Esox lucius) and yellow walleye (Stizostedion vitreum) were 0.40 and 0.45 µg/g (w.w.), respectively (Lapierre, 1989).

In the Fraser River in British Columbia, naphthalene and phenanthrene were detected in the prickly sculpin (Cottus asper), an abundant freshwater fish in the region (Chapman et al., 1981). In 1988, PAHs were measured in the livers of five species of fish (Swain and Walton, 1989). Naphthalene was the most frequently detected compound, with the highest concentrations in the liver of the northern squawfish (Ptychocheilus oregonensis).

For birds, the only Canadian data identified are from the late 1970s, for herring gulls (Larus argentatus) from Pigeon Island and Kingston, Ontario, where levels in liver were measured for 30 PAHs. For the compounds of concern in this report, the mean levels (in µg/kg lipid) were: 0.05 for naphthalene, 0.038 for acenaphthene, 0.152 for anthracene, 0.082 for fluoranthene, 0.038 for benzo[a]pyrene, 0.044 for fluorene, and 0.076 for pyrene (Rallett et al., 1977). These levels were slightly lower than those found for fish from Lake Ontario from the same time period (Rallett et al., 1984).

Food. Concentrations of PAHs in uncooked food depend principally on its source. For example, vegetables, fruits, and fish obtained from polluted areas generally contain higher concentrations of PAHs than those from less polluted regions.

In cooked food, the method of cooking is generally the primary determinant of the PAH content and this content varies considerably depending on cooking habits (Santodonato et al., 1981). In a report prepared under contract for Agriculture Canada, the results were reported of an analysis of 8 PAHs in 208 Canadian food composite samples (supplied by Health Protection Branch, Health Canada, for the Total Diet Program) (Das, 1987). Samples for analysis were prepared for consumption just as in the average household kitchen. Raw meats were cooked (methods unspecified); fresh vegetables were cooked (no salt added) or if not cooked, then properly peeled, trimmed, or otherwise cleaned for serving; processed foods were prepared following directions on the label. The concentrations of three of the five PAHs considered principally in the assessment of potential effects on human health ranged from not detected (detection limit unspecified) in all food types examined to 1.78 µg/kg in vegetables/fruits for B[b]F, not detected to 0.30 µg/kg in meat/poultry/fish for B [k]F, and not detected to 1.13 µg/kg in meat/poultry/fish for B[a]P. However, the percent recoveries were particularly low for B[a]P and B[k)F in dairy and cereal products (69.3 and 24.6%, respectively) and in the cooking fats/salad oil/margarine/butter group (34.1 and 43.8%, respectively).

2.4 Effects-related Information

2.4.1 Toxicokinetics

Polycyclic aromatic hydrocarbons are metabolized to a wide variety of compounds principally by enzymes of the cytochrome P-450 mixed-function oxidase (MFO) system and epoxide hydrolase (IARC, 1983; Wislocki and Lu, 1988). The most important step of primary metabolism is epoxidation. A small proportion of all PAHs will, after epoxidation and subsequent diol formation, be further epoxidized by the microsomal monooxygenases to diol epoxides. Benzo-ring diol epoxides, in which the epoxide forms part of the bay region1 of the hydrocarbon molecule, are likely highly reactive and are implicated as the ultimate biologically reactive intermediates involved in the binding of PAHs to macromolecules and resulting toxicity (Santodonato et al., 1981; Varanasi et al., 1989). They may also be detoxified and excreted as metabolites in bile, feces, and urine by conjugation with glutathione or glucuronic acid, or by further metabolism to tetrahydrotetrols (Gelboin, 1980). The cytochrome P-450 activities of the lung are less than those in the liver and intestinal tract (i.e., metabolism to the active species is slow); moreover, the activity of the conjugating enzyme systems and epoxide hydrase in the lung is very low (i.e., detoxification is slow). These differences may account for possible variations in potency of PAHs following inhalation and ingestion (Santodonato et al., 1981).

The bay region diol epoxide intermediates of PAHs are currently considered to be the ultimate carcinogen for most PAHs, although for some, other reactive intermediates may also be important (ATSDR, 1990a;b; IARC, 1983; Santodonato et al., 1981). Though examined principally in animal models in vitro and in vivo, results of studies on the metabolism of benzo[a]pyrene in primary cultures of human hepatocytes and a human hepatoma cell line indicate that these mechanisms are likely relevant to humans (ATSDR, 1990a).

2.4.2 Experimental Animals and in vitro

Non-neoplastic Effects. Carcinogenicity has been the principal focus of research on the biological effects of PAHs; identified limited information indicates that non-neoplastic effects of these compounds in experimental animals are restricted principally to proliferating tissues such as the bone marrow, lymphoid organs, gonads, and intestinal epithelium (Santodonato et al., 1981; ATSDR, 1990a;b). Observed effects include, for example, prominent hemosiderosis, reduction in spleen size with marked cellular depletion, alteration of the enzyme activity in the intestinal mucosa leading to increased production of reactive intermediates and tissue injury, aplastic anemia, myelosuppression, and decreased fertility. The cytotoxicity is probably a consequence of an interaction with DNA in the S-phase of cell proliferation. The doses required to produce such responses in various species of experimental animals, however, are in some cases an order of magnitude higher than those which induce neoplasms (ATS DR, 1990a;b).

Carcinogenicity. In numerous studies in laboratory animals, various PAHs have induced tumours following principally dermal exposure [see, for example, ATSDR (1990a;b), IARC (1983), and Santodonato et al. (1981)]. Data are quite limited, however, on induction of tumours following exposure by routes by which the general population are principally exposed (i.e., inhalation and ingestion).

The carcinogenic effects of exposure to PAHs by inhalation have been examined in only a few limited identified studies, all of which were restricted to B[a]P (Heinrich et al., 1986; Laskin et al., 1970; Thyssen et al., 1981); moreover, in two of the investigations, animals were concomitantly exposed to other compounds (Heinrich et al., 1986; Laskin et al., 1970). In the study by Heinrich et al. (1986), the incidence of lung tumours was increased in female Wistar rats exposed to combustion gases of a coal furnace for an average of 16 hours/day, 5 days/week over a maximum of 22 months. The incidence of respiratory tract tumours was also increased in rats that inhaled 10 ppm (103 mg/m3) B[a]P and the atmospheric irritant, sulphur dioxide (S02) (Laskin et al., 1970).

In a study by Thyssen et al. (1981), groups of 24 male Syrian golden hamsters were exposed by inhalation (nose only) to 0, 2.2, 9.5, and 45.6 mg/m3 B[a]P for 4.5 hours/day, 7 days a week for the first 10 weeks, and for 3 hours/day for the rest of the exposure period (up to 96 weeks). Though there was a decrease in body weight gain in exposed animals during the first 10 weeks of the study, from the tenth to the sixtieth week, the body weights of all surviving exposed animals were similar to those of the controls (with the exception of the high exposure group). Mean survival was also decreased in the group exposed to 46.5 mg/m3. The incidences of unspecified tumours of the respiratory tract (nasal cavity, larynx, and trachea) were 0/27 for controls, 0/27 for the low-dose group, 9/26 (34.6%) for the mid-dose group, and 13/25 (52%) for the high-dose group. Exposure-related neoplasms (unspecified) were present in the pharynx (0, 0, 23, and 56% for control, low-, mid-, and high-dose, respectively), oesophagus (0, 0, 0, and 8% for control, low-, mid-, and high-dose, respectively), and forestomach (0, 0, 4, and 4% for control, low-, mid-, and high-dose, respectively). Lung tumours were not observed.

A relationship between the ingestion of benzo[a]pyrene and the development of benign and malignant tumours has been documented in several limited studies in experimental animals (ATSDR, 1990a). In the most extensive, though limited, early study in which mice were fed a diet containing benzo[a]pyrene at concentrations of 40 to 45 ppm {equivalent to 5.2 to 5.9 mg/[kg (b.w.)·d]} for 110 days, the incidence of stomach tumours was increased 10% or less, whereas in mice fed a diet containing 50 to 250 ppm B[a]P {equivalent to 6.5 to 32.5 mg/[kg (b.w.)·d]} for 70 to 197 days, it exceeded 70%. In a second experiment in which a diet containing 250 ppm B[a]P {equivalent to 32.5 mg/[kg (b.w.)·d]} was fed to mice for different periods of time, incidences of tumours of the fore stomach (the only tissue examined) were as follows: 2 to 4 days of feeding, 10%; 5 to 7 days of feeding, 30 to 40%, 30 days of feeding, 100%. However, increases in the incidence of forestomach tumours were not observed following administration of a lower concentration of benzo[a]pyrene in the diet { 100 ppm or equivalent to 13 mg/[kg (b.w.)·d) } for up to 7 days (Neal and Rigdon, 1967).

The carcinogenicity of a wider range of PAHs has been examined in experimental animals following direct. introduction into the lungs (see, for example, ATSDR, 1990b; IARC, 1983; Santodonato et al., 1981). In the most extensive of these investigations in which the broadest range of PAHs was examined, various amounts of anthanthrene (ANT), B[a]P, benzo[e]pyrene (B[e]P), B[b]F, B[j]F, B[k]F, benzo[ghi]perylene (B[ghi] P), or IND dissolved in residue-free acetone and a mixture of 1:1 beeswax and trioctanoin were implanted into the left lung following administration of anaesthetic and thoracotomy of groups of 38 three-month-old, inbred Osborne-Mendel female rats (Deutsch-Wenzel et al., 1983). Operative and post-operative mortality was less than 5%. After surgery, rats were observed until their natural deaths, which occurred up to 32 months following implantation. At the site of implantation, a granulomatous inflammatory response was observed. In some animals, keratinized epidermoid carcinomas invading the extrapulmonary chest wall were observed; other tumours of this type metastasized predominantly into local and distant lymph nodes, heart, uterus, ovaries, adrenal glands, and kidneys. In a small number of animals, there were pleomorphic sarcomas. On the basis of histological and statistical analysis, there was evidence of an exposure-response relationship for increases in tumour incidence for B[a]P, B[b]F, IND, B[k]F, B[j]F, and ANT. The incidence of epidermoid carcinomas and pleomorphic sarcomas of the lung at the highest dose for all compounds administered were: control, 0/0; B[b]F at 1.0 mg, 13/35 (37.1%); B[e]P at 5.0 mg, 1/35 (2.9%); B[j]F at 5.0 mg, 18/35 (51.4%); B[k]F at 4.15 mg, 12/27 (44.4%); IND at 4.15 mg, 21/35 (60%); ANT at 0.83 mg, 19/35 (54.3%); B[ghi]P at 4.15 mg, 4/34 (11.8%); B[a]P at 1.0 mg, 33/35 (94.3%).

Based on data reviewed by IARC, B[a]P, B[b]F, B[j]F, B[k]F, and IND have also induced tumours in mice in several studies after dermal application, subcutaneous, and/or intramuscular administration (IARC, 1983).

The available data also indicate that the five PAHs considered principally here have genotoxic potential in both in vitro and in vivotest systems, and that a mammalian metabolic activation system is necessary for the activity in most cases (see for example IARC, 1983).

2.4.3 Humans

Though there have been increases in lung and skin tumour incidence in populations occupationally exposed to complex mixtures containing principally PAHs, these data have been considered inadequate as a basis for assessment of the weight of evidence of carcinogenicity in humans (IARC, 1983; 1987). Moreover, it is not possible, on the basis of these data, to assess effects of individual PAHs. In addition, the composition of mixtures to which these workers (principally those in coke production, roofing, oil refining, or coal gasification) are exposed may vary considerably from those in the general environment.

2.4.4 Ecotoxicology

Laboratory Studies. Most of the ecotoxicological studies on PAHs have been conducted on aquatic rather than terrestrial organisms. Effects on survival, growth, reproduction, and induction of neoplasms have been observed following exposure to PAHs. Ecotoxicological effects on aquatic organisms have been reviewed by several authors (see Germain et al., 1993). The broad range of the reported effects concentrations reflects the variation in experimental parameters and designs (Germain et al., 1993).

The most sensitive non-neoplastic test endpoint for freshwater organisms determined in the laboratory for each of the selected PAHs are presented in Table 7. Only data for environmentally relevant endpoints (lethality, growth, reproduction) and for test species indigenous to North America or closely related species were considered.

Lethal and sublethal effects of PAHs have been investigated in the laboratory for different species of aquatic invertebrates (Germain et al., 1993). For example, 96-hour LC50s (concentrations estimated to be lethal to 50% of tested organisms after 96 hours of exposure) for the water flea Daphnia pulex were 5 µg/L for B[a]P and 1000 µg/L for naphthalene (Trucco et al., 1983); the 2- day IC50 (concentration estimated to be inhibitory to 50% of tested organisms after 2 days) for Daphnia magna was 430 µg/L for fluorene; and the 30-day LOEC (lowest observed effect concentration) for emergence of the chironomid Chironomus riparius was 600 µg/L for fluorene (Finger et al., 1985). Generally, the short-term lethal and sublethal toxicities for Daphnia are greater for PAHs with higher octanol/water partition coefficients (EAG, 1990).

Polycyclic aromatic hydrocarbons in sediments have been associated with mortality of nymphs of the may fly Hexagenia limbata in toxicity tests conducted under laboratory conditions (Krantzberg and Boyd 1992; Murphy et al., 1993). Sediment grab samples from a station in Hamilton Harbour highly contaminated by PAHs induced nearly 100% mortality even after chemically treating the sediment in order to reduce toxicity associated with metals.

In aquatic vertebrates, death can occur following short- or long-term exposure to PAHs. For example, reported toxic thresholds for naphthalene include a 72-hour LC50 of 240 µg/L for embryos of largemouth bass (Micropterus salmoides), a 23-day LC50 of 120 µg/L for embryos of the rainbow trout (Oncorhynchus mykiss) (Black et al., 1983), and a 96-hour LC50 of 7900 µg/L for fathead minnow (Pimephales promelas) (DeGraeve et al., 1982). For phenanthrene, the reported 96-hour LC50 was 375 µg/L for juvenile

Table 7 Most Sensitive Toxicity Endpoints Reported for Polycyclic Aromatic Hydrocarbons for Freshwater Organisms

Substance

Organism/Life stage

Endpoint

Concentration (µg/L)

Study type

Temperature (°C) /pH

Dissolved O 2 (mg/L) /Hardness (mg CaCO3/L)

Reference

Naphthalene

Rainbow trout, Oncorhynchus mykiss/Early life stage

27-d LC50

110

Flow-through, measured

20.2 to 23.2 /7.4 to 8.1

7.1 to 8.4 /86.8 to 116.3

Black et al.,1983

Acenaphthene

Chironomid, Paratanytarsus sp./Larve

2-d LC50

60

Static, measured

19.0 to 22.4 /7.6

n.a. /46

Lemke and Anderson, 1984

Fluorene

Water flea, Daphnia magna/Adults

14-d LOEC (fecundity 44%

125

Flow-through, measured

25 + 1 /7.2 to 7.4

n.a./270

Finger et al.,1985

Phenanthrene

Rainbow trout, Oncorhynchus mykiss/Embryos

90-d LOEC (survival)

8

Flow-through measured

10.2 /6.9

74.8 /50.4

Call et al.,1986

Anthracene

Fathead minnow, Pimephales promelas/24-h post-hatching

96-h LC50

6.6

Static,

22 + 1 /7.5 + 0.06

6.2 + 0.4 /339 + 13

Oris et al.,1990

Pyrene

Water flea, Daphnia magna/ First instar larvae

48-h LC50

91

Static, Unmeasured

23/n.a.

5-9/n.a.

Abernathy et al., 1986

Fluoranthene

Blue-green algae, Anabaena flos-aquael Exponential growth phase

14-d LOEC (inhibition of growth)

38

Static, measured

25/ n.a.

n.a./ n.a.

Bastian and Toetz, 1982

Benz[a]anthracene

Blue-green algae, Anabaena flos-aquael Exponential growth Phase

14-d LOEC (inhibition of growth)

5

Static, measured

25/n.a.

n.a./n.a.

Bastian and Toetz, 1982

Benzo[a]pyrene

Water flea, Daphnia pulex/2 mm

96-h LC50

5

Static, measured

15 /7.5

n.a./n.a.

Trucco et al.,1983

n.a.- not available

In aquatic invertebrates, B[a]P (0.5 µg/L) was shown to be clastogenic (chromosomal abberations and secondary micronuclei) and teratogenic (deformed gastrula) to eggs of sea urchin (Paracentrotus purpuratus) following exposure for 48 hours (Rose, 1985).

Environmental Effects Observed in situ. In Sydney Harbour, surveys of the benthic fauna indicated a lower diversity and abundance in the more contaminated South Arm than in the Northwest Arm (Wendland, 1979; Hildebrand, 1982). The lowest diversity was observed near the mouth of Muggah Creek. The most common taxa reported in South Arm were polychaetes and sea anemones, accounting for 42% of the total biomass. Many species such as hermit crabs, limpets, and amphipods were present documented for various aquatic biota, including vertebrates, invertebrates, and macrophytes (Trucco et al., 1983; Newsted and Giesy, 1987; Oris and Giesy, 1987; Huang et al., 1993).

Exposure of fish to PAHs has been associated with various teratogenic effects. After a seven-day exposure to naphthalene at a concentration of 239 µg/L, teratogenic deformities were seen in 6% of embryos of largemouth bass (Micropterus salmoides) (Black et al., 1983).

Deformities were also observed in 43% of rainbow trout (Oncorhynchus mykiss) exposed for 27 days to 85 µg/L of phenanthrene (Black et al., 1983). Eye anomalies were detected in 7% of rainbow trout alevin exposed for 36 days to 0.2 µg/L of B[a]P and in 17% of those exposed to 0.3 µg/L. The lowest concentration at which effects were observed was 0.1 µg/L (36-day LOEC for eye anomalies in 2% of rainbow trout fry) (Rose et al., 1984). In addition, clastogenic effects resulting from DNA damage in fish were reported. For example, a formation of numerous secondary micronuclei is observed in red cells of rainbow trout embryos exposed for 36 days to 0.1 µg/L of B[a]P (Rose et al., 1984).

The growth of terrestrial flora may be stimulated or inhibited by PAHs, depending on the concentrations to which the plants are exposed. B[a]P, for example, stimulates growth in corn, soya, and wheat at soil concentrations of up to 50 mg/kg (d.w.), and inhibits growth at concentrations above this level (Slooff et al., 1989). Polycyclic aromatic hydrocarbons are of low toxicity to terrestrial plants because plants can degrade PAHs and even synthesize certain PAHs (Slooff et al., 1989).

The toxicity of an artificial mixture of 18 PAHs was tested on embryos of four bird species: chicken (Gallus domesticus), turkey (Meleagris gallopava), mallard duck (Anas platyrhynchos), and common eider (Somateria mollissima) (Brunström et al., 1990). The doses injected in the yolk sacs were 2.0 and 0.2 mg/kg of egg. Mortality was significantly increased in all four species at the higher dose, and in mallards at the lower dose. When the 18 compounds were tested individually, only benzo[k]fluoranthene, B[a]P, and indeno[1,2,3-cd]pyrene caused significant increases in mortality in chicken embryos at a dose of 2.0 mg/kg of egg. The B[k]F also caused significant increases in mortality of the other three species at 0.2 mg/kg of egg. The various PAHs thus appeared to differ substantially in toxicity to bird embryos.

Genotoxic and neoplastic effects have been reported in both vertebrate and invertebrate organisms following metabolism of certain PAHs. Of the compounds selected for this environmental assessment, such effects were observed under laboratory conditions for B[a]P, phenanthrene, and naphthalene (Shugart, 1988; Black et al., 1983; Hose et al., 1984; Hose, 1985).

Exposure of fish to PAHs can lead to clastogenic effects resulting from DNA damage. For example, a formation of numerous secondary micronuclei is observed in red cells of rainbow trout embryos exposed for 36 days to 0.1 µg/L of B[a]P (Hose et al., 1984). A 24-week exposure (with two six-hour periods per week) to 150 to 240 µg/L of B[a]P caused hepatic neoplastic tumours in 10% of the guppies tested (Hawkins et al., 1990).

Injecting organic extracts from Hamilton Harbour sediments into the sac fry of rainbow trout (Oncorhynchus mykiss) induced malignant hepatic tumours similar to those induced by PAHs such as B[a]P and 7,12-dimethylbenzanthracene (Metcalfe et al., 1990). These sediment extracts contained mainly 2 to 16 PAHs (0.01 to 52 µg/g for individual compounds) with traces of polychlorinated biphenyls (0.020 µg/g) and organochlorines such as hexachlorobenzene (0.0006 µg/g). The extracts were mutagenic in Sahnonella typhimurium and generated aromatic DNA adducts in a mouse cell culture.

Stein et al. (1990) have reported that sediment-associated PAHs are biologically available to two flatfish, English sole (Parophrys vetulus) and starry flounder (Platichthys stellatus). Under laboratory conditions, English sole and starry flounder metabolized B[a]P to intermediates that bind to hepatic DNA and develop chemically induced hepato-carcinogenesis (Stein et al., 1990).

English sole caught in an area of Puget Sound with minimal contamination by PAHs were exposed parenterally to organic extracts of urban marine contaminated sediment of Eagle Harbour and to B[a]P (Schiewe et al., 1991). Polycyclic aromatic hydrocarbons are the principal constituents (90%) of these sediment extracts. Following 13 exposures (every four weeks) over one year to 30 mg of extract/kg (b.w.), 75 mg of extract/kg (b.w.), or 12 mg of B[a]P/kg (b.w.), and after a six-month holding time, fish were examined for hepatic lesions. Preneoplastic foci of cellular alteration were observed in 13.6% of the fish exposed to 30 mg of extract/kg (b.w.), in 13.8% of those exposed to 75 mg of extract/kg (b.w.), and in 13.3% of those exposed to 12 mg B[a]P/kg (b.w.); in the controls, no alteration was observed. Controls included sediment extracts from an area of Puget Sound with minimal contamination, carrier extracts, and an untreated group of sole. The lesions induced in the laboratory were not distinguishable from those observed in English sole caught in Eagle Harbour, where the contaminated sediments were sampled.

In aquatic invertebrates, B[a]P (0.5 µg/L) was shown to be clastogenic (chromosomal abberations and secondary micronuclei) and teratogenic (deformed gastrula) to eggs of sea urchin (Paracentrotus purpuratus) following exposure for 48 hours (Hose, 1985).

Environmental Effects Observed in situ. In Sydney Harbour, surveys of the benthic fauna indicated a lower diversity and abundance in the more contaminated South Arm than in the Northwest Arm (Wendland, 1979; Hildebrand, 1982). The lowest diversity was observed near the mouth of Muggah Creek. The most common taxa reported in South Arm were polychaetes and sea anemones, accounting for 42% of the total biomass. Many species such as hermit crabs, limpets, and amphipods were present in Northwest Arm but not in South Arm. The concentration of total PAHs in sediments varied from 13 to 2800 µg/g (d.w.) in the South Arm and from 2.5 to 8.2 µg/g in the Northwest Arm (Matheson et al., 1983). Although high concentrations of heavy metals were also reported, no relationship was observed between the concentration of metal and faunal distribution. The colour and the strong hydrogen sulphide (H2S) odour of the sediments suggested that many metals were in sulphide forms, which are less available to biota (Hildebrand, 1982). The sediment ranged from a brown, sandy mud to a "black sludge" mud with a strong tar odour.

Stations with this strong odour had the lowest faunal diversity (Hildebrand, 1982). Diversity and abundance of benthic organisms in Sydney Harbour increased with distance from the mouth of the creek. The tidal flats of Muggah Creek [total PAHs: 2830 µg/g (d.w.)] were completely devoid of life.

Changes in populations of benthic invertebrates have been found to be correlated with concentrations of PAHs in sediments in Kettle Creek in Port Stanley, Ontario (AEL, 1988; Canviro, 1988). This stream drains the site of a former oil gasification complex and its sediments have been classified according to three levels of PAH contamination: <9 µg/g, 12 to 40 µg/g, and >80 µg/g (d.w.). A significantly lower population of oligochaetes (benthic worms), generally considered to be tolerant to the contamination, as well as a general impoverishment of the benthic fauna in the areas with the most contaminated sediments were observed.

In studies at several sites in Canada, including the Great Lakes and Vancouver Harbour, and locations in the U.S. (see Table 8), increases in the incidence of neoplasms in fish in both marine and fresh water have been associated with exposure to PAHs.

Baumann and Whittle (1988) reported that fish in the Great Lakes had higher tumour frequencies in areas with high sediment concentrations of B[a]P and fluorene. In Hamilton Harbour, where B[a]P and fluorene concentrations were greater than 1 and 10 µg/g (d.w.), respectively, 30% of white sucker (Catostomus commersoni) had carcinogenic papillomas. In the eastern side of Lake Ontario, B[a]P and fluorene were detected between 0.1 and 1 µg/g and 6% of the white suckers had papillomas. In Lake Superior, where B[a]P and fluorene were lower than 0.1 µg/g, papillomas were not observed. Brown bullhead (Ictalarus nebulosus) caught in Black River (Ohio) had B[a]P levels greater than 5 ng/g (w.w.) in tissues; liver tumours were observed in 25 to 35% of the fish. No tumours were observed for those caught in Lake Buckeye, a reference site.

Baumann and Mac (1988) detected hepatic tumours in 30 to 38% of brown bullhead in the Black River, Ohio, and 13 to 19% of those in the Cuyahoga River, Ohio. High concentrations of PAHs were detected in the sediments of these two rivers, individual compounds ranging from 10 to 1000 µg/g (d.w.) in the Black River and from 0.1 to 1 µg/g (d.w.) in the Cuyahoga River. The sediments contained benz[a]anthracene and B[a]P. Polycyclic aromatic hydrocarbons were found in the bullheads in the two rivers as well as high concentrations of PAH metabolites in the bile of bullheads in the Black River. Organochlorine compounds and metals were not found in high concentrations, compared to other sites where no hepatic tumours were detected.

In a study of carcinogenesis in fish, Maccubbin et al. (1985) reported that the stomach contents of white sucker (Catostomus commersoni) inhabiting the bottom of Lake Erie may contain appreciable concentrations of PAHs (0.78 µg/g). The concentrations measured in the fish were correlated with the concentrations in the sediments where the fish were taken. Moreover, the chromatographic profiles for the PAHs in the sediments and the fish were similar. Therefore, it is likely that a portion of the PAHs in the fish was taken up by ingestion of contaminated sediments during feeding. In another study, two species of fish from the Buffalo (New York) and Detroit (Michigan) rivers along the Great Lakes had a high incidence of liver cancer (27 and 20% respectively; Maccubbin et al., 1990). Sediments in these two rivers are highly contaminated by many chemicals, including B[a]P (0.12 to 17.7 µg/g), fluorene (0.12 to 34.6 µg/g), and pyrene (0.12 to 38.8 µg/g). To determine the role of these compounds in the etiology of the tumours, the authors measured DNA adducts in fish livers. Several genotoxic aromatic compounds were identified in the nucleotide adducts. An analysis of fish bile revealed recent exposure to polyaromatic compounds.

Table 8 Field Population Studies in Fish1

Species

Location

Observed effects and frequency of affected organisms (%)

Ranges of concentrations in sediments (µg/g dry weight)

References

Fresh water

White sucker, Catostomus commersoni

Hamilton Harbour (Ont.)

Papillomas
30

B[a]P> 1;

Baumann and Whittle, 1988

Lake Ontario (eastern side)

6

Flu> 10

Lake Superior

< 1

B[a]P <0.1;
Flu <0.1

Brown bullhead, Ictalurus nebulosus

Black River (Ohio)

Liver tumours
25 to 35

B[a]P > 5 ng/g in tissues; Flu > 1 µg/g in organs

Baumann and Whittle, 1988

Lake Buckeye (Ohio) reference site

0

 

Brown bullhead, Ictalurus nebulosus

Black River (Ohio)

Liver tumours 30 to 38

individual PAHs: 10 to 1000

Bauman and Mac, 1988

Cuyahoga River (Ohio)

13 to 19

individual PAHs: 0.1 to 1

Rainbow trout, Oncorhynchus mykiss

Hamilton Harbour (Ont.)

Hepatic neoplasms
8.9

Phe: nd to 40.8; Flu: nd to 52; 16 PAHs detected

Metcalfe et al., 1990

Oakville Creek (Ont.)

0

Phe: nd to 0.04; Flu: nd to 0.03; 4 PAHs detected

Brown bullhead, Ictalurus nebulosus

Buffalo River (N.Y.)

Tumours
27

B[a]P: 3; Flu: 13; Pyr: 17

Maccubin et al., 1990

Walleye, Stizostedion vitreum

Detroit River (Michigan)

Hepatic neoplasms
20

B[a]P:0.12 to 17.7; Flu: 0.12 to 34.6; Pyr: 0.12 to 38.8

Maccubin et al., 1990

Salt Water

English sole, Parophrys vetulus

Puget Sound (Wash.) Mukilteo Point

Hepatic neoplasms 7.5

Aromatic hydrocarbons 7.8 to 33

Malins et al., 1985a; b

Eagle Harbour

27

2.8 to 120 (90% are PAHs)

President Point (reference site)

0

1.1

English sole, Parophrys vetulus

Puget Sound (Wash.)

One or more types of lesions

B[a]P: 2.3; Flu: 59

Krahn et al., 1986

Eagle Harbour

90

B[a]P: 0.073; Flu: 0.44

Dunawish Waterway

67.2

B[a]P: 0.005; Flu: 0.019

Useless Bay

6.2

 

English sole, Parophrys vetulus

Vancouver Harbour (B.C.)

Hepatic tumours

Total PAHs

Goyette et al., 1988

Port Moody Arm

75

3.1 to 36.8

Burrard Narrow

30

nr

Central Channel

8

2.5

1 B[a]P- Benzo[a]pyrene;

Flu- Fluoranthene;

Phe- Phenanthrene;

Pyr- Pyrene;

nd- not detected;

nr- not reported

In marine waters, cancerous hepatic lesions have been observed in English sole (Parophrys vetulus) in Vancouver Harbour (Goyette et al., 1988; Brand and Goyette, 1989). The authors reported a high incidence of tumours (75%) where sediments were highly contaminated by hydrocarbons and PAHs (from 3.1 to 6.6 µg/g of total PAHs), including B[a]P and the [b] and [k] isomers of benzofluoranthene. Moderate frequencies of tumours (8 and 30%) were observed at other sites in the harbour.

Malins et al. (1984) studied the frequency of hepatic neoplastic and other lesions in English sole (Parophrys vetulus) collected in Puget Sound in Washington. They examined the relationship between the aromatic compounds present in sediments and the diet of the fish (Maims et al., 1984; 1985a;b). They observed a high frequency (27%) of cancerous hepatic lesions in English sole in Eagle Harbour of Puget Sound. The sediments in the harbour contained particularly high concentrations of aromatic hydrocarbons (from 2.8 to 120 µg/g, depending on the site), 90% of which were PAHs. At a control site with much lower contaminant concentrations (aromatic content: 1.1 µg/g), the bottom-dwelling fish showed no such hepatic lesions. Moreover, higher concentrations of aromatic hydrocarbons were observed in the stomach contents of sole from Eagle Harbour, along with higher PAH metabolite concentrations in their bile. McCain et al. (1990) also found higher PAH levels in the stomach content and higher B[a]P metabolites in the bile of juvenile chinook salmon (Oncorhynchus tsawytscha) caught in the Duwamish Waterway (discharging in Puget Sound near Seattle, Washington) than those found at the reference site. The salmon collected in Duwamish Waterway showed a reduced immunological memory; the consequence of this reduction in disease resistance is unknown (Arkoosh et al., 1991).

Krahn et al. (1986) statistically compared mean PAH metabolite concentrations in the bile of English sole with the incidence of hepatic tumours in the same fish collected at 11 sites in Puget Sound. They found a significant positive correlation between the incidence of neoplasms and other liver lesions and the concentration of PAH metabolite in the fish. These findings provided additional evidence of the relationship between PAHs and cancerous hepatic tumours in bottom-dwelling fish. Since PAHs are rarely the only contaminants present in sediments, their specific role in altering the health of fish is not fully understood. In a review of studies published on liver carcinogenicity, Myers (1990) noted that the livers of English sole collected from Puget Sound were affected by a broad range of hepatic lesions. The author advanced the hypothesis that these lesions are related to PAH contaminants, on the basis of the statistical associations drawn between PAH levels in sediments and lesion frequencies, and the significant correlations between lesion frequency and the level of aromatic compound metabolites (including numerous PAH metabolites) present in the bile.

Shugart (1990) analyzed B[a]P DNA adducts in the brains of belugas (Delphinapterus leucas) in the St. Lawrence and the Mackenzie River estuaries. Concentrations of 69 to 206 ng of B[a]P adducts per gram of DNA were observed in belugas from the St. Lawrence, while no adducts where detected in those from the Mackenzie.

In terrestrial mammals, higher concentrations of B[a]P adduct (geometric mean of 69.9 pmol of B[a]P adduct/g of albumin and 7.6 pmol of B[a]P adduct/g of hemoglobin) have been detected in the blood proteins of wild marmots (Marmota monax) captured 10 km from an aluminum smelter in the Saguenay region than in marmots from a control area (geometric mean of 7.6 pmol of B[a]P adduct/g of albumin and 0.4 pmol of B[a]P adduct/g of hemoglobin) located in Saint Roch des Aulnaies, Quebec (Blondin and Viau, 1991; 1992).


1 Bay regions occur in angularly fused benzo-rings; for example, in benzo(a)pyrene, the sterically hindered region between the 10- and 11-positions constitutes the hay region.

3.0 Assessment of "Toxic" Under CEPA

3.1 CEPA 11(a) Environment

Polycyclic aromatic hydrocarbons are emitted into the Canadian environment from both natural and anthropogenic sources. Forest fires are the single most important natural source of PAHs in Canada. Since forest fires are separated both in time and space across the country, and release PAHs to the atmosphere, a dispersive medium, they do not result in continuous or chronic release of, or exposure to, PAHs in any specific area. Anthropogenic sources are numerous and result in emissions of PAHs into all environmental compartments. The largest anthropogenic sources of PAHs to the atmosphere are aluminum smelters using the Horizontal Stud Söderberg smelting process. Other important sources to the atmosphere include residential heating with wood, agricultural burning, and burning of wood wastes. Major sources of PAHs to the aquatic and soil environments include creosote-treated products, spills of petroleum products, metallurgic and coking plants, and deposition of atmospheric PAHs. Most anthropogenic sources are point sources releasing PAHs regularly, whether continuously or intermittently (e.g., smelters, coking and metallurgic plants), or diffuse sources releasing PAHs in a circumscribed place or time (e.g., urban sources such as vehicle emissions or wood burning), and can therefore result in chronic exposure of biota to PAHs.

With the exception of some of the lighter compounds, PAHs are relatively non-volatile and of low solubility in water. In the atmosphere, they are mostly found adsorbed to particulate matter that can be deposited to water and soil. In aquatic systems, PAHs adsorb to suspended matter in the water column and are deposited in sediments where they will persist; heavy PAHs, such as benzo[a]pyrene, can persist in sediments for several years. Polycyclic aromatic hydrocarbons are removed from soils mainly by volatilization and by microbial activities. Thus, with most PAHs being emitted into the atmosphere, the atmosphere is the major medium for the transport of PAHs, while sediments are the major environmental sink for these compounds.

Polycyclic aromatic hydrocarbons have been measured in air, soil, groundwater, fresh and marine surface water, sediments, and terrestrial and aquatic biota in Canada. Concentrations in all media are generally higher in areas close to known anthropogenic sources of PAHs.

Considerable data are available on toxicity of PAHs to aquatic organisms. In reviewing these data, emphasis was placed on ecologically relevant test end-points for species occurring in Canada or for related species. Since toxicity of many PAHs is photo-activated, it is likely that those studies done in the dark or under conditions of ambient laboratory lighting underestimate the toxicity of PAHs in surface waters.

Concentrations of PAHs in Canada in ambient waters (i.e., not close to anthropogenic sources) are relatively low. The highest concentration for total PAHs in ambient water was 1.8 µg/L, measured shortly after ice breakup in the Mackenzie River.

Concentrations of PAHs reported in ambient water are a few to several orders of magnitude lower than the lowest effects thresholds reported.

The highest concentrations of PAHs in the water column were reported by Wan (1993) in drainage ditches along railway lines (total PAHs of 5356 µg/L) and utility rights-of-way (total PAHs of 23 µg/L), where PAHs were released from utility poles or railway ties treated with creosote. Based on limited sampling, median concentrations in railway ditches with poles (Table 4) were close to or greater than the effects thresholds recorded for aquatic biota (Table 7) for eight of the PAHs selected for this assessment: acenaphthene, fluorene, phenanthrene, anthracene, pyrene, fluoranthene, benzo[a]anthracene, and benzo[a]pyrene. Mean concentrations in utility right-of-way ditches (Table 4) were close to or greater than the effects thresholds recorded for aquatic biota (Table 7) for three of the PAHs selected for this assessment: phenanthrene, anthracene, and pyrene. No data were identified that would allow the estimation of effects to biota in receiving waters.

Therefore, based on concentrations of PAHs recorded in surface water in Canada and on laboratory-derived toxic effects thresholds, PAHs in ambient water likely do not directly affect aquatic biota, while PAHs in railway and utility ditches could adversely affect biota.

Numerous sites where sediments are contaminated with PAHs were identified in Canada. Sites discussed in this section were selected because concentrations of PAHs in sediments and the biota at the sites are relatively well characterized.

The discharge of wastes from a coking facility resulted in the formation of the tar ponds on the tidal flats of Muggah Creek in Sydney Harbour, Nova Scotia. Although the coking facility was closed in 1988, it is estimated that close to 1 tonne/year of PAHs are released in the estuary by tidal influence. It is recognized that the tidal flats of Muggah Creek contaminated with PAHs are devoid of living organisms. Generally, both the diversity and abundance of benthic organisms increase with distance from Muggah Creek. Also, it has been demonstrated that benthic communities in the more contaminated South Arm of the estuary were less diversified than those from the Northwest Arm.

Hamilton Harbour is a highly contaminated site in the Great Lakes, with the steel mills located close to the harbour being the main source of PAHs. Although they are not the only contaminants at this site, PAHs have been associated with mortality of nymphs of the mayfly, Hexagenia limbata, in toxicity tests conducted under laboratory conditions (Krantzberg and Boyd, 1992; Murphy et al., 1993). Sediments from a location in the harbour highly contaminated by PAHs induced nearly 100% mortality even after the sediments were chemically treated to reduce toxicity associated with metals.

Kettle Creek at Port Stanley, Ontario, is another site heavily contaminated with PAHs because of the presence of a former oil gasification complex. AEL (1988) reported a correlation between the concentrations of PAHs in the sediments (Canviro, 1988) and the abundance of the oligochaetes Limnodrilus cervix and L. hoffineisteri.

These worms are generally recognized as tolerant to contaminated environmental conditions and were the most abundant species in the study area.

Therefore, evidence from field studies and from laboratory studies using field samples indicate that biota are adversely affected at various Canadian sites contaminated by PAHs of different industrial origins.

Under laboratory conditions, neoplastic and genotoxic effects have been associated with exposure to PAHs for both terrestrial and aquatic organisms. Field evidence is supportive of such an association. In terrestrial systems, the only field evidence available is limited to that demonstrating exposure and possible preliminary stages of chemically induced carcinogenesis. For example, an increase has been observed in the incidence of B[a]P DNA adducts in the blood proteins of marmots captured downwind of an aluminum smelter in the Saguenay region. They have also been observed in the brain of aquatic mammals (i.e., beluga whales) in the St. Lawrence River. The ultimate neoplastic and genotoxic effects to wildlife resulting from chronic exposure to PAHs are not known.

In the case of bottom-dwelling fish, the evidence demonstrating that exposure to B[a]P is associated with cancer induction is stronger. Injecting organic extracts from Hamilton Harbour sediments into the sac fry of rainbow trout induced malignant hepatic tumours similar to those induced by PAHs such as B[a]P and 7,1,2-dimethylbenzanthracene (Metcalfe et al., 1990). Under laboratory conditions, English sole and the starry flounder metabolized B[a]P to intermediates that bind to hepatic DNA and develop chemically induced hepato-carcinogenesis (Stein et al., 1990). Also under laboratory conditions, Schiewe et al. (1991) demonstrated that English sole parenterally exposed to extracts from sediments contaminated with high levels of aromatic hydrocarbons or to benzo[a]pyrene developed neoplastic hepatic lesions. In Canada, similar cancerous hepatic lesions have been observed in English sole in Vancouver harbour (Goyette et al., 1988; Brand and Goyette, 1989). These authors reported tumour frequencies of up to 75% in English sole caught in Port Moody where sediments are highly contaminated by PAHs, including carcinogens such as B[a]P and benzofluoranthene ([b] and [k] isomers). Lower tumour frequencies were reported in other areas of the harbour.

Therefore, laboratory and field evidence indicates that PAHs induce neoplastic and genotoxic effects in aquatic biota. Data for mammals indicate that these animals may be susceptible to such effects, but no studies were identified documenting such effects in wild mammals.

Therefore, based on the neoplastic, genotoxic, and population-level effects observed in aquatic biota at sites contaminated with PAHs across Canada, PAHs are considered to be entering the environment in a quantity or concentration or under conditions that are having a harmful effect on the environment.

3.2 CEPA 11(b) Environment on Which Human Life Depends

Polycyclic aromatic hydrocarbons generally only weakly absorb light of wavelengths critical to global warming. Unlike substances associated with depletion of stratospheric ozone (Firor, 1989), PAHs are non-halogenated compounds of low to moderate persistence in the atmosphere. Given these properties and the low steady-state concentrations of PAHs in the atmosphere, they are not considered to contribute significantly to stratospheric ozone depletion, global warming, or ground-level ozone formation.

Therefore, based on available data, PAHs are not considered to be entering the environment in a quantity or concentration or under conditions that constitute a danger to the environment on which human life depends.

3.3 CEPA 11(c) Human Life or Health

Based on the limited available data on concentrations of PAHs in drinking water in Canada, it is likely that it contributes negligibly to total intake. Although data are available on the concentrations in food of several of the PAHs considered principally in this assessment, analytical recoveries were poor. Moreover, levels in food vary considerably depending on cooking habits and food sources. Available data were also inadequate to develop estimates of carcinogenic potency of a broad range of PAHs in the gastrointestinal tract following ingestion. The current assessment focuses, therefore, on exposure in ambient air.

Owing to the possible confounding by concomitant exposure to other substances that may have contributed to observed effects, available epidemiological data are considered inadequate to assess the health risks (including carcinogenicity) of PAHs in humans. The five PAHs considered principally in the assessment of potential risks to human health (benzo[a]pyrene, benzo[b]fluoranthene, benzo[j] fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd]pyrene) are classified in Group II ("Probably Carcinogenic to Humans") of the classification scheme for carcinogenicity developed for the determination of "toxic" under paragraph 11(c) of CEPA (ERD, 1994). This classification is based primarily on the results of carcinogenicity bioassays in which PAHs have been administered to experimental animals by inhalation (B[a]P only) and dermal application, and supporting data from studies involving routes of administration less relevant to environmental exposure (i.e., direct introduction into body cavities or tissues) and information on genotoxicity and mechanism of action. For such substances, where data permit, the estimated total daily intake or concentrations in relevant environmental media are compared to quantitative estimates of carcinogenic potency to characterize risk and provide guidance for further action (i.e., analysis of options to reduce exposure) (ERD, 1994). The carcinogenic potency is expressed as the dose or concentration that induces a 5% increase in the incidence of relevant tumours (TD0.05).

In order to calculate the exposure/carcinogenic potency indices (EPIs) for PAHs, two approaches have been adopted. One is based on the assumption that the carcinogenic potency of each of the components of a mixture of PAHs is equivalent to that of B[a]P on a weight basis; the other is based on calculation of relative carcinogenic potencies for several PAHs for which the data base is considered sufficient. Both of these approaches have considerable limitations (as described in this section); moreover, exposure from media other than air has not been taken into account1. The values are, therefore, considered "interim", pending generation of more suitable data for quantitative estimation of exposure and potency.

For the B[a]P equipotency approach, the TD0,05 for inhaled B[a]P has been estimated based on multi-stage modelling of the respiratory tract tumours in Syrian golden hamsters in the study reported by Thyssen et al. (1981). The highest exposure group was eliminated from the analysis due to the low average survival time. The TD0.05 for B[a]P estimated in this manner is 1.57 mg/m3.

Estimated interim EPIs for populations residing in the vicinity of aluminum smelters (Kitimat, Jonquière, and Shawinigan) developed based on this likely conservative approach and the total concentrations of the 17 to 28 PAHs at these locations (156 to 1690 ng/m3) determined in the national survey (Ringuette et al., 1993) range from 1.0 x 10-4 to 1.1 x 10-3 (156 to 1690 ng/m3 ÷1.57 mg/m3). Therefore, based on this approach, the priority for analysis of options to reduce exposure solely on considerations of potential health effects would be high.

This approach is necessarily limited to only those PAHs that have been monitored in ambient air and does not account, therefore, for all PAHs to which the general population is exposed. Also, the small increases in the incidence of tumours at sites other than the respiratory tract in the study by Thyssen et al. (1981) have not been taken into account. Nevertheless, it is likely that the carcinogenic potency of PAH mixtures is considerably overestimated based on the assumption that the potency of all PAHs is similar to that of B[a]P since the weight of evidence of carcinogenicity for each of the monitored PAHs has not been considered. Moreover, on the basis of studies in which the carcinogenicity of various PAHs has been examined principally by application to the skin of mice, the potency of B[a]P has consistently been greater than that of most other compounds (see, for example, U.S. EPA, 1982). In addition, there are some data that indicate that some PAH-containing mixtures containing B[a]P are less potent than B[a]P alone (Slaga et al., 1980). Furthermore, based on determination of the incidence of tumour formation after exposure to a combination of two or more PAHs, effects were not additive (see, for example, Grimmer, 1983).

With respect to the alternative approach, however, data available to serve as a basis for estimating the relative potencies of various PAHs in the target organ (i.e., the lung) relevant to one of the principal media of exposure in the general environment for which data were sufficient to estimate exposure (i.e., air), are very limited. Indeed, only two adequate studies have been identified relevant to estimation of potency in the lung for a sufficiently broad range of PAHs, those of Thyssen et al. (1981), in which there was clear evidence of an exposure-response relationship between inhaled B[a]P and respiratory tract tumours in Syrian golden hamsters, and Deutsch-Wenzel et al. (1983), in which there were exposure-response relationships for epidermoid carcinomas and pleomorphic sarcomas in Osborne-Mendel female rats administered B[a]P, B[b]F, IND, B[k]F, B[j]F, and ANT by pulmonary implantation.

For the relative potency approach, carcinogenic potencies were estimated for the selected PAHs on the basis of multi-stage modelling of tumour incidence (epidermoid carcinomas) in Osborne-Mendel rats exposed in the study by Deutsch-Wenzel et al. (1983) to each of the PAHs compared to that in the "solvent-vehicle"-exposed controls. Values were based on the dose that induced a 5% increase in the incidence of relevant tumours calculated on the basis of each fit. The potencies of B[b]F, B[j]F, B[k]F, and IND relative to that of B[a]P were computed by dividing the dose calculated to be associated with a 5% increase in tumours for B[a]P by those for each compound. The relative carcinogenic potency factors estimated on this basis were 0.06 for B[b]F, 0.05 for B[j]F, 0.04 for B[k]F, and 0.12 for IND (and 1 for B[a])P).

B[a])P equivalents for the five PAHs addressed in this assessment were calculated by multiplying the concentrations in ambient air by the relative carcinogenic potency factors derived above (Table 9). The values for total B[a]P equivalents/m3 range from 2.72 to 48.98 ng/m3 for cities near aluminum smelters using Horizontal or Vertical Stud Söderberg processes, 2.42 to 9.12 ng/m3 for cities where wood stoves are commonly used, 0.13 to 2.25 ng/m3 for urban cities, and 0.11 ng/m3 for a rural area.

The EPIs have been calculated on the basis of the TD0.05 for inhaled B[a]P estimated based on multistage modelling of the respiratory tract tumours in Syrian golden hamsters in the study reported by Thyssen et al. (1981), as described above for the B[a]P equipotency approach, and the B[a]P equivalents in ambient air at different types of sites in Canada for the five specified PAHs. For example, "interim" EPIs for the general population in Canada living near aluminum smelters that use the Horizontal or Vertical Söderberg process (i.e., the population that has the highest exposure to the selected PAHs) range from 1.7 x 10-6 to 3.1 x 10-5 (2.72 to 48.98 ng of B[a]P equivalent/m3 ÷ l .57 mg of B[a]P/m3). On the basis solely of considerations of potential health effects, based on this approach, therefore, the priority for further action (i.e., analysis of options to reduce exposure) would be considered to be moderate.

This approach to estimating relative potency is considered to offer several advantages over those reported previously by, for example, Thorslund and Charnley (1988) and Slooff et al. (1989). Estimation of potency is restricted to the target organ for the route of exposure for which sufficient data are available to estimate exposure of the general population (i.e., inhalation in air). In addition, potency estimates for the other PAHs considered here can be compared directly to that for B[a]P in the same experimental model. Moreover, the weight of evidence for carcinogenicity and study quality has been taken into account, factors that have often not been adequately considered in the previous estimations of potency for PAHs. However, since the type of respiratory tumours induced in the study by Thyssen et al. (1981) were not specified, it is unknown whether they were similar to those induced following implantation. In addition, only tumours of the respiratory tract have been taken into account in estimating carcinogenic potency. It should also be recognized that these "interim" EPIs based on relative carcinogenic potencies include only a very small proportion of PAHs in the general environment, for which the available data on concentrations and relative carcinogenic potency were considered adequate [i.e., 0.1% to 23.5% of a total of 13 PAHs determined at various locations in a recent survey of ambient air in Canada (Dann, 1992a)].

Table 9 Levels of PAHs in B[a]P Equivalents/m3 in Canadian Ambient Air

PAH*

Relative carcinogenic potency factors

ng B[a]P equivalent/m3 **

A***

B***

C***

D***

B[a]P

1.0

2.06 to 36.37

1.93 to 7.66

0.09 to 1.74

0.08

B[b]F

0.06

0.27 to 5.59

0.19 to 0.43

0.01 to 0.15

0.01

B[j]F

0.05

0.10 to 2.08

0.07 to 0.23

0.01 to 0.06

0.01

B[k]F

0.04

0.06 to 1.27

0.04 to 0.10

<0.01 to 0.07

<0.01

IND

0.12

0.23 to 3.67

0.19 to 0.70

0.02 to 0.23

0.01

Total

2.72 to 48.98

2.42 to 9.12

0.13 to 2.25

0.11

* B[a]P = benzo[a]pyrene; B[b]F = benzo[b]fluoranthene; B[j]F = benzo[j]fluoranthene; B[k]F = benzo[k]fluoranthene; IND = indeno [1,2,3-cd]pyrene

** ng B[a]P equivalent/m3 = mean concentrations in ambient air (ng/m3) x relative potency factor

*** A = cities near aluminum smelters (Jonquière, Kitimat, Shawinigan) B = wood stove burning (Whitehorse, Sept-Îles) C = cities (Montreal, Sydney, Toronto, Vancouver, Windsor, Winnipeg) D = rural area (Walpole Island)

Based on the two approaches presented here, both of which are limited by the inadequacies of the data base, the resulting "interim" EPIs and, hence, priority for further action (i.e., analysis of options to reduce exposure) would be moderate to high, on the basis solely of consideration of the potential effects on health of selected polycyclic aromatic hydrocarbons.

On the basis of available data, the PAHs benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd] pyrene are classified as "Probably Carcinogenic to Humans" (i.e., substances for which there is considered to be some probability of harm for the critical effect at any level of exposure) and considered to be "toxic" as defined under paragraph 11(c) of the Canadian Environmental Protection Act. This approach is consistent with the objective that exposure to substances for which the critical effect is considered not to have a threshold should be reduced wherever possible and obviates the need to establish an arbitrary de minimis level of risk for determination of "toxic" under the Act.

Therefore, based on available data, the polycyclic aromatic hydrocarbons benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd]pyrene are entering the environment in a quantity or concentration or under conditions that may constitute a danger to human life or health.

3.4 Conclusion

Therefore, based on available data, polycyclic aromatic hydrocarbons are entering the environment in a quantity or concentration or under conditions that are having or may have a harmful effect on the environment. They are not considered to be entering the environment in a quantity or concentration or under conditions that constitute or that may constitute a danger to the environment on which human life depends. The polycyclic aromatic hydrocarbons benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[1,2,3-cd]pyrene are considered to be entering the environment in a quantity or concentration or under conditions that may constitute a danger to human life or health.


1 On the basis of the limited available data on concentrations of PAHs in food (Das, 1987), the estimated intake (based on B[a]P equivalents) of three of the selected PAHs (i.e., B[a]P, B[b]F, and B[k]F) for an adult (ERD, 1994) is approximately two-fold greater than that inhaled; however, it is likely that there are considerable variations in toxicokinetics and potency by ingestion versus inhalation.

4.0 Recommendations for Research

To permit a more comprehensive assessment, it is recommended that the following research be considered.

  • Development of experimental and theoretical models for estimation of the carcinogenic potency of a broad range of PAHs. In particular, bioassays in which animals are exposed by routes by which the general population is principally exposed (i.e., inhalation and ingestion) to a range of PAHs selected to characterize a range of carcinogenic potencies are desirable (high priority).
  • Development of analytical methods and additional characterization of concentrations of a broader range of PAHs in ambient air, drinking water, food, and soil (high priority).
  • The highest concentrations of PAHs in water were reported from ditches along railway and utility lines. Given the prevalence of railway and utility lines throughout the country, it is recommended that data on effects on biota in the ditches and in receiving waters should be determined (medium priority).
  • The neoplastic and related effects of chronic exposure to PAHs should be determined for terrestrial organisms, notably mammals, in areas adjacent to major atmospheric sources of PAHs (low priority).
  • To better characterize the effects of PAHs in sediments on fish, fish should be exposed to individual sediment-bound PAHs using environmentally relevant exposure protocols (low priority).

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