Spatial and Temporal Variations in Chemical ...

riations in Chemica Contamination of frat4 Captured i the Estuary of the St. Lawrence River P.V. ~odson,'M . Castonguay, and C.M. Couillard Can. J. Fish. Aquat. Sci. Downloaded from www.nrcresearchpress.com by Fisheries and Oceans on 07/29/15 For personal use only.

MinisfPre des Pt;ches et des Oceans, %nstitutMaurice-Lamsntagne, G.P. 1000, 850, route de la Mer, Mnnt-jsli, QG 6 5 H 3Z4, Canada

C. Desjardins Minist&re des P&ches et des Oceans, Direction o'@ Iflnspection, 789, bod. Roland Therrien, Longueuil, QC 14H 4A6, Canada

INRS-Oc&anslsgie, Wniversite' du Que'bec, 3 70, w e . des Ursusines, Rirnouski, QC 65%3A 7, Canada

and R. Mebeod Henor-, Environmentai %absrcafsries,5555 North Service Rd., Bkirlington, O N k 7 L 5H7, Canada

Hodson, P.V., M. Castsnguay, C.M. Couiilard, C. Desjardins, E. Pel[etier, and R. Mcleod. 1994. Spatial and temporal variations in chemical contamination of American eels, Anguilh rostrata, captured in the estuary of the St. Lawrence River. Can. I. Fish. Aquat. Sci. 5 1 : 464-478. Levels of polychlorinated biphenyls (PCBs), mirex, and pesticides were 10-100 times higher in migratory adult American eels, Anguilla rostaata, sampled at Karnouraska, Quebec, in 1994) than in eels from an uncsntarr~inatedreference tributary; concentrations in eels within the estuary varied little among sites. In contrast, mercury levels were the same at estuarine and reference sites, suggesting natural sources or atmsspheric deposition of mercury. Dioxins, furans, and polynuclear aromatic hyalrocarbons were virtually absent at all sites. During the 7 wk of migration, levels of PCBs, n~irex,and pesticides in eels increased, while mercury did not. Since 1982, levels of PCBs and mirex have declined by 68 and 56%, respectively, and the percentage of eels exceeding human health guidelines for PCBs and mirex was about twofold lower in 1990 than in 4 982. bevels of other pesticides have also declined, except that dieldrin is unchanged since 1982. While chemical concentrations are declining, levels of specific, highly toxic PCB congeners are sufficiently high that eel consumption by beluga whales (Dejphinapterus leucas) is still likely to be hazardous. The hazard to the eels themselves is unknown due to a scarcity of toxicity data, but the highest concentrations of chemicals were observed in gonads.

Cher des anguifles d1Am6rique, AwguilBa rostrata, adultes en migration, echantillsnnees a Karnouraska (Quebec) en 1990, les concentrations de biphknyles polychlor4s (BPC), de mirex et de pesticides btaient 10 100 fois plus elevees que chez %esanguilles provenant d'un affluent tkmoin non pollue; dans I'estuaire, les concentrations variaient peu chez les anguilles d'un site 5 I'autre. Par contre, les teneurs en mercure btaient les rnemes dans I'estuaire et aux sites tbmoins, ce qui semble indiquer I'existence de sources naturelles s u de d6p6t d'origine atmospherique du rnercure. Les dioxines, Bes furanes et ies hydrocarbures aromatiques polynucleaires 6taient pratiquernent absents de tous les sites. Pendant les sept semaines de la migration, les concentrations de BPC, de mirex et de pesticides chez Ies anguilles augrnentaient, mais pas ies teneurs en rnercure. Depuis 1982, les concentrations de BPC et de rnirex ont baisse respectivement de 68 j, 56%, et le pourcentage d'anguilles chez lesquelles la teneur en BPC et en mirex dkpassait Ie niveaee autorisk pour la santk humaine 6tait envirsn cjleux fois plus bas en 1990 qea'en 1982. Les concentrations d'autres pesticides ont aussi baissk, sauf celle de dieldrine qui ne bsuge pas depuis 1982. Tandis que les concentrations de produits chirniques dirninuent, les teneurs en certains cong6ni?res de B?C extrernement toxiques restent assez klevbes pour que la consomrnation d'anguilles par le bkluga (Delphinapterus leucas) reste dangereuse. be risque pour Bes anguilles elles-rn@mesn'est pas csnnu, 3 cause de la rarett? des donnkes sur la toxicite, mais c'est dans les gonades qu'on a observe les concentrations de produits chirniques les plus 6lev6es. Received January 4, 7 993 Accepted September 7 4, 1 993 4563 747)

A

merican eels, Angegil!a rostrata, spawn in the Sargasso Sea, migrate as larvae to coastal streams s f North America, and spend most of their lives maturing in

freshwater (Schmidt 1925). En the St. Lawrence drainage basin, eels can ascend as far as Lake Ontario and may spend 12-16 y r growing from a size s f 91.0 g to 4808 g

' ~ u r r e n taddress: Environment Canada, National Water Research Institute, Canada Centre for Inland Waters. P . 0 . Box 5058, Burlington, OM k7W 4Ad. Canada. 464

Can. J. Fish. Aquat. Sci., Vo1. 51, 1994

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(Castonguay et al. 1994). Their diet ranges from macroinvertebrates to small fish, and they accumulate high concentrations of the chemical contaminants commonly found in the Great Lakes ecosystem (Castonguay et al. 1989). The accumulation of contaminants by eels may cause toxicity both to the eels and to their predators. Mass mortalities of eels have occurred in the freshwater portions of the St. Lawrence River, deformities have been noted by fishermen in the estuary, and the number of juveniles migrating towards Lake Ontario has diminished by more than 98% since 1985 (Homer 1 986; Dutil et al. 1987; Hendrick 1991). Mass mortalities were associated with gill damage and impaired ionoregulation, likely due to the combined effects of osmotic stress from saltwater adaptation and gill damage due to migration through industrial effluents (Dutil et al. 1987). Chronic toxicity was not suspected, since these effects did not correlate with tissue levels of persistent chemicals such as polychlorinated biphenyls (PCBs). However, the analyses did not measure highly toxic compounds now known to occur in the aquatic environment, such as specific congeners of PCBs, polychlorinated dibenzo[p]dioxins (BCBDs), and polychlorinated dibenzofurans (PCDFs) (Eisler 1986). Recent studies of beluga whales (Dekphknapterus leucas) in the estuary and the Gulf of St. Lawrence suggest that contamination of eels may have other important effects. Dead beluga contain levels of PCBs, BDT, pesticides, and mirex far above those found in beluga from the Arctic (Mass6 et al. 1986; Martineau et al. 1987; Muir et al. 1990). Marine mammals are probably exposed via their diet, since there is no contact of respiratory surfaces with water, which in fish permits direct chemical uptake. Therefore, the majority of contaminants in beluga must come from prey consumed over their lifetime. Since beluga are long-lived (20-30 yr), are predatory, and do not grow after 6-10 yr of age, they should accumulate high levels of lipophilic, persistent compounds such as PCBs and mirex. However, concentrations measured in beluga are too high to be explained simply by the consumption of fish from the estuary (BCland et al. 1987; Hickie et al. 1991). Furthermore, rnirex is not often found in estuarine fish, but was thought to be unique to fish from the upper St. Lawrence River and Lake Ontario, the only Great Lake to receive direct inputs (Kaiser 1978; Dutil et al. 1985; Slsterdijk 1985; Castonguay et al. 1989). Since eels contain contaminants at levels 10 times those of estuarine fish, and since a high proportion contain mirex due to their exposure in Lake Ontario, they may be an important transport route of chemicals to beluga (kum et al. 1987; BCland and Martineau 1988; Hickie et al. 1991). We collected and analyzed migratory adult eels ("silver eels") from the St, Lawrence estuary in 1990 to evaluate the effects of contaminants on their health, the role of eels in contaminating beluga whales, and the changes in contamination since the last major survey in 1982 (Desjardins et al. 1983a, H983b; Dutil et al. 1985; Castonguay et al. 1989). Since eels migrate from many tributaries to the St. Lawrence, as well as from Lake Ontario, it cannot be assumed that a sample taken at one glace or time represents all eels. Indeed, the freshwater origin of silver eels can be distinguished by their burden of rnirex (Dutil et al. 1985; Castonguay et al. 1989), and deformed eels are more common at the end of the annual run than at the beginning (Homer 1986). Therefore, heterogeneity in a sample of eels and in their contaminant Can. 3. Fish. Aquat. Sci., VooH. 51, 1994

burden could be induced by the distance they must migrate (e.g., those starting in the Great Lakes versus those starting in tributaries closer to the estuary), their ability to swim (toxicity, deformities), and their river of origin and distribution in the estuary. Since the eel fishery is coastal and eels can be captured on both the north and south shores, it is possible that the eels migrate along a coast according to their river of origin. Therefore, we tested the following null hypotheses: (a) levels of chemicals do not occur at concentrations that are toxicologically significant, (b) there are no temporal changes in average levels of chemicals in eels during the annual migration, as indicated by weekly sampling of migrating eels, (c) contamination has not changed since 1982, the last major sampling of migrating eels, and (d) contamination of migrating eels does not vary spatially, as shown by eels sampled within the St. Lawrence estuary and in a tributary to the estuary. Since eels from the St. Lawrence estuary were lager than those sampled in a tributary, size may confound comparisons among locations. Therefore we tested the hypothesis that (e) contamination of migrating eels does not vary with size, as measured by whole wet weights at the time of sampling. Furthermore, since beluga would eat whole eels, but only gutted carcasses are analyzed in surveillance programs, we compared the levels of chemicals in whole eels with levels in specific tissues, testing the hypothesis that (f) concentrations of contaminants do not vary among tissues or between tissues and "whole body".

Materials and Methods This study focussed on maturing eels migrating seaward for spawning. While the term "silver eels" has been applied to this stage, coHour was not a criterion for sampling, since colour Is highly variable and can change within hours (Scott and Crossman 1973). The criterion used was simply capture in commercial or experimental traps oriented to intercept those eels migrating seaward. While size varied considerably, advanced maturity was confirmed by histological examination of oocytes (C.M. Couillard, unpublished data); all fish sampled were female. Sample Collection Random samples of about 100 live silver eels were purchased from a commercial fisherman at Kamourash, Quebec (Fig. I). Samples were collected weekly during the entire 7 wk of the annual fishery, between September and November 1990. During this 7-wk period, live eels were also purchased on one occasion each from fishermen at Cacouna, on the south shore of the estuary, and at Saint-IrCnke, on the north shore. The eels were trapped during their seaward spawning migration by commercial weirs installed at angles to the shoreline. Most of these eels would be from the St. Lawrence River watershed, upstream of Quebec City (Dutil et al. 1985). Live eels were also captured by experimental trap nets during their seaward migration in Kvi&re aux Bins (Fig. I), an uncontaminated reference stream that enters the sea from the north shore of the St. Lawrence estuary. All eels were returned live to the laboratory where they were held in clean, flowing salt water (30 parts per thousand salinity) for 4-48 h before being killed for tissue analyses. Each was anaesthetized in MS-222 (tricaine


FIG. 1. Sampling sites. Eels were purchased at commercial weirs at Kamouraska, Saint-IrCnCe, and Cacsuna in the estuary in salt water and were captured by a trap set in freshwater in Rivikre aux Pins.

rnethanesulfonate), the viscera and gonads were removed, and the carcass with head attached was frozen in a csntaminant-free plastic bag (ARCAN Inc., Plainswell, Mich.). After all sampling was completed, the frozen eels from each week and each site were subsampled to provide two each with an original live weight of about 308, 688, 900, 1288, 1588, 1800, 2100, 2400, 2700, or 3000 g. The smallest and largest eels were scarce so that N varied from 7 to 16 fish per sample. For eight additional eels, analyses were made of the head, gutted carcass, gonad, and viscera (liver, stomach, intestines, etc.) to compare whole fish and tissue contaminant levels. For pslyn~acleararomatic hydrocarbons (PAH), PCDD, and PCDF analyses, only a limited number of fish were analyzed due to cost. Hsmogenates of two fish in the 980-g weight class were pooled to provide one sample for each s f the seven sampling times at Kamouraska and one for the fish collected at Rivikre aux Pins. The fish from Kamouraska were purchased at the same location as in 1982, and chemical analyses were by the same laboratory with the same methods as in 1982, to maintain continuity of monitoring data (Desjadins et al. 1983a, 198%). Sample Preparation Each eel was thawed, the head removed, and the remaining gutted carcass (muscle, skeleton, and skin) homoge466

nized with a Hobart meat grinder using a 4-mm hole size. The homogenate was collected in stainless steel bowls, rehomsgenized, and stored as 100-g aliquots in glass jars, using stainless steel or plastic kitchen utensils rinsed in ultrapure hexane. The glass jars were also rinsed in hexane, and a liner of hexane-rinsed foil separated the homogenate from the plastic jar lid. After each fish, the grinder was disassembled and cleaned, which included a hat water wash, scrubbing, rinsing with water, and rinsing with ultrapure hexane. Individual tissue samples were homogenized in the Hobart grinder (head), or in a smaller laboratory grinder (viscera and gonads) to avoid excessive loss of material. It was not necessary to empty or clean the digestive tract, since eels do not feed during their migrrdtion and the stomach is partially resorbed. All homogenates were frozen and maintained at -2Q°C until analy sis. sample

~~~l~~.~

Several chemicals were measured and detection limits varied according to the methods used. The methods are described in detail in Hsdscsn et al. (1992) and summarized briefly here. Lipids were not measured in most samples, except for the analyses s f individual tissues. Can. 9. Fish. Aquar. Sci., bolo 56, 6994


Carcasses Organochlorine compounds were extracted from homogenates with an acetone-hexane (1: I) mixture and cleaned by gel permeation chromatography (GPC) and 2% deactivated Florisil (Johnson et al. 1976; Mills et al. 1972). Hexachlorobenzene (HCB), DDE, PCBs, and mirex (fraction F1) plus 12 other pesticides (fraction F2) (McLeod and Ritcey 1978) were measured by gas chromatography with an electron capture detector (GC-EC) (Sherma and Beroza 1980; Freeman 1981), using 2% OV-17 and 2.6% 0V-210 packed columns. Three peaks were used to calculate total Aroclor 1254 (IUPAC Nos. 127, 147, and 177; Reynolds 1971). Recoveries ranged from 90 to 110% and were uncorrected. PCB congeners and pesticides were separated on a capillary column and quantified by peak-to-peak comparison (Martineau et al. 1987). Mercury was analyzed in acid-digested samples by selective reduction to elemental mercury by SnC12, using a Technicon autoanalyzer with a mercury monitor detector (Armstrong and Uthe 197 1; MPO 1989). For analyses of PCDDs and PCDFs, homogenates were mixed with anhydrous Na,SO, and radiolabelled surrogates and extracted several times with dichloromethane (BCM). The combined extracts were concentrated to 5.0 mL, diluted 50:50 with cyclohexane, and cleaned up by GPC after centrifugation to remove particulates. The extracts were solvent exchanged into isooctane, and recoveries ranged from 98 to 114%. The extracts were cleaned following Dow methodology (Lamparski et al. 1979), and final extracts were eva orated and taken up in 18 pE of isooctane with 250 pg [ HI,] triphenyleneepl-I as an internal standard. PCDDs and PCDFs were measured with a VG70-VSE high-resolution gas chromatograph mass spectrometer (GC-MS), and the elution windows for each congener group were established by injecting a mixture containing the first and last congener to elute. Concentrations were estimated from responses of individual 2,3,7,8-substituted standards. Average responses for the 2,3,7,8- substituted congeners of each congener group were used to calibrate the instrument response for all isomers of that group. Individual congeners were identified by their specific ions and ion ratios at an instrument resolution s f I in 10 (BOO. Identification was also made with a minimum signal-to-noise ratio of 3 to I and a retention time within a 3-s window of external standards after internal standard correction. None of the data were blank corrected but they were corrected for surrogate recoveries. For PAHs, tissue extracts were fractionated on a silica column. Fraction F2 was retained, reduced to 0.5 mE, and an internal standad added pkos to GC malysis with a flame ionization detector (GC-FID) (Hsdson et al. 1992). Compounds were identified by comparing retention times with authentic standards (Supelco PAH Kit 610-M; Supelco Canada Ltd., Bakville, Ont.). All peaks identified as PAHs by the retention time method were cross checked by GC-MS.

P

Tissue comparisoa Fish samples (10-12 g wet wt) were mixed with anhydrous Na,SO, and extracted three times with 50 mL of hexane-acetone (70:30) at 3 5 T . The extracts were reduced in volume, particulates were removed by centrifugation, and lipid was removed by GPC (Norstrom et al. 1986). Samples eluted in DCM-hexme (50:50) were evaporated to 0.5 mL and split into two fractions using a florisil column eluted Can. J. Fish. Aquat. Sci., Vol. 51, 1994

by hexane ( F l ) and hexane-DCM (75:25) (F2). Fraction volumes were reduced to 0.5 mL and an internal standard was added for analysis. Chlorinated compounds were analyzed by GC-EC by cornparing retention times with those of a chlorinated pesticides standard ( ~ u ~ e l ~ r e r n pesticides e @ - ~ ~ mix: Supelco Canada Ltd., Oakvil%e,Ont.) or a PCB standard containing Aroclors 1242, 1254, and 1260 (1: 1:1) (Supelco Canada Ltd., Bdcville, Ont.). Congeners were identified according to Schultz et al. (1989). Identities of some major congeners were confirmed by GC-MS. The overall variability of the method was 115% and the detection limit 0.00081 pg-g wet wt-'. Total mercury was analyzed in duplicate by flameless atomic absorption spectrophotometry (AAS) against aqueous standards (detection limit = 0.005 pg-g wet wt-') after Sample digestion at 55°C by nitric, sulphuric, and hydrochloric acids (after Hatch and Ott 1968). The overall variability of the method was 110%. Total cadmium, copper, chromium, and zinc were determined in samples digested t h e e times at 75°C in 58% hydrogen peroxide for 1 h, followed by concentrated H N 0 3 for 2.5 h. Extracts were analyzed by graphite furnace AAS against aqueous standards. Wet weight detection limits were 0.01 Fg-g.-' (Cd and Cr), 0.07 pg-g-l (Cu), and 3 pg.g.-' (Zn) and variability was 4% for Cd, Zn, and Cu and 10% for Cr. 'Toxicological Significance The toxicological significance of PCB, PCDD, and PCDF congeners can be judged by equating their toxicity to that of dioxin (TCDD). the highly toxic 2,3,7,8-tetrachlorodibenzo[p] Based on similar modes of toxic action, the toxicity of each congener may be expressed as a fraction of the toxicity of TCBD (Safe 1990). When these toxic equivalent factors (TEFs) are multiplied by the concentration of the corresponding congeners, concentrations are expressed as TCDD equivalents (TEQs), which are summed over all congeners to give the total TEQs. Equivalents developed from mammalian models are used to estimate the risk to humans from consuming food contaminated by PCBs, PCDDs, and PCDFs and were used here to judge the hazard to beluga whales of consuming eels. For other compounds, significance was judged by comparison of measured concentrations with Canadian Guidelines for Human Consumption for chemicals in fisheries products. Since the guidelines are intended to protect human consumers of fish, levels of chemicals exceeding the guidelines would presumably represent a risk to other mammalian consumers. Statistics All data except weights were converted to logarithms before statistical malysis due to the large ranges in values and the resultant nonnormal distributions and nonhornogenous variances. Log-transformed data met assumptions of normality and homogeneity of variance; contaminant levels are reported as geometric means. Many compounds were not detectable and a value of one half the detection limit was substituted for statistical analysis, a common convention in contaminant monitoring programs (B.M. Whittle, Department of Fisheries and Oceans, Burlington, Ont., personal communication).

TABLE1. Ranges of geometric mean concentrations of eight PCB congeners, pesticides, mirex, and mercury in 89 eels sampled weekly for 7 wk at Kmouraska in 1990. The F-statistics are from ANOVA testing the effect of week of sampling, and p is the probability that variations among weeks were by chance. An asterisk indicates statistical significance at the 0.05 probability level.


Compoupad

Range of geometric mean concentrations (pg-g-')

FWee,,

P

PCB congener 28 52 801 118 153

137 138 180 Sum of congeners

Total Arochlor 1254 Pesticides p,pl-BDE ppr-DDD p,pr-DDT Total DDT Hexachlsrsbenzene a-BHC

Weptachlor Heptachlor epoxide Lindane Aldrin Endrin Dieldrin Oxychlasrdane l -Chlordane 2-Chlordane Total pesticides Mercury

For the study of tissue distribution, the bbwhole-body9q concentrations of contaminants were calculated from the weights and chemical concentrations of the individual parts. For the eight eels tested, the carcass, viscera, gonads, and head averaged 83.3, 3.1, 4.9, and 8.7% of the whole-body wet weight. Hence, the concentration (C) of a chemical in the whole body was calculated as

Comparisons of chemical levels in carcasses among weeks were made by analyses s f variance (ANOVAs) using the general linear model (GEM) available as a SAS computer routine (SAS Statistics for Applied Sciences, version 6.0, SAS Institute, Inc., Box $000, Cary, NC 275 12-8000). Site comparisons by ANOVA were made in pairs, since chemical levels varied with weeks. Hence, results at Rivikre a m Bins, Cacouna, and Saint-Irknke were compared individually with results at Kmsuraska for the corresponding weeks of sampling. The comparison s f 1990 data with 1982 data was also made by ANOVA. All comparisons were made at the 95% confidence level, using Tukey's test to discrirnimate differences among means. To assess the effect of fish

size on contamination, coefficients of determination ( r 2 ) were calculated between chemical concentrations and total weight for all fish sampled at Kamouraska.

Results Levels of Chemicals in Relation to Toxicity Although only a small number of PCB congeners were measured in carcasses, they were found c~nsistentlyand at concentrations well above the detection limit (Table I). Congeners 28 and 137 (HUPAC nomenclature, Mullin et al. 1984) were the least concentrated and were found most frequently at levels less than 0.01 pg-g- l. Congeners 52 and 101 ranged from 0.004 to 0.336 in individual fish, about one half to one third of the concentration of congener 180, which ranged from 0.010 to 0.524 pg-g-'. Congeners 118, 153, and 138 were most concentrated, with levels from 0.001 to 0.833 The sum of the individual congeners ranged from 0.156 to 3.033 pg.g-l, while PCBs expressed as "Aroclor 1254", a specific mixture of congeners, ranged from 0.149 to 7.489 pg-g-'. However, these analyses were highly variable. Within any sample of 7-16 eels, concentrations (reference site) to could range from less than 0.2 Can. J. Fish. Aquat. Sci., Vol. 51, I994

TABLE2. Summary sf Canadian Guidelines for Human Consumption for the levels sf chemicals in fishery ~ B O ~ Uand C ~the S ,percentage of fish that exceeded these guidelines at Kamouraska in 1982 and 1990 and at Rivibre aux Pins in 1990. -

- - -

-

-

-

-

-

Percentage of fish exceeding the guideline" Guideline (pg0g-')


N PCB (ArschBor 1254) Mirex 2,3,7,8-TCDD

2.0

DDT

5 .O

Other pesticides Mercury

0.1 0.5

0.1 0.00002

A11 eels, 1982

Kamsuraska, 1982

396 62 31

104 80 52

-

-

0 33.1 9.3

0 13.5 8.6

Kamouraska, 1990

Riviere aux Pins, 1990

89 36 29 0 0 15.7~ 2.3

"Data for 11982 were derived from Desjardins et al. (1983a, 1983b), Castonguay et al. (19891, and C. Desjardins (unpublished data). '~ieldrinin all cases. greater than 2.0 The coefficients of variation of geometric means for fish sampled weekly at Kamouraska ranged from 104 to 185896, with a mean of 387%. For Aroclor 1254, 32 of 89 (36%) fish at Kamouraska exceeded the guideline for PCBs in fish products of 2.0 pg-g-' (Table 2). The analysis of PCB congeners in the eight fish used for the tissue comparison was much more detailed, and there was a wide variety in every tissue examined. Only 14 of 88 congeners or congener groups (Nos. 719, 34, 49, 96, 88, 136, 175, 167, 197, 193, 169, and 198) were undetectable in every tissue. In contrast, the average concentrations of seven congeners (Nos. 68156192184, 969110 1, 99, 7711 10, 123114911 18, 13211531105, and 128) exceeded detection in every tissue; congeners 12311491118 and 1321153/105 were most concentrated. Overall, there were many congeners that were not measured in the larger sample of 89 fish, but which contributed significantly to the total levels of PCBs in the eight individual eels. Not all PCB congeners are equally toxic; coplanar structures most similar to that of TCDD are most toxic. Hence, the hazard of consuming eels can be estimated by reference to only a small subset s f congeners (TEFs): Nos. 126 (0.1), 169 (0.05), 77 (0.01), 118 (0.001), 105 (O.OBBl), 123 (0.001), 156 (0.001), 157 (0.001), 167 (0.6901), and 189 (0.001) (Safe 1998). These congeners were only measured in the eight eels sampled for tissue distribution, and the range of total concentrations within this sample (0.097-1.427 pgeg-L) was lower than the range of values for the more extensive data set on carcasses. Unfortunately the analytical techniques used did not completely separate all congeners. Hence, No. 126 coeluted with Nos. 129 and 78 and its contribution to TEQs is likely overestimated. The same is true of Nos. 77, 118, 185, 123, 156, and 157. If these mixtures are treated simply as the congener of interest, the TEQ for all eight fish due to all of the toxic congeners was 0.001885 Fgg-' or 1885 parts per trillion (ppt), with 86% of the TEQs contributed by No. 126. Despite the unknown bias in treating a mixture of congeners as pure congener 126, it is likely that PCB congeners in eels pose a hazard for beluga whales. For fish, far less is known about the toxicity of PCB congeners and only a few TEFs are available to predict their hazards. Furthermore, some congeners that are not considered toxic in mammals have significant effects on fish. For example, congener 81 induces the mixed function oxygeCan. J . Fish. Aquat. Sci.,

k701. 51,

1994

nase enzymes of fish (Janz m d Metcalfe 1991) and is embrystoxic (C. Metcdfe, Trent University, Peterborough, Ont., personal communication). This congener was not included in our analyses but it is not often found in environmental samples (M. Servos, Department of Fisheries and Oceans, Burlington, Ont., personal communication). For other congeners of known toxicity, No. 169 was not detectable and No. 126 (fish TEF = 0.005) coeluted with Nos. 129 and 178, as did No. 77 (fish 'TEF = 0.002) with No. 110. Hence, there is not a solid basis for judging the hazards of toxicity. Assuming that all of the mixtures containing congeners 126 and 77 could be assigned to these congeners, and that there are no others toxic to fish, the geometric mean TEQs in muscle and gonads were 0.0001 17 (117 ppt) and 0.000130 (130 ppt), respectively. PCDDs and PCDFs were found only at very low csncentrations, a ~ many d congeners were not detected (Table 3). The most toxic congener, 2,3,7,8-TCDD, was detected only once, at 0.0000814 yg(1.4 ppt). well below the guideline of 0.000020 kg-g-$ (20 ppt) (Table 2). The only congeners detected frequently were hepta- and octachlorinated PCDDs and PCDFs, at levels up to 0.00012 i*g.g-l (120 ppt). Hepta- and octachlorinated congeners are the most commonly observed in environmental samples, but sf least concern toxicologically, being 100- to 1000-fold less toxic than 2,3,7,8-TCBD (Table 3). Other tetra-, penta-, hexa-, and heptachlorinated congeners were also observed, but even on the basis of total hornologeres, concentrations were low relative to the guideline in Table 2. Based on the mammalian TEFs for PCDDs and BCDFs (Table 3), TEQs in eels ranged from 0.0000001 to 0.0000011 Fg.g-' (0.1-1.1 ppt), levels far below the guideline of 20 ppt (Table 2). The hazard of dioxins was difficult to judge for eels. The tissue dose associated with TCDD toxicity to rainbow trout (Orzcorh~~nchu.~ rnykis,~)embryos is 0.000065 yg-g ($5 ppt), about 50 times higher than levels s f TCDD found in eels (Table 3). However, there are no TEFs for the other congeners measured, so that TEQs cannot be calculated. For other tetra-, penta-, and hexachlorinated PCDDs and PCDFs, TEFs for fish range from 8.028 to 0.73 (Walker et al. 1991). If these are applied to the congeners of Table 3, the total TEQs do not approach or exceed 65 ppt. No eels were found with con~entratafionsof PAHs above the limits of detection of 0.01 pg-g-'; hence, no results are

-'

-'

449

TABLE3. Levels of PCDDs and PCDFs (pg-g-l) in homogenates pooled from two eels per week or site. Congeners that did not exceed the detection limit (1.0 pg.g-l) are listed in footnote "a". Toxic equivalent factors (TEFsj express the toxicity of each congener as a fraction of the toxicity of 2,3,7,8-TCBD (NATOIGCMS 1988).

Kamouraska, week:


Congener

1

2,3,7,8-Tetrachlorodibenzo[p]dioxin c1 1,2,3,6,7,8-Hexachlorodibenzo[p~dioxin 1.1 B ,2,3,4,6,7,8-HeptachIorodibenzo[p]dioxin 6 1,2,3,4,5.6,7,8,9-0ceach10rodiRenzo[p~dioxin 55 Sum of tetrachlorodioxins