6.0 AQUATIC LIFE

6.1 Freshwater

6.1.1 Lethal and acute effects

The acute (96-h LC₅₀) and lethal effects of PAHs in the freshwater environment are shown in Table 16. Figure 4 shows a graphic summary of the data.

The toxicity of PAHs to aquatic organisms is determined by several factors which include: (a) the PAH type (e.g., molecular weight, alkyl substitution, etc.), (b) the species of the organism exposed, and (c) the duration and the type of exposure to a given PAH (Table 16). In general, fish appear to be the most sensitive of the aquatic organisms to PAHs (Figure 4). However, there are exceptions to this general trend. For instance, the 96-h LC₅₀ for acenaphthene (ANA) was lower for the alga Selenastrum capricornutum (EC₅₀ for cell count = 520 µg/L) than for brown trout (Salmo trutta ) (LC₅₀ = 580 µg/L) or fathead minnow (Pimephales promelas ) (LC₅₀ = 610 µg/L). The longer exposure periods reduce the LC₅₀s for both cladoceran (Daphnia magna) and bluegill (L. macrochirus) exposed to acenaphthene and fluoranthene (FLAN) (LeBlanc, 1980; Buccafusco et al., 1981).

The minimum LC₅₀ for relatively more soluble and lower molecular weight PAHs, containing 3 or less aromatic (benzene) rings in their structure (Table 1), was found for rainbow trout (Oncorhynchus mykiss ) exposed to phenanthrene (LC₅₀ = 30 µg/L, Table 16).

The higher molecular weight PAHs (containing more than 3 aromatic rings in their structure) such as benz[a]anthracene and and benzo[a]pyrene, have also been shown to be acutely toxic to invertebrates at low concentrations (5-10 µg/L - Table 16). In natural aquatic environments, this condition may not be achieved because of the low solubility of HPAHs. Note that the 96-h LC₅₀ (5 µg/L) for Daphnia pulex exposed to B[a]P was higher than its solubility (3.8 µg/L) in water. The solubility of B[a]ANTH (Table 1) and the 48-h LC₅₀ for Daphnia pulex exposed to B[a]ANTH were nearly identical.

Alkyl homologues of PAHs are generally more toxic to aquatic life than the parent compound. For instance, the 48-h EC₅₀ for Daphnia pulex exposed to anthracene (750 µg/L) was much higher than that obtained when the organisms were exposed to methyl anthracene (48-h EC₅₀=96 µg/L) or 9-methoxy anthracene (48-h EC₅₀=400 µg/L) (Table 16).

1 S= static; FT= flow through; M= measured; U= unmeasured

Figure 4

6.1.2 Sublethal and chronic effects

Sublethal and chronic effects of PAHs on growth and physiological processes of aquatic algae and plants are shown in Table 17. The data suffer from some major drawbacks: (a) short exposure periods, (b) exposure levels higher than PAH aqueous solubilities, and (c) lack of constancy in the PAH concentration during the experiments. Bastian and Toetz (1982) exposed Anabaena flos-aquae in open flasks to several PAHs for 14 days. Within 7 days, acenaphthene, fluorene, naphthalene, and pyrene completely disappeared from solution, whereas the benzanthracene, phenanthrene, chrysene and fluoranthene concentrations were reduced to 85%, 77%, 62%, and 49%, respectively, of the initial value at the end of the 14-d period. Additionally, the concentrations of several PAHs used in the experimental solution were greater than their individual aqueous solubilities (Table 1). In the study on nitrogen fixation by Anabaena flos-aquae, Bastian and Toetz (1985) used short-term exposure (2 h) to minimize losses of PAHs during the experiment. Several PAHs were observed to reduce the nitrogen fixation by the alga, but the long-term effects of the PAHs are difficult to predict from these short-term studies (Table 17).

The data on long-term or chronic effects of PAHs on freshwater animals are few and suffer from the same drawbacks as noted above for aquatic plants (Table 18). Brown et al. (1975) exposed bluegill (Lepomis macrochirus ) to 1 000 µg/L benz[a]anthracene and found 87% mortality in 6 months. The B[a]ANTH concentration used by these investigators in their study was much higher than the aqueous solubility of the PAH (Table 1). Finger et al. (1985) reported 12% mortality in bluefish exposed to 500 µg/L fluorene for 30 days.

Teratogenic effects during organogenesis (7- to 24-d post fertilization) were studied by Hannah et al. (1982) and Hose et al. (1984) in rainbow trout (Oncorhynchus mykiss ) exposed to B[a]P-contaminated sand (1-500 µg/g) (Table 18). Gross anomalies (e.g., microphthalmia) were noted in a significant population of fish (6.8%) exposed to the contaminated sand; the average aqueous concentration was 0.2 µg B[a]P/L (Hose et al., 1984).

The minimum concentrations of lower molecular weight PAHs: naphthalene, acridine, and phenanthrene, causing gross developmental anomalies in rainbow trout, were found to be much higher (than B[a]P) at 230, 410, and 85 µg/L, respectively (Black et al., 1983).

6.1.3 Photo-induced effects

Several PAHs, accumulated by aquatic organisms during exposure, have been shown to be severely toxic when the contaminated organisms were exposed to sunlight or ultraviolet radiation (Table 19). For instance, Bowling et al. (1983) found that 12.7 µg/L anthracene was fatal to bluegill sunfish (Lepomis macrochirus ) in 48 hours in an outdoor channel in bright sunlight. No mortality was noted in fish exposed to the PAH in the shaded area of the channel. But, when shading was removed after day 4 (when anthracene concentration in water had dropped to zero and fish were allowed to depurate for 24 hours), all fish previously in the shaded area died within 24 hours. It was concluded that direct sunlight exposure of anthracene-contaminated fish, and not the toxic anthracene photoproducts in the water, was responsible for the mortality of the bluegill.

Photo-induced toxicity of PAHs due to ultraviolet (UV) radiation was subsequently studied by other investigators (Oris and Giesy, 1985, 1987; Newsted and Giesy, 1987). For the purpose of photochemical considerations, UV light is divided into three bands of varying wavelengths: UV-A (390-315 nm), UV-B (315-285 nm), UV-C (285 nm and lower). Although much of the incident solar ultraviolet radiation (SUVR) is filtered out by the atmosphere, some SUVR of longer wavelengths (290-400 nm) passes through. Several PAHs have shown absorption maxima in the 290-400 nm wavelength range. Previous to the discovery of the photo-induced toxicity reactions in contaminated aquatic organisms, most of the laboratory studies with PAHs were conducted in conventional cool fluorescent lighting to avoid photo-oxidation of the compounds; PAHs absorb little radiation in the visible band.

Bearing in mind that LC₅₀ increases as the period of exposure decreases, a comparison of the data in Tables 16 and 19 suggests that photoactivation of PAHs was responsible for the observed increase in acute toxicity seen in invertebrates and fish. Also, higher molecular weight PAHs (e.g., B[a]P), which previously were not considered to be acutely toxic to fish because of their low aqueous solubility, could cause an acute toxic reaction if photoactivation occurred.

The phototoxicity of a PAH is a function of several factors: (a) PAH concentration in tissue, (b) length of exposure to and absorption of SUVR by the organism, (c) the efficiency of conversion of ground-state molecules to the excited triplet state, and (d) the probability of the excited intermediate reacting with a target molecule (Newsted and Giesy, 1987).

Newsted and Giesy (1987) investigated several descriptors (e.g., lowest energy triplet state, phosphorescence lifetime, octanol/water partition coefficient, etc.) to describe photo-induced toxicity of PAHs to Daphnia magna. The phototoxicity of PAHs was highly correlated with the lowest energy triplet state⁴ (LETS, kJ/mol). It was also found that compounds with long phosphorescence lifetimes5 (PLT > 3.5 s) were not toxic. Newsted and Giesy used these relationships to classify PAHs into toxic categories of very toxic (LT₅₀6 < 900 min), moderately toxic (900 < LT₅₀ < 9999 min), and non-toxic. Working with the larvae of fathead minnow (Pimephales promelas), Oris and Giesy (1987) used a similar technique to separate toxic (i.e., causing phototoxic reaction within 96 h) and non-toxic (non-phototoxic in 96 h) PAHs. The results were similar in most of the cases (Table 20). Although promising, these modelling techniques need further improvement to become useful as a tool in the environmental regulation of numerous PAHs which have not been tested yet.

While modelling photo-induced acute (96-h LC₅₀) and chronic toxic effects in bluegill sunfish (Lepomis macrochirus) exposed to anthracene, Oris and Giesy (1986) found that both acute and chronic toxicities were dependent on the length of exposure to solar ultraviolet radiation. As the daily exposure to SUVR was increased, the threshold concentration for predicted effects decreased. Furthermore, the absolute difference between acute and chronic threshold concentrations decreased as the daily exposure to SUVR increased (e.g., at daily exposure of 5 h, the predicted acute concentration = 55 µg/L and chronic concentration = 17 µg/L; and at daily exposure of 20 h, the acute concentration = 7.8 µg/L and chronic concentration = 2.2 µg/L).

The evidence for the photo-induced toxicity of PAHs is recent and its significance in natural aquatic systems is yet to be understood. Aquatic organisms in deep and turbid waters and shaded areas may not be affected by this phenomenon. However, juveniles of most fish are found in the shallow areas of the littoral zone and are subject to photo-induced toxicity of PAHs. Landrum et al. (1986) calculated that SUVR of sufficient intensity, which may cause 50% immobilization in the cladoceran (Daphnia pulex) exposed to 1.2 µg/L anthracene over a 14-h

+ Results based on bioassays; ++ Predicted results from toxicity modelling techniques;
*Very toxic = LT₅₀ < 900 min, and Moderately toxic = 900 < LT₅₀ <9999 min;
**Toxic = LT₅₀ < 96 h, and Non-toxic = LT₅₀ > 96 h

daylight cycle, could penetrate to a depth of 7 metres in a lake. Since photo-induced toxicity of PAHs is a function of several factors which can not be duplicated in a laboratory environment, sophisticated experimentation and evaluation techniques are needed to determine the extent to which phototoxicity may proceed in natural aquatic environments.

6.2 Marine water

6.2.1 Lethal and acute effects

Most of the literature on acute and lethal toxic effects (as EC₅₀ for aquatic plants and LC₅₀ for aquatic animals) in estuarine and marine environments is related to the lower molecular weight PAHs (LPAH), containing 3 or less benzene rings in their structure (Table 21). These compounds are relatively more soluble in water than the higher molecular weight PAHs (HPAH); at saturation, their concentrations in water (Table 1) can exceed the LC₅₀s shown in Table 21.

The PAH concentrations causing lethal effects in marine organisms vary widely. The lowest 96-h LC₅₀ of 40 µg/L was recorded for juvenile mysid shrimp (Mysidopsis bahia) exposed to fluoranthene (USEPA, 1978). Since this test with shrimp was conducted in a static (as opposed to a flow-through) environment and the PAH concentration was not measured during the experiment, the results of this test were considered to be of secondary⁷ importance.

Several trends were established from Table 21: (a) The toxic (LC₅₀) PAH concentration for an organism decreased with longer exposure periods; for instance, the 96-h LC₅₀s (1 900 µg/L for 1-MNA, and 1 300 µg/L for 2-MNA) for Dungeness crab (Cancer magister ) exposed to two methylnaphthalenes were about four times lower than their 48-h LC₅₀s (8 200 µg/L for 1-MNA, and 5 000 µg/L for 2-MNA) (Caldwell et al., 1977); (b) the degree and position of methylation affected PAH toxicity; e.g., dimethylnaphthalene (96-h LC₅₀ = 600 d-MNA/L) was more toxic to Dungeness crab than a methylnaphthalene (96-h LC₅₀ = 1 300 2-MNA/L),

1 S= static; U= unmeasured; FT= flow through; M= measured

whereas, between the two methylnaphthalenes, 2-MNA was more toxic than 1-MNA (96-h LC₅₀ = 1 900 1-MNA/L) (Caldwell et al., 1977). With the addition of each methyl group, Ott et al. (1978) found that the 24-h LC₅₀ for the copepod Eurytemora affinis exposed to naphthalene was reduced by approximately one-half; (c) the acute and lethal reactions of HPAH (e.g., benzo[a]pyrene, chrysene, and dibenz[a,h]anthracene) were limited to concentrations much above their solubilities in seawater.

6.2.2 Sublethal and chronic effects

Chronic toxicities of various PAHs are shown in Table 22. The lack of data and a wide variety of end points chosen by the investigators during experimentation prevent a comparison of chronic toxicity levels among PAHs.

The most sensitive chronic effects of naphthalene were observed by DiMichele and Taylor (1978) while studying histopathological and physiological responses in mummichog (Fundulus heteroclitus ). These investigators found gill hyperplasia in 80% of the fish after a 15-d exposure to 2 µg/L naphthalene; only 30% of the control fish showed the effect. All of the fish exposed to 20 µg NA/L demonstrated necrosis of tastebuds, a change not observed in the control fish.

Ott et al. (1978) found that lethal toxicity (24-h LC₅₀) of naphthalene and its alkylated derivatives was determined by the degree of methylation (see Section 6.2.1). This trend was not obvious from low level chronic studies due to the lack of appropriate data (e.g., organism tested, experimental end points, experimental test conditions, etc. were different in different studies). Ott et al. found that chronic exposure of copepod E. affinis to various naphthalenes (2-MNA, d-MNA, and t-MNA) at a concentration of about 10 µg/L in sea water for the duration of their adult life (maximum 29 days) resulted in significant reductions in the organisms' length of life, total numbers of nauplii produced, and mean brood size. Exposure to all naphthalenes at 10 µg/L resulted in reduced rates of egg production which were, on average, about 50% of those of control animals.

Miller et al. (1982) found that the concentration of 1 µg/L chrysene in water increases incidence of molts in pink shrimp (P. duorarum ) after 28 days.

Among PAHs studied, B[a]P was found to be the most toxic (Table 22). Five percent of sand sole (Psettichthys melanostichus ) eggs exposed to 0.1 µg B[a]P/L in water (as compared to 0% in control fish) showed gross anomalies such as overgrowth of tissue originating from the somatic musculature, and arrested development (Hose et al., 1982). Also, the hatching success of eggs exposed to 0.1 µg B[a]P/L (average = 28.1%; range = 7.0% to 67.6%) was significantly lower than that of controls (average = 57%; range = 21.6% to 89.6%).

6.3 Mutagenicity, carcinogenicity, and tumor induction

Several PAHs, especially those containing 4 to 6 aromatic rings in their structure, have been shown to be mutagenic, carcinogenic, and inducers of tumors in mammals exposed to high doses of the contaminants in the laboratory. Studies directly linking PAHs to these effects in fish are not only few in number but also have used excessively high exposure levels of PAHs.

Hendricks et al. (1985) exposed rainbow trout (Salmo gairdneri now classified as Oncorhyncus mykiss ) for 6 to 18 months to a diet containing 1000 µg B[a]P/g dry weight. Hepatic neoplasms were observed in 25% of the trout that were fed the B[a]P diet for 18 months; the fish on the control diet did not show the effect at all. Of the affected (i.e., 25%) population, 21% had livers with at least one hepatocellular carcinoma. In the same experiment, Hendricks and coworkers also found that at the end of 18 months the body weight was lower in the B[a]P-fed fish than the control fish; however, the number of mortalities was higher in the control population (5% as opposed to 3% in treated fish) which suggested that B[a]P was not acutely toxic. Similar results were obtained by Hendricks et al. (1985) in 10-month old trout (weighing 45-55 g) intraperitoneally injected with B[a]P (dissolved in 0.4 mL propylene glycol) at the rate of 1 mg B[a]P/month for 12 months.

Krahn et al. (1986) examined hepatic lesions in English sole (Parophrys vetulus ) caught from 10 sites in Puget Sound, Washington. A strong positive correlation was found between B[a]P metabolite equivalents (67-2100 µg/kg wet weight) in bile and the prevalence of hepatic lesions (6.2-90%). However, the concentration of metabolites in the bile was highly variable within locations and was not correlated with the PAH (e.g., naphthalene, phenanthrene, and B[a]P) concentration in the sediment. In the same area, Malins et al. (1985a, 1985b) found elevated hepatic neoplasms in association with sediments which were heavily contaminated with creosote and also contained PAHs (e.g., acridine and carbazole).

Repeated short-term (five successive 6-h periods separated by weekly intervals) aqueous exposure of viviparous8 fish (Poeciliopsis lucida and P. monacha ) to 5 000 µg/L 7,12-dimethylbenzo[a]anthracene (DMB[a]AN) induced hepatocellular neoplasms in 25 of 60 fish in 6 to 7 months; tumors were not found in any of the controls (0/59) (Shultz and Shultz, 1982). However, short-term exposures to 250 µg/L 7,12-DMB[a]AN failed to induce neoplasms. Since these results were obtained using concentrations much above the aqueous solubility of the PAH (1.5 µg/L at 25 C - Table 1), they cannot be considered indicative of a natural exposure situation.

In a preliminary work by Hawkins et al. (1988), young guppies (Poecilia latipes ) and medakas (Oryzias latipes ) were exposed (two 6-h exposures one week apart) to B[a]P and DMB[a]AN concentrations of <5 µg/L (0.45 µm filtrate with no carrier), 30-50 µg/L (carrier-mediated 0.45 µm filtrate), and 150-250 µg/L (carrier-mediated glass fiber filtrate). Preliminary analyses indicated that hepatic neoplasm occurred in both fish exposed to B[a]P and DMB[a]AN. In the guppy, B[a]P-induced hepatic neoplasms (10% at 24 weeks post initial exposure) were limited to the high exposure group. DMB[a]AN was much more carcinogenic to fish than B[a]P. Medakas, which were affected by all concentrations of DMB[a]AN, also showed numerous non-hepatic lesions at the high DMB[a]AN concentration. The prominence of both hepatic and non-hepatic lesions at high exposure concentrations (which contained particulates) suggested that the insoluble or particulate fraction of PAHs may play an important role in carcinogenesis. Both guppies and medakas are not native to British Columbia; also, PAH concentrations used in these tests are much higher than their aqueous solubilities (Table 1).

Metcalfe and Sonstegard (1984) exposed rainbow trout (O. mykiss) embryos at the eyed stage to DMB[a]AN using a micro-injection (0.5 µL) technique. Injections of 0.25 DMB[a]AN/embryo produced grossly visible neoplasms and carcinomas at a frequency ranging from 3.5 to 6.3%. Using a similar technique, Black et al. (1985) noted that injections of 10 µg B[a]P/embryo (136.1 µg B[a]P/g) produced a tumor incidence of 8.7% after nine months in rainbow trout, while all control fish appeared healthy.

6.4 Mixed-function oxidases

Mixed-function oxidases (MFOs) are a group of enzymes which are located in hepatic microsomes and play an important role in the biotransformation of PAHs (and other xenobiotics e.g., PCBs, dioxins, and furans) before their excretion from the body. The action of MFOs can either detoxify PAHs or produce more reactive intermediates which are cytotoxic and possibly carcinogenic.

There are several enzyme systems which make up the mixed-function oxidases. They include, for instance, aryl hydrocarbon hydroxylase (AHH), ethoxycoumarin O-deethylation (ECOD), ethoxyresorufin O-deethylation (EROD), etc. Several fish and invertebrate species have been reported to show a mixed-function oxidase activity in response to a PAH exposure.

Gerhart and Carlson (1978) exposed rainbow trout to several PAHs (e.g., 1,2,4-trimethylnaphthalene, phenanthrene, pyrene, fluoranthene, chrysene, and benzo[a]pyrene) in water and through intraperitoneal (ip) injections to screen for MFO activity. Chrysene and B[a]P were found to induce MFO activity in the fish injected with the PAHs. In water, bioaccumulation of B[a]P (average aqueous conc.=0.4 µg/L) resulted in MFO induction in 10 days, whereas bioaccumulation of pyrene (3.9 µg/L) and fluoranthene (3.31 µg/L) did not. Based on these experiments, it was predicted that tissue (excluding liver and gut) concentrations in excess of 0.3 µg/g B[a]P would cause MFO induction in rainbow trout. These investigators did not correlate MFO induction with other effects (e.g., carcinogenic, mutagenic, tumor induction, etc.), but a comparison of these results with those of Black et al. (1985) in section 6.3 suggests that the tissue concentration of 0.3 µg B[a]P/g may induce MFO activity in rainbow trout without producing a tumor.

Walters et al. (1979) found naphthalene (200 µg/L), 2-methylnaphthalene (100 µg/L), 2,6-dimethylnaphthalene (100 µg/L), 3-methylcholanthrene (100 µg/L), and benzo[a]pyrene (50 µg/L) increased AHH activity in the marine zooplankton Calanus helgolandicus by 191%, 125%, 29%, 146%, and 129%, respectively, over the control population. However, in all cases, the observed mortality in the animals exposed to the PAHs for 7 days was not significantly different from those of the control population.

Recently, Hendricks et al. (1985) studied the hepatocarcinogenicity of B[a]P to rainbow trout by dietary exposure (500 and 1 000 µg/g dry weight of diet) and intraperitoneal injection (10-month old fish weighing 45-55 g injected with 1.0 mg B[a]P once a month for 12 months). In 9 weeks, the MFO activity (e.g., EROD) increased significantly in the fish fed both 500 and 1 000 µg/g B[a]P diets over the control population. In the long-term feeding experiments (up to 18 months) with a diet containing 1 000 µg/g B[a]P, these investigators found that the treated population had significantly lower body weight than the control (mortality was similar in both populations). No lesions were found in the fish for the first 6 months, but the incidence of liver neoplasms (basophilic foci+carcinomas) increased to 15% (12%+3%) and 25% (4%+21%) after 12 and 18 months, respectively of feeding B[a]P in the diet. The whole fish B[a]P level at the end of 18 months was estimated to range from 860 to 1 000 µg/g wet weight.

In the intraperitoneal injection experiments, Hendricks et al. (1985) found that the incidence of the carcinoma of the liver in rainbow trout increased to 46% (as opposed to 4% in control; based on fish that survived) in 6 months after 12 monthly ip injections of B[a]P. At the end of the experiment, the B[a]P concentration in whole fish, based on the starting weight of the fish at 55 g, was about 218 µg/g. The investigators did not measure the MFO activity in this portion of the study, but mortality of the fish was high (44% in treated fish versus 46% in control), which was attributed to propylene glycol used as the PAH carrier.

6.5 Mixtures of contaminants

In aquatic environments, organisms are exposed to several contaminants at a time. To assess the impact of a PAH on aquatic environments, therefore, interactions between contaminants must be considered. Landrum (1983) found a 50% reduction in the uptake of B[a]P and anthracene by the amphipod Pontoporeia hoyi in presence of toluene (expressed as I₅₀9 values). It was noted that I₅₀ was a function of several factors, including (a) co-contaminant solubility- I₅₀ increases as the co-contaminant solubility increases, and (b) aqueous solubility of the primary contaminant- the reduction in the uptake of relatively insoluble and hydrophobic B[a]P (I₅₀ = 740 µg/L toluene) was more sensitive than anthracene (I₅₀ = 2 200 µg/L toluene).

Stein et al. (1984, 1987) exposed English sole (Parophrys vetulus ) to sediments labelled with 3H-benzo[a]pyrene and ¹⁴C-Aroclor 1254 (a PCB formulation) either singly or together, and found that the accumulation of B[a]P-derived radioactivity was enhanced in the fish exposed to both contaminants simultaneously relative to exposure to the PAH alone. The formation and accumulation of potentially toxic metabolites of carcinogenic B[a]P in sole liver were also increased by the simultaneous exposure to other contaminants present in a sediment. However, while investigating accumulation of naphthalene, a PCB mixture, and B[a]P by the oyster Crassostrea virginica., Fortner and Sick (1985) found several instances where multiple components had antagonistic effects on PAH accumulation.

The literature on interactions between co-contaminants is limited. More work is needed on modelling and quantifying these interactions before they can be applied to real aquatic environmental situations.

6.6 Other modifying factors

Smith et al. (as quoted in OMOE, 1990) found that lethal toxicity of waterborne phenanthrene to Daphnia pulex increased over a narrow temperature range of 17 °C to 20 °C, but only a marginal effect was noted with 1,3 dimethylnaphthalene. On the other hand, a decrease in temperature increased bioaccumulation of PAHs. This increase in bioaccumulation may be attributed to several factors including decreased rates of depuration and metabolism with decreasing temperature. Higher PAH concentrations were observed in fish and invertebrates at lower temperatures, even though uptake kinetics may be slower (Varanasi et al., 1981). In general, the effect of temperature on PAH toxicity and bioaccumulation is minor (Varanasi et al., 1981).

Complexation by dissolved organic matter (DOM) reduces the bioavailability of PAHs (Leversee et al., 1983; Spacie et al., 1983; McCarthy and Jimenez, 1985b; McCarthy et al., 1985). Leversee et al. (1983) found that the presence of humic acids reduced bioaccumulation of unsubstituted PAHs in proportion to their Kow (B[a]P > anthracene > naphthalene). A reversal in the bioaccumulation trend with humic acid (i.e., higher bioaccumulation at higher humic acid content) was observed for methylcholanthrene, but no explanation was offered.

Humic acids were also shown to increase salting-out of PAHs initiated by increasing salinity. Spacie et al. (1983) found that the addition of dissolved humic acids to water decreased B[a]P accumulation in bluegill sunfish (Lepomis macrochirus ); however, in the same experiment, the accumulation of anthracene was not affected by the added humic acids.

McCarthy and Jimenez (1985b) and McCarthy et al. (1985) investigated the binding and dissociation of several PAHs (B[a]P, benzanthracene, and anthracene) with dissolved humic material (DHM) and PAH uptake and accumulation by bluegill sunfish (L. macrochirus) and Daphnia magna. A positive logarithmic relationship between Kow and an association constant with dissolved humic material (i.e., Kp) was found. Also, the binding of PAHs with humic acids was completely reversible. The presence of humic acids dramatically reduced the availability of PAHs for uptake by organisms, which lead these investigators to suggest that dissolved organic material has the potential to mitigate effects of PAH in aquatic systems.

Landrum et al. (1987) found that the partitioning (expressed as partitioning coefficient Kb) of PAHs binding to dissolved organic carbon was not adequately predicted by Kow; only 46% of the variance in Kb was explained by Kow. Thus, while Kow was an adequate descriptor for PAH binding with dissolved humic material (DHM) of consistent composition, it clearly lacked the capacity for predicting partitioning of PAHs on the basis of dissolved organic carbon (or Kb). Obviously, the composition and complexing properties of the dissolved organic carbon used in the Landrum et al. (1987) study were different than those of the standardized DHM used in the McCarthy and coworkers studies. Landrum et al. (1987) noted that Kb was more closely related with a sorption coefficient derived through reverse phase separation.

PAHs are also complexed by sediments. However, sediment-bound PAHs may become available to organisms at higher trophic levels through ingestion of benthic organisms living in these sediments. For instance, Landrum and Scavia (1983) found that sediment-associated uptake of anthracene by the amphipod Hyalella azteca was slower than aqueous uptake; however, the sediment-associated anthracene accounted for 77% of the steady state body burden in these organisms.

6.7 Bioaccumulation

Aquatic organisms can accumulate PAHs from water, sediment, and food. The literature suggests that PAH uptake by aquatic organisms depends upon several factors: (a) physical and chemical properties of the PAH (e.g., molecular weight, octanol/water partition coefficient, etc.), (b) environmental variables (e.g., suspended matter, dissolved organic matter, bioavailability, temperature, presence of other contaminants, biodegradation, etc.), and (c) biological factors (e.g., PAH metabolism and depuration rates, feeding characteristics of organisms, fat content of tissue, lifestage, etc.) (McElroy et al., 1989).

The octanol/water coefficient (Kow) has been shown to be a good descriptor for accumulation for several PAHs in aquatic organisms (expressed as bioconcentration factor BCF, which is equal to the contaminant concentration in the organism ÷ the contaminant concentration in the aqueous phase) (Southworth et al., 1978; Pruell et al., 1986). The relationship between Kow and BCF may be modified by the PAH affinity for dissolved organic and solid phase fractions (McCarthy et al., 1985). In some cases, the relative availability of PAHs does not appear to be a simple function of their physical and chemical properties. For instance, Varanasi et al. (1985) found 4-ring PAHs to be more available than either 3- or 5-ring PAHs from contaminated sediments in Puget Sound, Washington, for accumulation by two species of amphipod (Rhepoxynius abronius and Pandalus platyceros) and one species of clam (Macoma nasuta).

The accumulation of petroleum hydrocarbons by the two populations of the oyster, Crassostrea virginica, was positively correlated with their fat content (Stegeman and Teal, 1973). Bioconcentration of PAHs in the fat tissue, however, may be influenced by the lipid composition. Schneider (1982) found that the difference in PCB accumulation in different organs of cod was eliminated when residues were normalized to the neutral lipid or fat (mostly triglycerides) content (rather than extractable lipid content) of the organ. No such data are available for PAHs.

In general, waterborne PAHs are taken up relatively rapidly as compared to sediment-associated PAHs. McElroy (1985) exposed the polychaete, Nereis virens, to ¹⁴C-benz[a]anthracene introduced directly to the water column or already sorbed to the sediment. The tissue/sediment ratio for N. virens and the degree to which accumulated residues were metabolized in a 7-d period, were significantly higher when B[a]ANTH was introduced via the water column. Regarding short-necked clams, Tapes japonica, exposed to PAHs added to the water column in a circulating aquarium, or to PAHs in contaminated sediments collected from an urban harbour, Obana et al. (1983) observed that the clams reached apparent equilibrium with water concentrations in one day in the first case, while concentrations of most individual PAHs were still increasing in clams exposed to contaminated sediment for 7 days. Investigating the influence of sediment on uptake of PAHs, Landrum and Scavia (1983) estimated that anthracene associated with sediment and pore water contributed 77% of the amphipods' (Hyalella azteca) steady-state body burden. Considering reversibility between aqueous phase and sediment-associated PAHs, the differential availability of dissolved versus sorbed PAH should primarily be a kinetic consideration. That is, the particulate PAH reservoir should be the primary source, even though actual uptake may occur via a dissolved pathway.

Aquatic organisms are capable of accumulating PAHs through diet. In cases where uptake from food versus sediment has been compared, the dietary route appears to be more efficient (McElroy, 1985; Corner et al., 1976). For instance, in comparing bioaccumulation and metabolism of ¹⁴C-anthracene by the omnivorous deposit-feeding polychaete in microcosms where the PAH was introduced either already sorbed to sediments or in a prepared protein-based diet, McElroy (1985) noted that the isotope in the prepared diet was more rapidly metabolized than the isotope that bound to sediments. However, studies comparing direct uptake from solution to that from dietary routes are contradictory. In hard clam larvae (Mercenaria mercenaria), Dobroski and Epifanio (1980) found greater efficiency of ¹⁴C-benzo[a]pyrene uptake from the water column than from contaminated algae (Thaslassiosira pseudeonana), although the contaminated algae contributed significantly to the PAH body burden of the larvae. Lu et al. (1977) compared uptake of ¹⁴C-benzo[a]pyrene (added to water column) by fish (Gambusia affinis), mosquito larvae (Culex pipiens quinquefasciatus), and snails (Physa sp.) exposed individually and collectively in a model ecosystem. In single species exposures, no bioaccumulation was observed in fish, whereas mosquito larvae and snails attained B[a] P levels (in terms of radioactivity) 40 to 2 000 times that of the water column. When the organisms were exposed together, bioconcentration factors increased dramatically for all groups including fish (up to 1 000 times). On the other hand, in an experimental food chain consisting of diatoms, mussels, and snails exposed to ¹⁴C-naphthalene, Clark (1983) found that partitioning from seawater across membranes is a much more important route for PAH accumulation than the trophic transfer. Obviously, as noted from the discussion above, the topic of PAH bioaccumulation needs further exploration.

The bioconcentration factors in aquatic organisms exposed to PAHs in water are shown in Tables 23 and 24.

6.8 Sediment toxicity

The toxicity of PAH adsorbed to sediments is not well studied. It is, however, generally recognized that association with sediments reduces the bioavailability of PAHs.

Swartz et al., (1988) calculated 10-d LC₅₀s of 3.68 µg/g (dry weight) phenanthrene and 4.20 µg/g (dry weight) fluoranthene for the marine amphipod Rhepoxynius abronius exposed to these PAHs in sediment. However, working with contaminated sediments of Eagle Harbor in Puget Sound, Washington, these investigators found different results (Swartz et al., 1989). To assess toxicity, Rhepoxynius abronius were exposed to several sediment samples obtained by