4.0 OCCURRENCE in the ENVIRONMENT

4.1 Natural sources

In nature, PAHs may be formed three ways: (a) high temperature pyrolysis of organic materials, (b) low to moderate temperature diagenesis of sedimentary organic material to form fossil fuel, and (c) direct biosynthesis by microbes and plants (Neff, 1979).

4.1.1 Fires

Forest fires, prairie fires, and agricultural burning contribute the largest volumes of PAHs from a natural source to the atmosphere. The actual amount of PAHs and particulates emitted from these sources varies with the type of organic material burned, type of fire (heading fire vs. backing fire), nature of the blaze (wild vs. prescribed; flaming vs. smouldering), and intensity of the fire. PAHs from fires tend to sorb to suspended particulates and eventually enter the terrestrial and aquatic environments as atmospheric fallout (Eisler, 1987).

In the atmosphere, PAHs may undergo photolytic and chemical (ozone) transformations. However, most of the material does not degrade quickly in the atmosphere and thus may reside in the environment for extended periods of time. During this atmospheric entrainment, winds may distribute these particle-sorbed PAHs in a global manner such that they appear even in remote areas of the Arctic or Antarctica. There has been some contention, however, that the world-wide distribution of PAH may actually be due to natural PAH sources in proximity to these remote areas (Clarke and Law, 1981; Platt and Mackie, 1981).

4.1.2 Fossil fuels

PAHs occur naturally in bituminous fossil fuels, such as coal and crude oil deposits, as a result of diagenesis (i.e., the low temperature, 100-150 °C, combustion of organic material over a significant span of time). This process favors the formation of alkylated PAHs; the unsubstituted (or the parent) compounds being relatively low in abundance in these sources (NRCC,1983).

It has been suggested that 70-75% of the carbon in coal is in aromatic form; the 6-membered ring aromatics are dominant with a small 5-membered ring fraction present as well (Neff, 1979). PAHs such as benz[a]anthracene, benzo[a]pyrene, benzo[e]pyrene, dibenzo[c,d,m]pyrene, perylene, and phenanthrene have been identified in coal samples (Woo et al., 1978). Atwater and Mavinic (1985) analyzed wastewater and sludge samples from 11 coal operations across Canada In wastewater, naphthalene and phenanthrene were detected at levels >10 µg/L, whereas anthracene, benzo[k]fluoranthene, and dibenzo[a,h]anthracene levels were <10 µg/L. Naphthalene, phenanthrene, anthracene, fluorene, and pyrene were usually found in sludges at µg/g levels.

The PAH make-up of crude oil and refined petroleum products is highly complex and variable and no two sources have the same composition (Table 4).

Under natural conditions, fossil fuels contribute a relatively small volume of PAHs to the environment. Because most oil deposits are trapped deep beneath layers of rock, there is little chance to emit PAHs to the surface environment. There are some petroleum bodies (e.g., tar sands) which, being near the surface, are capable of contributing PAHs to both atmospheric and aquatic surroundings. These deposits are small in number and are likely to contribute little to the overall volume of PAH in the environment.

4.1.3 Other sources

Volcanic activity and biosynthesis by bacteria and plants are other natural sources of PAHs. Relative to fires, these sources contribute small amounts to the environment. There is still some uncertainty as to whether or not biosynthesis of PAH in vegetation, fungi and bacteria is actually occurring or whether PAH levels in these organisms have been acquired from other sources (Neff, 1979). More sophisticated experimental techniques and equipment are required to resolve these questions.

4.2 Anthropogenic sources

Incomplete combustion of organic matter at high temperature is one of the major anthropogenic source of environmental PAHs. The production of PAHs during pyrolysis (i.e., partial breakdown of complex organic molecules during combustion to lower molecular weight

free radicals) and pyrosynthesis (i.e., combination of free radicals containing one or more carbons) is a function of the temperature. In studying the effects of temperature (550 to 1 000 °C) on pyrosynthesis of PAH from styrene, Commins (1969) found that the yields of all PAHs (ranging in molecular weight from naphthalene to coronene) peaked at 780 °C, decreasing at higher and lower temperatures.

The environmental sources of PAHs of pyrolitic origin are many (Neff, 1979):

(a) Charcoal-broiled steaks, and commercially available smoked food products have been identified to contain PAHs.
(b) Conditions are ideal for PAH pyrosynthesis within a cigarette flame.
(c) Burning of fossil fuels is an important source of PAHs in the environment. Significant quantities of benzo[a]pyrene and other PAHs have been identified in vehicular exhaust.
(d) Many heat and electrical generating facilities burn fossil fuels and produce, as byproducts, liquid, solid, and gaseous wastes that may be rich in PAHs.
(e) Catalytic breakdown of crude petroleum to produce hydrocarbon fuels and other refined products results in the production of PAHs. Many of the PAHs thus produced become concentrated in the high boiling residual oil (e.g., Bunker C and No. 2 fuel oils - Table 4) and asphalt. Significant quantities of PAHs may also be released in flue gas.
(f) The production of coke involves subjecting hard coal to high temperatures (1400 °C) in a reducing atmosphere, conditions ideal for pyrosynthesis of PAHs. Lao et al. (1975) identified 75 PAHs in air-filter samples of gaseous coke oven emissions.
(g) Coal tars, produced by the high temperature treatment of coal, are also known to contain a host of PAHs. These PAHs are derived either from PAHs indigenous to the coal or from pyrolysis of coal hydrocarbons.
(h) Incineration is a valuable means of waste disposal and waste reduction. PAHs in the stack gases, solid residues, and wastewaters from municipal incinerators have been identified (Davies et al., 1976). It has also been found that PAHs released each day in solid residues were 10 times more than in the stack gases and 100 times more than in the wastewater (Davies et al., 1976).

There are many other anthropogenic sources of pyrolytic PAHs. In fact, any industrial or domestic process in which organic carbon is subjected to high temperature will result in the production of some PAHs. Treated wood has also been recognized as a source of PAHs in water and sediments.

In general, anthropogenic sources can be divided into two categories: sources that discharge directly into a body of water, and sources that discharge into the atmosphere.

The sources of PAHs which may discharge directly into aquatic environment include: accidental spillage and/or leakage of PAH-containing fluids (e.g.,waste oils, gasoline, etc.), industrial and domestic wastewaters, urban runoff, discharges originating from landfills, and use of creosoted pilings for docks and other shoreline structures.

Atmospheric PAH emissions fall into two groups: (i) those which originate from stationary sources, and (ii) those which originate from non-stationary sources. Stationary sources include coal and gas-fired boilers; coal gasification and liquifaction plants; carbon black, coal tar pitch and asphalt production; coke-ovens; catalytic cracking towers; petroleum refineries and related activities, electrical generating plants; industrial incinerators; municipal incinerators, agricultural and refuse burning, and any other industry that entails the use of wood, petroleum or coal to generate heat and power. These sources contribute PAHs to the environment either through the formation of these compounds during industrial processing or through pyrolysis of the above mentioned fuels for energy generation. These PAHs, if not degraded in the atmosphere, are sorbed onto particulates in the air and are then deposited onto bodies of water, as well as the surrounding terrestrial environment.

Non-stationary sources of PAHs usually refer to automobiles or other vehicles which use petroleum products as a fuel. Temperatures within an internal combustion engine are often sufficient enough to convert a fraction of the fuel or oil into PAHs via pyrolysis. These compounds are then emitted to the atmosphere through exhaust fumes whereupon they sorb onto particulates. Most PAHs are then photolytically degraded or are deposited onto street surfaces. Precipitation then washes these PAHs into stormwater drainage systems eventually flushing them into the aquatic environment.

4.3 Aquatic environmental loading

According to Eisler (1987) approximately 228 000 metric tons of PAHs are discharged to the aquatic environment per annum as a result of human activity (Table 5). Petroleum spillage and/or leakage of a major and/or a minor nature is the largest contributor to this loading and amounts to 170 000 tons (roughly 75%) of this total. The other major contributor is the atmospheric fallout from the sources listed in section 4.2, which adds an accumulated total of 50 000 tons to aquatic systems. The remaining mass of PAH is contributed through industrial wastewater effluents, sewage effluents and from runoff. The PAH mixtures disposed of in this manner are highly variable and complex due to the large number of sources contributing to this discharge.

4.4 Levels in sediment, water and biota

4.4.1 Water

PAH concentrations in fresh waters vary widely, depending upon such factors as proximity of the waterbody to the source, source type, and season (Moore and Ramamoorthy, 1984).

From a review of data collected in Europe, Neff (1982a, b) noted that drinking water from various sources (e.g., ground water, reservoirs, rainwater, etc.) typically contains 0.2 to 80 ng/L B[a]P and 4 to 4 000 ng/L total PAH. The average concentration of total PAH in drinking water from U.S.A. and Europe, respectively, was quoted to be 15 and 50 ng/L by Lee and Grant (1981). According to Lee and Grant, the concentrations ranging from 50 to 250 ng PAH/L represent the low level contamination of fresh surface water by PAHs, whereas the concentrations ranging from 200 to 1 000 ng PAH/L represent the medium level contamination.

The Great Lakes Science Advisory Board (GLSAB) (1983) has reported concentrations of several PAHs in open waters of the Great Lakes water system (Table 6). In general, the Great Lakes are relatively uncontaminated by PAHs. Although the data for each of the lakes in this system were not available, it is likely that there would be a considerable discrepancy between them as substantial portions of certain lakes (e.g., Lake Ontario, Lake Erie) are more impacted by human activity than others.

^a n=6

PAH concentrations in ambient estuarine and oceanic waters are not well addressed in the literature. The available data are based on estimates of total aromatics by ultraviolet, infrared, and fluorescence techniques, which may be subject to considerable interference from non-PAH materials. The PAH concentrations in marine waters from national and international sources are shown in Table 7. It can be seen that, in each water body, the PAH concentration was a function of the sampling depth with the maximum value recorded near the surface. Marty et al. (1978) also indicated that PAHs (e.g., phenanthrene, alkylphenanthrene, perylene, fluoranthene, and pyrene) in seawater tend to concentrate in the surface microlayer. Several organisms (plankton, fish eggs) are located in this microlayer and may potentially be impacted to a greater extent than those organisms located in sub-surface waters.

Reports on specific PAH concentrations in marine waters are few. Niaussat and Auger (1970) found 1 600 ng/L B[a]P and 3 050 ng/L perylene in the Clipperton Lagoon in the Pacific Ocean. Levels of B[a]P ranging from non-detectable to 400 ng/L were found in the Polynesian atolls of Moruroa and Hao (Niaussat et al., 1975). Gschwend et al. (1982) found that naphthalene concentrations in Vineyard Sound, USA, ranged from < 1.0 to 35 ng/L over sixteen months.

In water 15 m away from an oil separator platform and brine outfall in Trinity Bay, Texas, USA, Armstrong et al. (1977) detected single ring aromatic hydrocarbons (e.g., benzene, toluene, xylene, etc.) as well as naphthalene (0.40 µg/L), 1-methylnaphthalene (0.20 µg/L), 2-methylnaphthalene (0.60 µg/L) and dimethylnaphthalenes (0.70 µg/L). No other LPAH or HPAH were detected in the water although the effluent discharged contained significant quantities of fluorene, phenanthrene and their alkyl derivatives.

Data on PAHs in British Columbia waters are limited. Wan (1991) measured concentrations of 16 PAHs (see Table 12 for the list of PAHs) in the ballasts from five railway rights-of-way and the adjacent ditches (6 locations) flowing to salmon streams in the Lower Mainland of British Columbia. Unlike the ballasts and ditch sediments, PAHs were not consistently found in the ditch water. The average2 concentrations ranging from 0.4 µg/L for acenaphthylene and benzo[a]pyrene to 208 µg/L for fluoranthene were found in the ditch water. Highest concentrations in the water were detected where power and telecommunication line poles were erected in the railway ditches. Among lower molecular weight PAHs, high mean concentrations were found for acenaphthene (8.3 µg/L), anthracene (9.7 µg/L), naphthalene (82.7 µg/L), and phenanthrene (112.9 µg/L). Among high molecular weight PAHs, mean concentrations for benz[a]anthracene (32.3 µg/L), benz[b]fluoranthene (25.5 µg/L), benz[k] fluoranthene (14.0 µg/L), chrysene (76.0 µg/L), fluoranthene (207.7 µg/L), and pyrene (125.8 µg/L) were the highest.

The more recent samples collected by the British Columbia Ministry of Environment, Lands and Parks from Duteau Creek and Christina Lake in the Okanagan area, and Spectacle, Old Wolf, Quamichan, Lizard, and Maxwell Lakes on Vancouver Island indicated that the concentrations of the 16 PAHs (see Table 12 for the list of PAHs) were mostly less than the detection limit (0.01 µg/L); only one sample (Quamichan Lake) recorded a significant number (0.03 µg/L) for naphthalene (Nagpal, 1992). Note that no anthropogenic sources were detected in the vicinity of these creeks and lakes.

4.4.2 Sediment

PAH concentrations reported in this section are expressed on a dry weight (dw) basis in surface sediments at the bottom of water columns, unless indicated otherwise.

PAHs are slightly soluble in water. Binding to particulate matter (especially organic), they tend to accumulate in the bottom sediments. Levels of PAH in sediments vary, depending on the proximity of the sites to areas of human activity. Sediment concentration and distribution of PAHs may also fluctuate due to biodegradation of these chemicals, a process which is reliant upon abiotic and biotic factors which are dependent on site characteristics.

In the surface sediment samples collected from the Great Lakes system, 27 PAHs were identified. Among those commonly found were perylene, pyrene, benzopyrenes, benzoperylenes, fluoranthenes, benzofluoranthenes, and chrysene. The total PAH concentrations in sediments from Lakes Ontario, Erie, and Huron, respectively, were 14 µg/g, 54 µg/g, and 1.2 µg/g (GLSAB, 1983).

PAHs in sediments are elevated near industrial and urban centres. In British Columbia, this trend was evident in the Greater Vancouver area. Dunn and Stich (1975) demonstrated the impact of municipal effluent on sediment PAH concentrations in samples collected near the Iona Island sewage treatment outfall when it discharged onto Sturgeon Bank in shallow water. B[a]P levels of 121 µg/g were detected at a distance of about 0.7 km from the sewage outfall. As this distance increased, however, the concentrations of B[a]P dropped rapidly, registering a value of <1.0 µg B[a]P/g past 5 km.

Recently, Fanning et al. (1989) sampled sediments near the Iona sewage treatment plant outfall which now discharges at depth beyond Sturgeon Bank. They found that the total PAH concentration did not exceed 0.10 µg/g. Levels ranging from 0.166 to 0.177 µg total PAH/g were measured in sediments from the same area by Harding et al. (1988). Sediments sampled from Sturgeon and Roberts Banks were below 0.10 µg total PAH/g (Harding et al., 1988).

Goyette and Boyd (1989a) noted that the sediment PAH concentrations for Vancouver Harbour (Table 8) were considerably higher than those reported for the Fraser River estuary (results from Fanning et al. reported above). The major PAH compounds found in sediments were phenanthrene in the low molecular weight PAH range (i.e., LPAH) and fluoranthene, pyrene, chrysene, benzo[k]fluoranthene, and benzo[b]fluoranthene in the high molecular weight PAH range (i.e., HPAH). Carcinogenic PAHs including benzo[a]pyrene and indeno[1,2,3-cd]pyrene were also present. B[a]P concentration ranged from 0.73 to 1.6 µg/g in the Inner Harbour sediments and 1.9 to 3.0 µg/g in the Port Moody Arm sediments. During a two-year sampling period, the heavily industrialized Port Moody Arm and moderately industrialized Inner Harbour test sites yielded significantly elevated PAH levels compared to the lesser impacted Outer Harbour site. These investigators also concluded that PAH data for sediments in Vancouver Harbour were insufficient to estimate a baseline level.

Sediment samples taken from Estevan Sound, British Columbia, were found to contain 0.0034-0.010 µg/g, 0.0034-0.016 µg/g, and 0.027-0.068 µg/g of LPAH (phenanthrene+ anthracene), HPAH (chrysene+triphenylene+benz[a]anthracene+benzofluoranthenes) and TPAH, respectively (Cretney et al., 1983). This site is located approximately 120 km or greater (linear distance) from the aluminum smelter in the Kitimat Arm and is not subject to any other forms of human impact; consequetly, it may be considered sufficiently uncontaminated to reflect background PAH levels.

# Low molecular weight PAH include naphthalene, acenaphthylene, acenapthene, anthracene, phenanthrene, and fluorene; High molecular weight PAH include fluoranthene, pyrene, chrysene, benzo[a]anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, benzo[ghi]perylene, dibenzo[ah]anthracene, and indeno[1,2,3-cd]pyrene.

Cretney et al. (1983) also sampled and dated sediment cores collected off Emsley Point at the southern extension of north Kitimat Arm, possibly impacted by the aluminum smelter (established in early 1950s). The results showed a dramatic decrease in the total PAH concentration from 2.0 µg/g at 0-3 cm depth to 0.026 µg/g at 75-78 cm depth. It was also evident that a rapid accumulation began sometime between 1944-1959. Prior to this period, the TPAH concentration of 0.031 ± 0.006 µg/g sediment (or 1.8 ± 0.3 µg/g Carbon) was fairly constant for over a century.

Similar trends (i.e, decrease in PAH with increasing depth) were observed by Heit et al. (1981) in sediment core samples collected from Sagamore and Woods lakes in the Adirondack region of New York. The sudden increase in the PAH concentrations of surface sediments was credited to the increase in atmospheric particulates originating from various combustion sources, such as industry, vehicles and heating processes (Heit et al. , 1981; NRC, 1983). The average background PAH concentrations in the sediments are shown in Table 9.

Swain and Walton (1990a, b) measured PAH concentration in sediments collected from several sites in the Fraser River (freshwater sediments) and Boundary Bay (marine sediments), in British Columbia. In most samples PAH levels were below the detection limits (i.e., 0.005 µg/g for acenaphthene, acenaphthylene, anthracene, fluorene, naphthalene, and phenanthrene; 0.02 µg/g for benzo[a] pyrene, benzo[b]fluoranthene, benzo[g,h,i]perylene, benzo[k]fluoranthene dibenzo[a,h] anthracene, and indeno[1,2,3-c,d]pyrene; and 0.01 µg/g for benz[a]anthracene, chrysene, fluoranthene, and pyrene). The maximum PAH concentrations for the freshwater and marine sediments are shown in Table 10.

Wan (1991) measured concentrations of 16 PAHs (see Table 12 for the list of PAHs) in the ballasts from five railway rights-of-way and the adjacent ditches (6 locations) flowing to salmon streams in the Lower Mainland of British Columbia. All 16 PAHs were found in the sediments of all ditches sampled adjacent to the rights-of-way. The mean concentrations for LPAHs varied between 0.25 µg/g for acenaphthylene and 36.7 µg/g for phenanthrene, and for HPAHs between 0.29 µg/g for dibenz[a,h]anthracene and 91.3 µg/g for fluoranthene. The author noted that the PAH mean concentrations in the ditches were several times higher than those in the Inner Harbour sediments of Burrard Inlet (Goyette and Boyd, 1989 a); also, the 16 PAHs were not detected³ in the ditch sediments of pristine parkland and agricultural pump stations.

4.4.3 Biota

PAH concentrations in biota depend upon their proximity to the source of pollution, species ability to biotransform, and bioavailability of soil- or sediment-sorbed aromatic hydrocarbons.

The level of anthracene in seaweed collected from Osaka Harbour, Japan, averaged 4 ng/g dw. The average HPAH concentrations in seaweed for this site ranged from 2 ng/g dw for dibenzo[a,h]anthracene to 72 ng/g dw for benzo[a]pyrene (Obana and Kashimoto, 1981). Harrison et al. (1975) reported 60 ng/g ww TPAH for marine algae from Greenland while Lee and Grant (1981) reported up to 60 ng B[a]P/g dw for marine algae. Seuss (1976) observed that 10-50 ng/g dw of B[a]P was taken up by the freshwater alga, Chlorella vulgaris. Lee and Grant (1981) stated that the worldwide B[a]P concentration for marine plankton ranged up to 400 ng/g dw.

To study baseline levels of B[a]P, Dunn and Young (1976) collected mussels (Mytilus californianus, and Mytilus edulis) from 19 mainland and 6 island stations situated throughout the Southern California Bight. The coastal area in this region is inhabited by about 5% of the U.S.A. population. At both mainland and island stations, levels of contamination in mussels taken from locations at least 1 km from piers and wharfs were generally at or near the detection limit of 0.1 ng/g ww. The samples which recorded elevated levels of B[a]P were those in which the mussels were growing directly on creosoted pilings (e.g., up to 8.2 ng/g ww), or were growing near large harbours or marinas (e.g., up to 2.3 ng/g ww). The data collected in Oregon from a relatively pristine Alsea bay site showed non-detectable levels of B[a]P (< 0.4 ng/g dw or < 0.10 ng/d ww) in the tissues of gaper clams (Tresus capax ), blue mussels (Mytilus edulis ) and softshell clams (Mya arenaria) (Mix et al., 1977).

In British Columbia, PAHs in shellfish were first reported by Dunn and Stich (1975). Levels up to 0.2 ng B[a]P/g ww were measured in mussels (M. californianus ) from the open west coast of Vancouver Island, 5 km from human activity; 42.8 ng/g ww B[a]P were found in mussels from a poorly flushed inlet (False Creek) with heavy boat and industrial use. At four out of five sites in the Vancouver Harbour, B[a]P uptake by mussels fluctuated seasonally. These seasonal fluctuations were attributed to variations in pollution pattern rather than physical differences such as temperature, or physiological differences related to the breeding cycle of the organisms.

Duncan (1984) monitored PAHs in commercial shellfish from seven British Columbia locations, with pacific oysters (Crassostrea gigas ) collected at five of the stations (e.g., Henry Bay, Denman Island, Comox Harbour, Cortes Island, and Barkley Sound), butter clams (Saxidomas giganteus ) from the sixth (i.e., Seal Islets), and geoducks (Panopea generosa ) from the seventh location (i.e., Courtenay area). The low to moderate levels of PAH (and metals, which are not shown here) led the investigator to conclude that the major shellfish harvesting sectors of B.C. were located in areas of good water quality (Table 11).

Crustaceans possess an MFO system which is capable of converting most PAHs into water- soluble metabolites. Most of these compounds and their resulting products are distributed in the hepatopancreas (the main site of MFO), although residual levels are often detectable in other organs and tissues (Dunn and Stich, 1975). For example, American lobsters (Homerus americanus) held in creosoted tidal ponds in Nova Scotia were found to have 35 times as much PAHs in the hepatopancreas than in the tail muscle (Uthe et al., 1984). This study found that hepatopancreatic levels of these substances tended to be greater than tail muscle concentrations regardless of whether the lobster was held in a contaminated pond or was freshly obtained from a relatively pristine site. In a similar study of PAH uptake in H. americanus from a minor diesel oil spill in Arnold's Cove, Newfoundland, Williams et al., 1985 noted that PAHs preferentially concentrate in the hepatopancreas of the animals.

Goyette and Boyd (1989a) analyzed Dungeness crab (Cancer magister ) tissues (muscle and hepatopancreas) for PAHs. The results of Goyette and Boyd are reproduced in Table 12. The animals were caught from False Creek, upper Indian Arm, Coal Harbour, and Port Moody Arm (Ioco) off Vancouver Harbour. PAH concentrations in both muscle and hepatopancreas tissues were non-detectable in samples from upper Indian Arm (detection limit = 0.02 µg/g dw except for indeno(1,2,3-cd)pyrene, dibenzo(a,h)anthracene, and benzo(g,h,i)perylene which was 0.06 µg/g dw). Crab samples from False Creek had the highest concentrations in both muscle (0.025 - 0.169 µg total PAH/g dw) and hepatopancreas (0.148 - 1.24 µg total PAH/g dw) tissues. Similar to the Dunn and Stich (1975) observations above, hepatopancreas tissue was the primary site for PAH accumulation.

MFOs in fish have been extensively studied and are believed to be very similar to those of mammals. These enzymes are located in microsomal tissues present in the livers of these animals and serve as a detoxifier of toxic substances. Enzymatic activity of MFOs in fish is more effective in metabolizing PAHs than it is in lower animals (e.g., invertebrates). As a result, these vertebrates tend to bioaccumulate few PAHs (Lawrence and Weber, 1984). West et al., (1984) observed that higher molecular weight PAHs, which include the largest class of chemical carcinogens, do not accumulate in fish. Muscles of six species of fish in Lake Ontario were found to contain 3-8 µg/kg ww of TPAH (Eisler, 1987), while trout tissues from Lake Maskinonge, Ontario, did not exceed individual PAH concentrations of 1.5 µg/kg ww (Pancirov and Brown, 1977). Carp (Cyprinus carpio , a herbivore) from Hamilton Harbour and Detroit River contained 0.003-0.243 µg/kg ww of PAH (Perylene, Benzo[k]fluoranthene, Benzo[a]pyrene, and Coronene), while Northern Pike (Esox lucius , a carnivore) from the same

*Levels below the detection limit are semi-quantitative estimates of PAHs.
+Hepatopancreas

sources contained 0.016-0.074 µg/kg of the same (GLSAB, 1983). English Sole (Parophrys vetulus ) sampled from two urban bays and two non-urban bays in Puget Sound, Washington were all found to contain <0.05 µg/kg dw of total aromatic hydrocarbons in muscle tissues.

Goyette and Boyd (1989b) examined PAH levels in the liver and muscle tissues of English Sole from Vancouver Harbour. Concentrations ranging from 0.001-0.037 µg/g dw of LPAH and trace-0.074 µg/g of HPAH were detected in the fish livers from the outer Harbour. The inner Harbour fish liver samples contained 0.013 µg/g fluoranthene, 0.001 µg/g anthracene and 0.014 µg/g phenanthrene. In Port Moody Arm, only phenanthrene was detected at 0.019 µg/g dw. By comparison, the muscle samples contained non-detectable levels of both LPAH and HPAH except for phenanthrene and fluoranthene which were present in trace amounts and 0.013 µg/g dw, respectively.

In several fish species collected from the North and Main Arms of the Fraser River, British Columbia, Swain and Walton (1989) found PAHs in both muscle and liver tissues. The PAH concentrations were much greater in the liver than in the muscle samples (Table 13).

2 (Based on 6 measurements at 6 locations, including those containing non-detectable levels; the non-detectable levels were considered to have a zero value. The dection limits for LPAHs and most of HPAHs were 0.1 µg/L and 0.5 µg/L, respectively; the detection limit for benz[g,h,i]perylene, dibenz[a,h]anthracene, and indeno[1,2,3-cd]pyrene was 0.8 µg/L.

3 the detection limits for LPAHs, and chrysene = 0.01 µg/g; benz[a]anthracene, fluoranthene, and pyrene = 0.05 µg/g; benz[b]fluoranthene, benz[k]fluoranthene = 0.08 µg/g; benz[a]pyrene, benz[g,h,i]perylenedibenz[a,h]anthracene, and indeno[1,2,3-cd]pyrene = 0.10 µg/g.