Lipid content of all tissue samples was measured because many of the anthropogenic contaminants measured in this study are lipophilic compounds. Blubber tissue generally has high lipid content in most marine mammals; the lipid content of blubber of many marine mammals is often found to be between 60% and nearly 100% (Aguilar 1985, Martineau et al. 1987, Geraci 1989, Henry and Best 1983). However, for most of the whales sampled in this study, the lipid content in blubber was relatively low, ranging from less than 1% to only 16%, with the exception of CRC 395 that had over 70%.
The low lipid content of these whales can potentially be attributed to three factors:
1) leaching of oil from the tissues after death, 2) poor nutritional condition because of fasting during the winter breeding season, and 3) poor nutritional condition due to ill health or other factors. Though some leaching of oil from tissues may have occurred, it appears unlikely that this could have accounted for the low lipid contents observed. All of the animals sampled were still intact and did not have visible oil leaching from the areas where the tissues were sampled. Additionally, the most decomposed animal examined (CRC 395) had the highest lipid content. Part of the reason for the low lipid content is likely a result of the postbreeding season timing of the deaths. The only whale sampled that had died prior to the breeding season (CRC 395) was the only whale with a high lipid content in the blubber (70%). This individual was sampled in February and was very decomposed suggesting it had probably been dead at least a month. This finding, in addition to the fact that CRC 395 was sampled earlier than the other whales, would indicate that its death occurred during the period of the southbound migration from the feeding to the breeding grounds. Rice and Wolman (1971) report that northbound migrating whales (postbreeding season) have decreased weights, girths, blubber thicknesses, and oil content compared to southbound migrants (postfeeding season).
The lipid content of blubber from most of the whales sampled
is lower than what would be expected as a result of their seasonal
status alone. The difference in overall body oil yield of northbound
and southbound migrants reported by Rice and Wolman (1971) was
small (38.1% oil versus 39.6% oil, respectively). The overall
percent oil in the body they reported for northbound migrants
was much higher than the lipid content we found in the blubber
of these animals. The low lipid content in the sampled whales
appears to have likely been the result of poor nutritional condition
from other than just the postbreeding season timing. It is difficult
to verify the poor nutritional status of these animals from the
gross examinations. Some animals showed below normal girths and
blubber thicknesses (J. Calambokidis, pers. commun., June 1992).
Blubber thickness, however, is a poor indicator of nutritional
status in gray whales (Rice and Wolman 1971), with a thick fibrous
layer remaining even in whales with no lipids. The girth of decomposing
whales also could not reliably be used because of potential distortion
due to bloating.
The results show that all whales contained detectable levels of CHs in most tissues
(Table 5 and Appendix A); however, the summed concentrations of CHs in blubber, liver, brain, and stomach contents of animals were generally less than 5,000 ng/g wet weight (parts per billion) with a high concentration of 20,000 ng/g wet weight in blubber of a whale (GLR 101) stranded along the Strait of Juan de Fuca (Fig. 1). The relatively low concentrations of lipophilic CHs in most of the whales may be related to the low lipid content in analyzed tissues. However, we found that the whale (CRC 395) having the highest lipid content in blubber did not have exceptionally or proportionately higher concentrations of these anthropogenic compounds. Moreover, whether the data on concentrations of CHs were based on wet weight or lipid content, the relationships with respect to contaminant concentrations did not change significantly or alter any of the conclusions with regard to region-specific differences in contaminant concentrations.
Blubber, liver, brain, and stomach contents were available from one whale (CRC 334, Hartstene Island). The highest summed concentrations (wet weight) of PCBs, DDTs, and chlordanes, and of dieldrin and hexachlorobenzene were observed in blubber and following in descending order were the concentrations in liver, brain, and stomach contents. This order does not strictly parallel the order of the total lipid content among the tissues and stomach contents; thus, the proportion of specific lipids (e.g., triglycerides and nonesterified fatty acids) in each compartment may better correlate with the concentrations of these lipophilic CHs in tissues and stomach contents.
The summed concentrations of PCBs were higher than other CHs, including DDT and its derivatives, in both blubber and liver tissues (Table 5) regardless of the site of the stranding, whereas the summed concentrations of DDT and its derivatives were generally higher than other chlorinated pesticides, such as the summed concentrations of chlordanes (Table 1). The same relationship between summed concentrations of PCBs and pesticides was observed in brain tissue obtained from whale CRC 334.
Information is not available on the levels of CHs in either sediment, invertebrates, or fish from the exact locations where the whales stranded. However, for Puget Sound in general, the range in summed concentrations of CHs in livers of adult English sole (Pleuronectes vetulus) is 200 to 14,000 ng/g (wet weight), with the lowest values for sole from relatively uncontaminated nonurban areas and highest values for areas such as the urban Duwamish Waterway (Johnson et al. 1988). Thus, the summed concentrations of CHs (110 to 2,500 ng/g) in these whale livers were comparable to those measured in livers of adult English sole from areas of relatively low contamination. In addition, in the San Francisco area where four of the whales stranded, the range of CHs in livers of starry flounder (Platichthys stellatus) and white croaker (Genyonemus lineatus) is 990 to 2,400 ng/g wet weight (Varanasi et al. 1989a) and the summed concentration of CHs (190 ng/g) in the one liver sample obtained from a whale stranded in San Francisco Bay was comparable to the values for fish livers from relatively uncontaminated sites.
A comparison of summed concentrations of PCBs and DDTs in tissues from these stranded gray whales to the values for baleen whales from other studies (Wagemann and Muir 1984), shows that the gray whales from this study had comparable or lower concentrations (wet weight) of these contaminants. For example, the ranges in summed concentrations of PCBs and DDTs in blubber of humpback (Megaptera novaeangliae), fin (Balaenoptera physalus), and minke (Balaenoptera acutorostrata) whales were less than 10 to 7,000 and less than 10 to 23,000 ng/g wet weight, respectively, whereas the ranges in summed concentrations of PCBs in blubber of gray whale in the present study were 120 to 10,000 ng/g wet weight. In contrast, summed concentrations of PCBs and DDEs (major metabolite of DDTs) in toothed species, such as killer whales (Orcinus orca), harbor porpoise (Phocoena phocoena), range from 150 to 250,000 ng/g wet weight, and 550 to 640,000 ng/g wet weight, respectively (Calambokidis et al. 1984). The lower concentrations of these pollutants in baleen whales is consistent with the gray whales' primary food source being invertebrates which is in contrast to predatory seals, porpoises, and toothed whales (e.g., killer whales) which feed on organisms higher in the food chain, such as fish and other marine mammals.
As discussed above, the extensive metabolism of PACs by marine mammals makes it nearly impossible to quantitatively assess exposure of these animals to PACs, which are widespread and toxicologically significant environmental pollutants (Varanasi et al. 1989b). Hence, in the present study, the only assessment of PAC exposure was by measurement of PACs in stomach contents, which contain sediment and also contain invertebrates that generally metabolize PACs much less than do mammals. The concentrations (wet weight) of PACs (Table 1, Appendix B) in stomach contents, which were available from eight of the whales sampled, ranged from 7 to 2,100 ng/g, with a mean value of 440 ng/g (Table 6). For Puget Sound, the range of concentrations (wet weight) of PACs in sediment is approximately 4 to 3,000 ng/g, with PACs in stomach contents of benthic fish ranging from 29 to 14,000 ng/g (Varanasi et al. 1989a).
In the future, the assessment of exposure to PACs can potentially be enhanced by the application of techniques developed for fish that measure metabolites of PACs in bile (Krahn
et al. 1986) and tissues (Krone et al. 1992) and metabolites bound
to hepatic DNA (i.e., DNA-xenobiotic adducts). Application of
these techniques to marine mammal tissues and bile, however, is
hampered by the often poor quality of the samples obtained and
the absence of gall bladders in many marine mammals (e.g., cetaceans).
For example, levels of DNA-xenobiotic adducts (Varanasi et al.
1989a) which represent the binding of potentially carcinogenic
PACs to DNA may be a useful indicator of exposure to PACs, but
only samples collected from tissues that have undergone minimal
autolysis can be analyzed.
In the present study, the concentrations of 16 elements in liver, kidney, and stomach contents of whales stranded in all areas were generally low (Table 7). The only elements in stomach contents for which relatively elevated concentrations (wet weight) were found were as follows: aluminum (190,000 to 3,400,000 ng/g, n = 8); manganese (5,100 to 54,000 ng/g, n = 3), iron (5,800 to 810,000 ng/g, n = 3), chromium (360 to 11,000 ng/g, n = 8), and barium (310 to 24,000 ng/g, n = 5). Elevated concentrations of aluminum were also found in stomach contents of the whale stranded near Port Angeles in 1984. The mean concentration (± SE) of aluminum in stomach contents of these whales (n = 8) was 1,700,000 ± 450,000 ng/g wet weight, which is equivalent to 1.7% by weight with the highest value being 3.4%. These values are comparable to the concentration (4 to 9%) of aluminum found in marine sediments (Cerundolo et al. 1988 and Loreti 1988). Further, the high concentrations of aluminum in the stomach contents of these eight whales were also accompanied by correspondingly elevated values for iron and chromium (Table 7). Moreover, the mean ratios of the concentrations of iron and chromium to aluminum in stomach contents were comparable to the ratios in sediment and suspended particulate matter ( Table 8). Thus, the findings of high concentrations of aluminum, chromium, and iron in stomach contents are consistent with ingestion of sediment in the natural feeding process of gray whales (Nerini 1984).
The high concentrations of aluminum in the stomach contents of the gray whales were accompanied by relatively high concentrations in liver and brain. In contrast to aluminum, barium, which was found in high concentrations in stomach contents, was not elevated in liver or kidney, indicating lower bioavailability of this element. Evaluating whether the concentrations of aluminum in tissues of gray whales are comparable to other marine mammals is not possible, because data on concentrations of aluminum in tissues of other marine mammals species are essentially nonexistent. However, the aluminum concentrations (2,300 ng/g) in brain of a gray whale in the present study and in the gray whale stranded in 1984 (8,800 ng/g) are within the range for some terrestrial mammals. For example, one review study (NRC 1980) listed the normal (control) concentrations (wet weight basis) of aluminum in the brain of rats as 7,100 ng/g, in cattle as 6,400 ng/g, and in sheep as 4,500 ng/g. These species may receive high concentrations of aluminum in their diet and hence their normal tissue burden may be higher than other species. Other control values of aluminum in various tissues for these species are 120 to 15,000 ng/g (liver) and 2,700 to 8,600 ng/g (kidney). These results were from toxicity studies in which aluminum was administered at dosages of 1,200,000 to 2,800,000 ng/g in water or food (NRC 1980). No effects were observed in cows and sheep fed these dosages of aluminum for 84 and 77 days, respectively. The range in control levels of aluminum indicates a broad range in the tolerance to aluminum in mammals, and hence demonstrates the importance of comparing concentrations for stranded animals to values in apparently healthy gray whales. These results, together with the fact that gray whales ingest sediment as part of feeding, would suggest that the concentrations of aluminum in tissues of gray whales reported here are within the normal physiological range. This can be confirmed, however, only by measuring tissue concentrations of aluminum in healthy gray whales.
The highest concentrations of mercury in livers were observed in the whales stranded on Hartstene Island (CRC 334) in south Puget Sound and near Lyre River (GLR 101) on the Strait of Juan de Fuca (100 ng/g and 120 ng/g wet weight, respectively). The mean concentrations (± SE) (wet weight) of mercury in liver and kidney (56 ± 12 ng/g, n = 10 and 34 ± 6 ng/g, n = 10, respectively) of the gray whales sampled in this study were relatively low when compared to the data for other cetaceans (Wagemann and Muir 1984). For example, the range of concentrations of mercury in liver of minke whales, a baleen whale, were found to range from 61 to 390 ng/g wet weight (Honda et al. 1987), while the range in mean concentrations of mercury in liver of porpoises and narwhal (Monodon monoceros) from several studies was 700 to 31,000 and in kidney 680 to 3,600 ng/g wet weight (Wagemann and Muir 1984). However, the range (nondetectable to 500 ng/g) of concentrations of mercury in the stomach contents of gray whales were within the range of concentrations (20 to 1,700 ng/g) found in stomach contents of porpoises and seals (Fujise et al. 1988 and Yamamota et al. 1987). As noted above, gray whales feed on benthic invertebrates and in the process ingest sediment that can be contaminated with toxic elements such as mercury. In contrast, porpoises and seals feed on higher trophic level organisms and would not ingest much sediment during feeding. Thus, the lower concentrations of mercury in gray whale liver and kidney compared to porpoises and seals would appear to indicate that mercury associated with sediment was not readily bioaccumulated by gray whales.
The concentrations of nickel, copper, zinc, cadmium, and lead in liver and kidney (Table 7) were similar among gray whales from different regions. Additionally, the concentrations of these elements in gray whales were comparable to the concentrations in liver of minke whales harvested between 1980 and 1985 (n = 135) in Antarctica (Honda et al. 1987) where the concentrations (wet weight) of nickel, copper, zinc, cadmium, and lead ranged from 20 to 430 ng/g, 6,900 to 25,500 ng/g, 74,300 to 175,000 ng/g, 6,600 to 100,000 ng/g, and 80 to 1,900 ng/g, respectively. Further, several other elements were present at low concentrations in the gray whales in this study. For example, the mean concentrations of silver and tin in liver and kidney were less than 40 ng/g wet weight, while the mean concentrations of selenium in liver and kidney were less than or equal to 2,000 ng/g wet weight. The concentrations of selenium reported (Wagemann et al. 1983) for narwhal liver and kidney (4,000 ± 1,800 and 3,100 ± 850 ng/g wet weight, respectively) are slightly greater than the concentrations of selenium in these stranded gray whales (Table 7).
Analysis of variance was used to assess if there were significant differences in concentrations of CHs, PACs, or elements in animals from different regions. A summary of the statistical analyses for selected CHs and elements is given in Tables 9 and 10. First, the data for all tissues and CH analytes were analyzed for region-specific differences without controlling for either sex of the animal or year of sampling; subsequently, a three-way ANOVA was performed for blubber, which was sampled from all animals, to determine if there were region, sex, or year specific effects. While controlling for the variances of sex and year of stranding, there were no significant regional differences in the concentrations of CHs in blubber (wet weight or lipid normalized basis) in these gray whales (Table 8). These results are consistent with the concentrations of CHs in blubber, reflecting more chronic exposure rather than relatively recent exposure, even if it is assumed that animals were feeding actively at the site where they were stranded.
The concentrations of CHs (Table 5) and elements (Table 7) in liver of whales from Puget Sound were compared to the concentrations in whales from the other four areas combined because only a small number of liver samples (10) was available. The statistical analyses revealed no site differences for any of the analytes. Further, no significant differences in concentrations of CHs, PACs, or elements were found for stomach contents of animals from Puget Sound as compared to those from whales from all other sites combined, excluding whales from California for which no stomach content samples were obtained. The only exception was the finding of a significantly higher concentration of higher molecular weight PACs (Table 6) in stomach contents of whales from Puget Sound, which has several urbanized areas with high concentrations of sediment-associated PACs, as compared to whales from other sites. However, because no significant differences were found in concentrations of other anthropogenic contaminants (i.e., CHs) in stomach contents of these animals, which can contain large amounts of sediment, and because only a portion of sediment-associated PACs is known to be bioavailable (Varanasi et al. 1985, Farrington et al. 1983), the significance of higher levels of PACs in stomach contents of these animals is not known. It becomes important, therefore, that methodologies (DNA-xenobiotic adducts or PAC metabolites in tissues and bile) for measuring the level of exposure to PACs need to be validated for use in stranded marine mammals to assess PAC exposure. However, currently these biochemical measurements require relatively fresh tissue samples, thus complicating the task of assessing PAC exposure in stranded animals that have been dead for various periods of time and are exhibiting tissue autolysis.
Data on concentrations of a number of elements in liver, kidney, and stomach contents were also statistically evaluated (Table 10) to determine if there were significant differences between the concentrations for animals from Puget Sound in comparison to the values for whales from the Washington outer coast, Strait of Juan de Fuca and Strait of Georgia treated as a single group. No significant differences attributable to region of stranding were observed for any of these elements. These findings demonstrate that neither longer term exposure as reflected by concentrations in liver and kidney, nor short-term exposure, as reflected by concentrations of these elements in stomach contents, showed any regional specificity for whales that were stranded in Puget Sound as compared to those that stranded at other sites in Washington waters. Additionally, the finding of no significant stranding site related difference in concentrations of aluminum is consistent with a geological rather than anthropogenic origin of aluminum in stomach contents and tissues of gray whales, as discussed above.
A further assessment of whether whales spending time in more urban areas is related to tissue levels of chemical contaminants could be assessed for a subgroup of whales stranded in Puget Sound (Table 11). These four whales (CRC 334, CRC 397, CRC 398, CRC 401) were observed in Puget Sound from 33 to 67 days before they died, and all were observed exhibiting foraging behavior one or more times. One-way ANOVA of the concentrations of CHs in blubber of these four whales showed no significant difference in concentrations of CHs from the concentrations in gray whales stranded on the Washington outer coast or in the Strait of Juan de Fuca. These findings further support the conclusion that the concentrations of CHs in tissues of gray whales do not show a clear association to the environmental chemistry of areas in proximity to the stranding sites.
Although the benthic feeding strategy of gray whales that is
unique among the baleen whales would suggest them as a sentinel
species of environmental quality in specific regions, the present
findings, the whales' extensive migrations, and their variable
periods of fasting are factors that appear to limit their use
in this regard. Recent studies on the biology of gray whales
in Puget Sound have revealed that some individuals return to feed
for extended periods in the same areas (Calambokidis et al. 1992,
1993). In 1992 for example, four gray whales that were observed
feeding in Puget Sound were individuals that were returning for
at least their second or third consecutive year (Calambokidis
et al. 1993). So far none of the animals that has died and been
examined has been any of the individuals known to have returned
multiple years to Puget Sound. The relatively territorial marine
mammals, such as harbor porpoises or harbor seals (Phoca vitulina),
may be more appropriate as sentinel organisms in assessing regional
differences in marine environmental quality. For example, a recent
study (Calambokidis and Barlow 1991) with harbor porpoises showed
that the ratios of the concentrations of PCBs to DDEs in blubber
were significantly different among porpoises from southern California
(0.36 ± 0.03), Oregon (0.53 ± 0.06), and Washington
(1.22 ± 0.17). In southern California, the levels of DDE
in sediment are higher than in sediments from sites in Washington
or Oregon (Varanasi et al. 1989a). In the present study, the
mean PCB/DDE ratio in blubber of gray whales from Washington waters
(6.4 ± 0.6) was very similar to that for gray whales from
San Francisco Bay (5.4 ± 0.99), while the ratios in blubber
of the two gray whales from Alaska were 6.0 and 16.7, thus indicating
no marked regional differences. Further, the PCB/DDE ratio in
gray whale blubber was markedly higher than in harbor porpoise
from the Washington coast, indicating relatively lower concentrations
of DDE in gray whales than in porpoise.
Statistical analyses of concentrations of CHs in blubber, which was the only tissue available for all 22 animals, revealed no marked consistent difference between sexes in the concentration of persistent CH contaminants. The results showed significant differences (P < 0.05) between males and females only for mirex, and for the summed concentrations of DDEs and DDTs (Table 9). Findings from other studies (Cockcroft et al. 1989, Aguilar and Borrell 1988) suggest that there is enhanced excretion in sexually mature females of lipophilic CHs as a result of the redistribution of CHs that occurs during pregnancy and lactation. However, in the present study, tissues were obtained from only six females and only two were greater than 1,000 cm in length; the average length at sexual maturity of female gray whales is 1,200 cm (Rice and Wolman 1971). Thus, additional tissue samples from adult gray whales are needed before sex differences in concentrations of CHs can be fully evaluated.
There are 209 possible PCBs that differ in the degree and position of substitution of chlorines on the biphenyl ring structure, and both factors are important in the toxicity of individual congeners. The PCB congeners (e.g., 3,4,3',4'-tetrachlorobiphenyl, 3,4,5,3',4',5'-hexachlorobiphenyl) that are isostereomers to the highly toxic 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) are particularly hepato- and immunotoxic (Safe 1990), while unidentified and more highly chlorinated congeners present in commercial PCB mixtures, such as Aroclor 1260, appear to be more carcinogenic (Kimura and Baba 1973). The profile of PCB congeners in tissues is, therefore, important in evaluating the toxic potential of PCBs. In the present study, the pattern of accumulation of PCBs by congener class (tri- to nonachloro) in stomach contents, liver, and blubber was examined. Because sex-specific differences in concentrations of some CHs were observed in this study and may be attributed to cyclic depletion of these lipophilic contaminants from females, we first examined profiles of PCBs by congener class in blubber of males and females (Fig. 2); no substantial differences among males and females were observed in these profiles (Fig. 2). Comparison of profiles of PCB congener classes among blubber, liver, and stomach contents was conducted for males only (Fig. 3), for which a larger number of samples was available. The results showed that in stomach contents of males, tetrachlorobiphenyls were present in the highest proportion (40 ± 3.5%). The proportion of trichlorobiphenyl was quite low and the proportions of penta- to nonachlorobiphenyls decreased with increasing degree of chlorination. However, in gray whale liver there was a shift towards PCB congeners with a higher degree of chlorination when compared to the pattern in stomach contents. Specifically, the proportion of tetrachlorobiphenyls was slightly lower and the proportion of heptachlorobiphenyls was slightly higher in liver compared to stomach contents, although relative ranking of the proportion of a congener class was similar in stomach contents and the liver (Fig. 3). Examination of PCB congener profiles in blubber compared to the profile for stomach contents showed a clearer trend towards preferential accumulation of PCBs with a higher degree of chlorination, as evidenced by a substantial increase in proportions of the more hydrophobic penta- and hexachlorobiphenyls, confirming that blubber is a storage depot of PCBs in gray whales and may reflect relatively chronic exposure, whereas liver shows a pattern reflecting more recent exposure. This finding is in concordance with the hypothesis that biphenyls with a lower degree of chlorination are more subject to oxidative metabolism and more readily excreted (Safe 1984, Birnbaum 1985, McFarland and Clarke 1989). Thus, the congeners retained in an organism chronically exposed to PCBs would exhibit a pattern of more highly chlorinated classes as compared to those in sediment, the main repository for hydrophobic contaminants in aquatic systems.
It should be noted that the blubber of gray whales shows PCB profiles with a higher proportion of less chlorinated congeners than profiles in blubber from fish-eating cetaceans and pinnipeds, or liver of benthic fish from urban areas. Specifically, in our analyses of blubber of cetaceans such as bottlenose dolphin from the Gulf of Mexico coast and harbor porpoise from the Atlantic coast or liver of English sole from Puget Sound, the predominant congeners are hexachlorobiphenyls (37 to 42%), followed by penta- and hepta- chlorobiphenyls, which are followed by lower proportions of tetra- and octa- chlorobiphenyls, with trichlorobiphenyls present at the lowest proportion (Tilbury, unpubl. data). The pattern in gray whale blubber showed that both tetra- and penta-chlorobiphenyls were present in higher or comparable proportions to those of the hexachlorobiphenyls. It appears, therefore, that differences in feeding strategy and possibly differential metabolic capacity of gray whales in comparison to other cetaceans and pinnipeds leads to profiles of PCBs in blubber exhibiting higher proportions of less chlorinated PCB congeners.
Metabolism is an important factor for excretion of highly lipophilic compounds, and the highly substituted congeners tend to be resistant to metabolism and subsequent excretion. The route of excretion shifts from urine to feces with increasing size and number of halogen atoms. Laboratory studies show that clearance of PCBs by invertebrates, fish, and terrestrial mammals occurs primarily through excretion into the digestive tract, followed by elimination with fecal material (Waid 1986). Therefore, bile in fish and mammals may be an important fluid in which to measure levels and profiles of CHs to understand further the mechanisms of excretion of these compounds. It would also be important to examine profiles of PCBs in other tissues such as reproductive organs, kidney, and brain to further understand the accumulation in extrahepatic tissues. In the present study, the pattern of PCBs in brain tissue (not shown) more closely resembled that in the blubber than the profile in liver; however, because only one complete set of tissue samples was available, no firm conclusions can be drawn with regard to the pattern of deposition of PCBs in this extrahepatic tissue.
The present findings for stranded gray whales sampled from several areas on the West Coast of the United States showed that the concentrations of a broad spectrum of anthropogenic contaminants were relatively low in all of the whales analyzed. Further, the results suggest that elevated concentrations of aluminum detected in stomach contents appeared to come from geologic and not anthropogenic sources. Moreover, the low tissue concentrations (wet weight) of toxic elements (e.g., mercury 56 ± 12 ng/g in liver) and CHs (e.g., summed concentrations of PCBs, 590 ± 140 ng/g in liver and 1,600 ± 450 ng/g in blubber; summed concentrations of DDEs, 100 ± 28 ng/g in liver and 310 ± 96 ng/g in blubber) would suggest that toxic chemical induced effects did not significantly contribute to the death of these whales. Results from studies with other marine and terrestrial mammals, as summarized by Wagemann and Muir (1984), indicate that concentrations (wet weight) of mercury in liver and summed concentrations of PCBs in blubber exceeding 100,000 ng/g and 50,000 ng/g, respectively, are levels of toxicological concern. However, the lack of samples from apparently healthy gray whales limits further assessment of the role of these or other anthropogenic factors in the mortality of the whales. Additionally, natural factors such as the increase in recent years in the population of these protected species (Rice et al. 1984) must also be considered when evaluating incidents of stranding or sightings in near coastal waters.
We are grateful to Dr. Nancy Foster of the Office of Protected Resources, National Marine Fisheries Service, NOAA, who provided valuable assistance and funding support in organizing this project. We appreciate the additional assistance of John Cook (National Marine Fisheries Service, Kodiak, Alaska), Richard Ferrero (National Marine Mammal Laboratory, Seattle, Washington), Tag Gornall (Marine Animal Resource Center, Seattle, Washington), and Robert Jones (Museum of Vertebrate Zoology, Berkeley, California) for the collection of samples. A number of EC Division scientists and technicians ably assisted in the collection of samples, sample analyses, and data management. In alphabetical order they are Jill Andrews, Stacie Aspey, Lynn Berggren, Kristin Bryant, Jennie Bolton, Daryle Boyd, Richard Boyer, Todd Crawford, Katherine Dana, Tho Dang, Heather Day, Donald Ernest, Joy Evered, Tara Felix-Slinn, Rebecca Hastings, Stephanie Johansen, Ronald Modjeski, John Shields, David Rees, Ronald Pearce, Susan Pierce, and Dana Whitney. We also thank Dr. Bruce McCain for reviewing the manuscript and Mark Myers for reviewing the manuscript and histopathological examination of blubber tissue. Sharon Giese and Barbara Bennett for editorial comments and assistance, and Shirley Perry for typing the manuscript.