i1, il/^ ' 9 EXPOSURE OF MARINE BIRDS TO ENVIRONMENTAL POLLUTANTS SKIOIWORF^ ^y SAR/UO0A h?mu, N, Y, ^^Ov^XVV^"^ UNITED STATES DEPARTMENT OF THE INTERIOR FISH AND WILDLIFE SERVICE Wildlife Research Report 9 WILDLIFE RESEARCH REPORTS This series comprises reports of research relating to birds, mammals, and other wildlife and their ecology, and specialized bibliographies on these, issued for wildHfe research and management specialists. The Service distributes these reports to official agencies, to libraries, and to researchers in fields related to the Service's work; additional copies may usually be purchased from the Division of Public Documents, U.S. Government Printing Office. Library of Congress Cataloging in Publication Data Ohlendorf, Harry M Exposure of marine birds to environmental pollutants. (Wildlife research report; 9) Bibliography: p. 1. Sea birds— Physiology. 2. Pollution -Environmental aspects. 3. Pollu- tion—Toxicology. 4. Oil spills and wildlife. I. Risebrough, Robert W., joint author. II. Vermeer, Kees, joint author. III. Title. IV. Series. QL698.044 598.2 '2 '4 78-5240 Use of trade names does not imply U.S. Government endorsement of commercial products. EXPOSURE OF MARINE BIRDS TO ENVIRONMENTAL POLLUTANTS By Harry M. Ohlendorf Robert W. Risebrough Kees Vermeer -•^O wittA.^'"' I From the collection of International Bird Rescue Research Center Cordelia, California in association with 7 ^ ^ i— m 0 Prelinger u V t a ibrary San Francisco, California 2006 UNITED STATES DEPARTMENT OF THE INTERIOR FISH AND WILDLIFE SERVICE Wildlife Research Report 9 Washington, D.C. • 1978 Ji Digitized by the Internet Archive in 2007 with funding from IVIicrosoft Corporation http://www.archive.org/details/exposureofmarineOOohlerich Contents Page Abstract 1 Petroleum Hydrocarbons 2 Sources of Oil in United States Waters 3 Transfer and Dissipation of Oil in the Marine Environment 4 Exposure of Marine Birds to Oil 5 Biological Effects of Oil on Marine Birds 5 Feather-oiling 6 Toxicology, physiology, and pathology 7 Reproduction 8 Behavior 9 Organochlorines 9 Exposure of Marine Birds to Organochlorines 10 Biological Effects of Organochlorines on Marine Birds 15 Toxicology, physiology, and pathology 15 Reproduction 18 Behavior 22 Heavy Metals 22 Exposure of Marine Birds to Heavy Metals 23 Biological Effects of Heavy Metals on Marine Birds 25 Toxicology, physiology, and pathology 25 Reproduction 27 Behavior 29 Plastic and Other Artifacts 29 Recommendations 30 Acknowledgments 31 References 31 Exposure of Marine Birds to Environmental Pollutants^ by Harry M. Ohlendorf U.S. Fish and Wildlife Service Patuxent Wildlife Research Center Laurel, Maryland 20811 Robert W. Risebrough and Kees Vermeer Canadian Wildlife Service Department of the Environment Ottawa, Ontario (RWR) Delta, British Columbia (KV) Abstract It is unlikely that any marine birds remain uncontaminated by the synthetic organochlorine compounds that have become ubiquitous pollutants. Marine birds also are increasingly exposed to petroleum compounds as a result of the exploitation of undersea petroleum deposits, increased tanker traffic, and ex- pansion of coastal petrochemical industries. Lethal and reproductive effects of organochlorines on marine birds have been most pronounced in coastal areas receiving effluents discharged by manufacturing plants. For example, particularly severe DDT contamination in southern California and elevated levels of dieldrin and related chemicals in the Netherlands have killed local marine birds or inhibited their reproduction. Eggshell thinning, apparently resulting from exposure to DDE, is widespread among estuarine species, and eggshells of peregrine falcons {Falco peregrinus) have become thinner in all areas of the species' range thus far studied. In more contaminated coastal areas, reproductive success of the peregrine falcon is low. Adverse effects of organochlorines on the reproduction of other species also have been found. The oihng of feathers and the associated mortality of marine birds are not the only adverse effects of oil pollution; ingestion of oil may cause death by dehydration by interfering with ion transport and water balance in the gut. Surfactants used to disperse oil spills also have serious consequences for ma- rine birds. Dissolved oil fractions may kill or poison the biota the birds feed on. The physiological effects of the incorporation of more persistent com- pounds into marine food webs are unknown. Contamination of marine birds by most metals and certain trace elements has not been documented, although elevated mercury levels have been ob- served in birds of certain estuarine and local marine environments. The signifi- cance of elevated mercury levels and small plastic particles found in the stomachs and pellets of marine birds is not yet known. ' A summary of this paper was presented at the 13-15 May 1975 international symposium on "Conser- vation of Marine Birds of Northern North America," in Seattle, Washington. The paper was written in 1975-76, and certain portions have been amended or updated as references were pubHshed. Proceed- ings of the symposium are being prepared for publication, but this paper on environmental pollutants is being published separately because of its great length and the delay in pubhcation of the entire Proceedings. Marine birds are exposed to several types of environmental pollutants: petroleum hydro- carbons, organochlorines, heavy metals, and others. Because few data are available for northern North America, potential problems for marine birds there must be judged from observations in other geographical areas. Certain marine birds may serve as indi- cators of environmental pollution on a global scale because (1) they usually can be identified even in an advanced state of decomposition, after a long period of submergence in sea- water, or when thickly covered with oil; (2) they are geographically widespread, often are very numerous, and feed on a wide range of marine organisms; and (3) most species nest colonially and lay large, distinctively marked eggs that are often easily collected and consti- tute distinct units for comparison between species (Vermeer and Reynolds 1970; Prestt 1971). Eggs serve as particularly useful sample units for analysis of organochlorines and certain heavy metals, particularly mercury, because they do not decompose rapidly and are easily handled. Some marine bird species lay additional clutches if the first is removed; therefore, eggs may be taken without severe adverse effects on populations. This charac- teristic is of particular importance because studies are sometimes not begun until it is ap- parent that a population is declining (Prestt 1971). Organochlorine concentrations in the egg are about equal to whole body concentrations found in the female at the time the egg was laid (Keith and Gruchy 1972). Although some microbes have the ability to metabolize or- ganochlorine pesticides under certain condi- tions (Matsumura 1974), putrefaction does not significantly affect residue analysis for DDT and its metabolites (Mulhern and Reichel 1970). During incubation, however, the developing embryo appears to metabolize DDT to DDD and DDE (Abou-Donia and Menzel 1968; Blus et al. 1974b). Chemical resi- due concentrations can be adjusted for the loss of moisture and lipids that occurs during incubation (L. F. Stickel et al. 1973). Eggs may not be the best tissue for mea- surement of all metals, because certain heavy metals apparently are not readily transferred to them (Anderlini et al. 1972). This, however, is not true of mercury. Under certain circum- stances, feathers may be the best tissue to analyze for mercury residues (Stickel 1971). However, unlike liver and muscle tissue, mer- cury residues in feathers tend to reflect body burdens at the time the feathers were grow- ing. The liver, which is a major organ of me- tabolism, or muscle appear to be the best tis- sues for measuring current exposure to mer- cury (Backstrom 1969; Vermeer and Arm- strong 1972b; Fimreite 1974). Other heavy metals may be concentrated in other tissues. For example, residues in the bones, kidneys, and brain, as well as in the liver, appear to be the best measure of exposure to lead (Long- core et al. 1974b). The transfer of petroleum hydrocarbons to eggs has not been reported, but may be expected to occur. Unless otherwise indicated, all chemical residues in the present report are expressed on a wet-weight basis. Petroleum Hydrocarbons Because much of the current information concerning the significance of oil pollution in the estuarine and marine environment has been included in recent reviews (National Academy of Sciences 1973, 1975a; Moore and Dwyer 1974; Evans and Rice 1974; Vermeer and Vermeer 1974a, 1974b; Farrington 1977), we have avoided an extensive review here. However, some of the general information, taken in part from these reviews, is pertinent to our subject and has been included along with that more specifically related to birds. Crude oil and petroleum products are com- plex mixtures of chemicals with individual compounds numbering in the tens of thou- sands with wide molecular weight ranges (Farrington 1977). No one method of analysis is available that will provide reUable esti- mates of the concentration of the entire range of petroleum compounds, and there has yet to be a complete analysis of a single crude oil. Therefore, reports of the presence or absence of petroleum pollution should be carefully evaluated to be certain that the methods of chemical analysis employed would indeed pro- vide the information reported. Vanadium and nickel are present in appre- ciable quantities (> 100 ppm) as organometal- lics indigenous to crude oil, and other trace metals are picked up during production or transportation of crude oil (Whisman and Cotton 1971). Oil pollutants have been detected in sedi- ments, water, and organisms in areas of large oil spills as well as from areas where no large spills have occurred in past months or years (Farrington 1977). These areas are near sources of small spills and chronic inputs. No more than an estimated 300 analyses for pe- troleum pollutants in sediment, water, and or- ganism samples have been reported in the lit- erature exclusive of reports of visible sheens on the water. The scarcity of published measurements of the extent and severity of oil pollution in sedi- ments and organisms is probably related to the difficulty of making meaningful analytical measurements to detect petroleum pollution (Goldberg 1972). Sources of Oil in United States Waters The amount of oil entering the marine envi- ronment from known sources has recently been estimated on a worldwide basis (5.3 million metric tons) as well as for the United States (1.3 million metric tons; Na- tional Academy of Sciences 1973, 1975a; Far- rington 1977). The largest amounts of oil come from normal transport and refining operations and are intentional discharges (Farrington 1977). Accidents account for only 3% of the oil that reaches marine waters of the United States and for only 5% of the world total. Oil tdnker operations account for 26 times as much oil as offshore production in the United States and 24 times as much in the world total. The oil that reaches the coastal waters from rivers and from land operations accounts for 65% of the total (Farrington 1977). The oil that reaches the oceans from the air, by dry fallout and rain, is estimated to be less than 5% of the total. The relative importance of the various sources of oil entering the marine environ- ment varies with location and time (Farring- ton 1977). For example, a large well blowout would introduce a massive amount of oil to a given location and even if averaged over a 10- year period would be the dominant source for that geographical location. The effect of the oil from the various sources can be very different (Farrington 1977). For example, accidental spills may have both immediate acute effects and long- term chronic effects. Municipal or industrial effluents, in contrast, may have no measur- able immediate impact but may have long- term chronic effects as the concentration of the petroleum chemicals builds up in the ecosystem. Two important points relate to control of oil pollution discharges (Farrington 1977): (1) The largest source is the chronic drib- bling of oil into the coastal zone by industrial and municipal effluents, urban runoff, and river runoff carrying oil from inland areas. A substantial amount of oil, therefore, will be discharged to the coastal zone regardless of source. This amount will increase as oil con- sumption increases unless control steps are taken. Evidence suggests that chronic low- level pollution could be potentially more dam- aging to ecosystems than isolated cata- strophic spills (Evans and Rice 1974). (2) It may be safer for the total marine envi- ronment to drill and produce oil in offshore areas than to import equal quantities of oil. Approximately 0.014% of the oil produced offshore is discharged to the marine environ- ment, in contrast to about 0.16% of the oil transported by tanker. However, this does not take into account the ecological damage that may occur in coastal areas as a result of the construction and maintenance of pipelines and onshore facilities. Mystery oil spills, those of unknown source, account for 30% of the oil spills in U.S. waters (National Academy of Sciences 1973, 1975a). There are two possible ways to identify mys- tery oil spills. The first method is to tag oil tanker cargoes, pipeline loads, and storage tank contents with microscopic spheres or special chemicals. However, the size of the bu- reaucracy necessary to ensure accurate rec- ords renders this method impractical (Far- rington 1977). The second method is to make detailed chemical analysis of the spilled oils and poten- tial sources. The chemical characteristics are then compared and the best match of a poten- tial source with the spilled oil is attempted. This technique, which is called "passive tagging," makes use of the unique chemical composition of each oil to distinguish one from another and to match oils from source and spill samples. The technique is also re- ferred to as "fingerprinting," which is perhaps unfortunate. Many nonscientists in the field of oil pollution control have mistak- enly equated "fingerprinting" in identifying mystery oil spill sources with fingerprinting in criminology. Although there are some simi- larities, the identification of oils is very diffi- cult and in its infancy as a technique (Lee et al. 1974; Farrington 1977). However, follow- ing an extensive investigation by the U.S. Coast Guard, charges have been filed in the first case that was based on chemical similari- ties in spilled oil and a sample taken from a suspect tanker (Anon. 1975). Transfer and Dissipation of Oil in the Marine Environment A basic understanding of the various path- ways of transfer and fate of oil has been de- rived from laboratory studies, field studies, and the application of knowledge of processes in the marine environment (Farrington 1977). Many of the processes that act on the oil re- sult in a fractionation and selective removal of certain components more rapidly than others (Farrington 1977). Lower molecular weight components of the type found in kerosene, gasoline, and in varying concentrations in crude oils and fuel oils will evaporate more rapidly than the heavier molecular weight components such as those that make up the bulk of lubricating oils. The lower molecular weight components also are more soluble than the heavier components (Moore and Dwyer 1974; Farrington 1977). When oil is placed in contact with seawater, the lower molecular weight aromatic hydrocarbons are dissolved or accommodated in the water to a greater extent than are the saturated hydrocarbon components (Boy Ian and Tripp 1971; Frank- enfeld 1973; Boehm and Quinn 1974; Ander- son et al. 1974a, 1974b; Lee et al. 1974; American Petroleum Institute 1973). When a spill occurs, however, oil may enter marine sediments and be released essentially un- changed months later (Blumer et al. 1970). Extensive laboratory research has been di- rected toward a better understanding of the biodegradation of oil, and of the individual compounds or classes of compounds in oil (Davis 1967; ZoBell 1969; Ahearn and Meyers 1973; National Academy of Sciences 1973, 1975a). Several species of microorganisms, e.g., bacteria and yeasts, will completely de- grade certain components of oil under the right conditions in the laboratory or in the field (Farrington 1977). Bacteria capable of partially degrading oil have been isolated from several locations in the world's oceans (Farrington 1977). How- ever, the rates of degradation in the various types of coastal areas are unknown. The po- tential pathogenicity of some species of bac- teria that might increase in number near or in an oil spill area also is unknown and little is known about the toxicity of the chemicals produced by microbial degradation of oil (Na- tional Academy of Sciences 1975a). Knowl- edge of the biochemical pathways and prod- ucts of the biochemical degradation of oil is only rudimentary (Davis 1967; ZoBell 1969; Ahearn and Meyers 1973; National Academy of Sciences 1973, 1975a; Farrington 1977). Oil may enter marine organisms by inges- tion of contaminated food and may also enter from water across membrane surfaces such as gills (Farrington 1977). Oil incorporation into some shellfish, lob- sters, and fish is reversible to some extent when the animals are placed in clean water for a period of time. Most, but not all, of the oil taken up from water by the animals was dis- charged within weeks to months in clean water (Blumer et al. 1970; Lee et al. 1972a, 1972b; Anderson 1973; National Academy of Sciences 1973; Stegeman and Teal 1973; An- derson et al. 1974b; Fossato 1975). However, oysters exposed for 2 months to oil from an oil spill did not appreciably reduce their oil content even after 6 months in cleaner waters (Blumer 1971; Blumer and Sass 1972). The more toxic cyclic hydrocarbons were retained longer than the less toxic straight chain com- pounds (Blumer et al. 1970). Fish tested in the laboratory partially me- tabolized several different aromatic hydrocar- bons of the type found in crude oils and fuel oils (Lee et al. 1972b). Mussels, however, did not metabolize these compounds under simi- lar conditions, showing the undesirability of extrapolating from one group of organisms to another (Lee et al. 1972a). Equal caution is ad- visable in extrapolating from results of tests of those few compounds that have been tested because differences in the molecular structure can have profound effects on the rates at which they are absorbed and metabohzed (Farrington 1977). Aside from these few examples, we have found no studies of retention of petroleum hy- drocarbons after oil spills. Neither have we found studies of the uptake, retention, and discharge of petroleum hydrocarbons taken in with food. Some data suggest that food web magnification of oil does not occur in certain communities of marine organisms (Lee et al. 1972a, 1972b; Burns and Teal 1971, 1973; An- derson 1973; Stegeman and Teal 1973). There may, however, be magnification of the higher boiling fractions of the contaminants higher up in the food web (Burns and Teal 1971). Chemical communication is highly im- portant among marine organisms, for both interspecific and intraspecific message sys- tems. Because very low concentrations of or- ganic stimuli are required for communication, such processes are especially susceptible to in- terference by pollutants at low concentrations (Blumer 1971; Blumer et al. 1973; Jacobson and Boylan 1973; Atema and Stein 1972, 1974). Small quantities of crude oil (0.9 ml in 100 liters of sea water) interfere with some specific, possibly chemosensory, behavior of the lobster (Homarus americanus). The delay period between noticing food and going after it doubled when oil was added. The water-sol- uble fraction of the oil alone (in the 50-ppb range) did not have a noticeable effect on be- havior and feeding times. Morphological changes in odor receptors after oil exposure were not detected by light and electron mi- croscopy. The results indicate that small amounts of oil mixed in seawater constitute a bad odor in the lobster's environment, de- pressing its appetite and chemical excitability (Blumer etal. 1973). Exposure of Marine Birds to Oil Following the 1969 spill of 650,000 to 700,000 liters of No. 2 fuel oil into Buzzards Bay and the adjacent Wild Harbor Marsh near West Falmouth, Massachusetts, essen- tially all the marsh organisms living in the contaminated area were affected; they all ac- cumulated oil hydrocarbons in their tissues. Two processes apparently occur as the oil passes through the marsh ecosystem: a pro- gressive loss in the straight chain hydrocar- bons in relation to the branched chain, cyclic, and aromatic hydrocarbons, and a greater re- tention of the higher-boiling fractions of the contaminants by organisms higher in the food chain (Burns and Teal 1971). Although its feathers were not oiled, a herring gull {Larus argentatus) that was col- lected in Wild Harbor had substantial amounts (584 ppm) of the whole spectrum of fuel oil hydrocarbons in its muscle but con- tained mostly those with straight and slightly branched chains. The brain of this bird also contained high residues (535 ppm), but with a higher proportion of the aromatic compounds. Another herring gull, collected outside the spill area, had much lower oil hydrocarbon residues in its muscle (10 ppm) and brain (15 ppm), and the aromatic compounds were not detected in the brain (Burns and Teal 1971). Three birds that died in the 1971 San Fran- cisco Bay oil spill contained very high resi- dues of oil hydrocarbons in their tissues. A composite sample of liver, kidney, brain, fat, and heart tissue of a common murre (Uria aalge) contained 8,820 ppm, composite liver and kidney tissue of a surf scoter (Melanitta perspicillata) contained 1,250 ppm, and the liver of a western grebe (Aechmophorus occi- dentalis) contained 9,100 ppm oil hydrocar- bons. The composite tissues (liver, kidney, brain, fat, and muscle) of a murre that had not been exposed to the oil spill contained no de- tectable oil hydrocarbons (Snyder et al. 1973). Body fat of herring gulls breeding on Lake Ontario in 1973 contained a number of aro- matic hydrocarbons including several polynu- clear aromatics. Naphthalene, 2-methyl-naph- thalene, acetonaphthalene, and biphenyl were identified from their retention times (Fox et al. 1975). The sources of these aromatic hydro- carbons remain undetermined. Accumulation in the food chain from water or sediments through fish is probable, but these com- pounds, which presumably are of petroleum origin, may have been ingested at garbage dumps. Thus, it appears likely that aquatic birds living in oil-polluted environments may accumulate residues of the relatively more persistent compounds. Biological Effects of Oil on Marine Birds Aside from the reports on mortality and re- habilitation of oiled birds, the biological ef- fects of oil on marine birds are little known. Important biological effects include both acute and chronic toxicity as well as adverse effects on physiology, reproduction, and be- havior. Indirect effects involving the food web and changes in habitat and food supply are relatively unknown. There is a distinct possibility that oil and other environmental contaminants such as the organochlorine compounds may act syner- gistically (Farrington 1977). Circumstantial evidence suggests that oil pollution has seriously reduced populations of certain species of marine birds in some areas (Tuck and Livingston 1959; Tuck 1960; Hawkes 1961; Buck and Harrison 1967; Par- slow 1967, 1970; Bourne 1968; Clark 1973). An oil spill can have significant effects on populations of marine birds such as the alcids, which often are numerous among the birds that die in spills. Although alcids are long- lived and have few predators once they are at sea, they often do not breed until 3 or more years old, most lay only a single egg per clutch, not all adults breed every year, and they produce an average of only one chick per five breeding adults. These species require more than 50 years to double their population under optimal conditions. More than half a century would be required for a colony to re- cover its numbers (excluding immigration) if reduced by one-half as the result of a large oil spill (Clark 1969). The potential effects of oil spills on aquatic birds and their feeding habitat on the Cana- dian west coast were assessed by Vermeer and Vermeer (1975). They concluded that the pres- ent shipping of oil plus the increased tanker traffic along the entire British Columbia coast that is expected to be in progress in 1977 will result in enough oil spillage to threaten the coastal populations of seabirds with destruc- tion. Concentrations of seabirds will be most vul- nerable to spills (Vermeer and Vermeer 1975). Three major colonies along the coast of British Columbia are the Langara Region, the southeast coast of the Queen Charlotte Islands, and the Scott Islands. Alcids and storm petrels [Oceanodroma spp.) are the most numerous seabirds along the British Columbia coast. Alcids are among the birds most vulnerable to oil pollution, whereas storm petrels are less threatened by spills be- cause they spend more time in the air and dive only occasionally. Waterfowl, especially div- ing ducks, will be vulnerable to spills during the winter as they concentrate in large numbers in estuaries and inlets along the British Columbia coast. The large wintering populations of ducks, geese, and grebes along the Fraser Delta foreshore and Boundary Bay will be vulnerable because of their near- ness to tanker and shipping traffic. Approxi- mately 1 million loons, shearwaters, phal- aropes, ducks, gulls, and geese migrate north in the spring along west Vancouver Island. These migrants, because of their concentra- tion in large numbers, may be very tempo- rarily but critically vulnerable to oil pollution. The birds most likely to be directly affected by spills are breeding populations of alcids and wintering diving ducks, whereas ducks, geese, and shorebirds, which feed in the inter- tidal zone, may be hardest hit indirectly through destruction of their feeding habitat (Vermeer and Vermeer 1975). Of the ducks threatened by destruction of their feeding habitat, sea ducks are most vulnerable be- cause they rely most on the marine habitat for feeding purposes. Feather-oiling Large numbers of marine birds die each year as a result of oil spills. Estimates of mor- tality are based primarily on beach counts of oiled birds, but such estimates may be highly inaccurate because a significant percentage, perhaps 50-90%, of the dead birds never wash ashore (Clark and Kennedy 1968; Coulson et al. 1968; Tanis and Morzer Bruyns 1968; Hope-Jones et al. 1970). An estimated 30,000 marine birds, of which about 97% were common murres and razor- bills (Alca tarda), died as a result of the Torrey Canyon disaster (Bourne et al. 1967). Earlier, in the winter of 1951-52, approximately 100,000 birds were lost to oil pollution on the coasts of the British Isles (ZoBell 1962). At least 10,000 birds, including alcids, ducks, gulls, and kittiwakes, were killed by oil appar- ently derived from ballast pumped from tankers entering Cook Inlet, Alaska, during February and March 1970 (U.S. Department of the Interior 1970). The population decline of murres {Uria spp.) along the coast of Newfoundland has been as- sociated with oil pollution (Tuck 1960), al- though the effects probably are not related only to those caused by feather-oiling. Numer- ous other instances of mortality related to oil are documented in reviews on this subject (Clark and Kennedy 1968; Aldrich 1970; Ver- meer and Vermeer 1974a). The effects of oiled plumage on marine birds vary with the properties of the oil, degree of contamination, quantity absorbed, environ- mental conditions, and the original condition of the bird. Even a small patch of oil on the feathers may mean that without care the bird will die (Tuck 1960; Smith 1975), but in some instances it appears that birds are able to clean their own plumage (Phillips 1974; Smith 1975). Oiling of a bird's plumage increases me- tabolism and causes an increased loss of body heat to the surrounding cold water that can readily be fatal (Lincoln 1936; Hartung 1967; Boyle 1969; Greenwood 1970; McEwan and Koelink 1973). Feather-oiling appears to be a more signifi- cant problem in cold-water areas than in areas where water is warmer. Warm water appar- ently causes the spilled liquid oil to form tar- balls that are comparatively less hazardous to birds (Bourne and Bibby 1975). Damage to feathers may result long after exposure and may be reflected by abnormal wear of the plumage (Bourne 1974). After the Torrey Canyon grounding in March 1967, 7,849 oiled birds were captured for cleaning and rehabilitation. One month later, however, fewer than 6% were still alive (Clark and Kennedy 1971). An estimated 3,180,000 hters of bunker C fuel oil were spilled in the massive 1971 oil spill that occurred near the entrance to San Francisco Bay. The California Department of Fish and Game estimated that 7,000 aquatic birds were exposed to the fuel oil, and more than 4,000 of these were taken into captivity for treatment and rehabilitation. Two weeks after the spill, 90% of the birds had died in spite of efforts to save them, and within 3 months mortality exceeded 96% (Orr 1971; Snyder et al. 1973). Grebes, murres, and loons apparently died more rapidly than the other species affected, and ducks appeared most hardy (Snyder et al. 1973). Progress has since been made in the reha- bilitation of oiled birds, and modified methods are being used (Hay 1975). In 1973, the Inter- national Bird Rescue Research Center treated 523 oiled birds of which 49% survived (Smith 1975). Toxicology, Physiology, and Pathology The great diversity of chemical compounds in oil increases the difficulty of determining its toxicological and physiological effects. In addition, oil dispersants used to clean up a spill area are also toxic and the toxicity of oil plus dispersant usually is greater than the toxicity of either alone (Clark and Kennedy 1968; Tarzwell 1970; Linden 1975). There also are important species differences in suscep- tibility (Swedmark et al. 1973). The toxicity of some oils to ducks has been measured under different environmental con- ditions. Single doses of several industrial oils produced lipid pneumonia, gastrointestinal ir- ritiation, fatty livers, and adrenal cortical hyperplasia. Birds that received a cutting oil in combination with diesel oil exhibited acinar atrophy of the pancreas. Those that received diesel oil and a fuel oil developed toxic nephro- sis. Cholinesterase activity was significantly inhibited by administration of the cutting oil and somewhat depressed by the diesel oil (Hartung and Hunt 1966). Ducks that had been killed by oil pollution exhibited changes that were similar to those in the experimentally fed birds, suggesting that toxicity of oils is a major factor in mor- tality of exposed birds (Hartung and Hunt 1966). Toxicity apparently is reduced through aging of the oil because the more volatile com- pounds are also the more toxic (Clark and Kennedy 1968). Birds that died after the San Francisco Bay oil spill in 1971 were examined for pathologi- cal changes that might have resulted from ex- posure to oil. Intoxication from oil ingestion appeared to be an important factor contribut- ing to the high mortality, although the evi- dence was circumstantial. Birds that died in the period of high mortality had ingested oil and exhibited dehydration, ulceration of the intestinal mucosa, enteritis, hepatic fatty changes, and renal tubular nephrosis (Snyder et al. 1973). Similar pathological changes as well as adrenal lesions and pulmonary dis- eases have been observed in other oiled sea- birds (Guillon 1967; Beer 1968). Following the large 1974 oil spill in the Straits of Magellan, a high percentage of South American tern (Sterna hirundinacea) chicks on an island in the spill area died (Smithsonian Institution 1974). Although the cause of mortality is unknown, it is possible that the small fish that the terns ate and fed to their young were contaminated with some fraction of the spilled crude oil in concentra- tions that did not harm the adults but were toxic to the young. It is possible, however, that the chicks died of starvation after the adults were killed or were unable to catch enough food for the young. The rates at which water and sodium are transported across the intestinal mucosa in- crease when Pekin ducklings (Anas platyrhyn- chos) are transferred from fresh water to a diet containing hypertonic saline drinking water (Crocker et al. 1974). These rate in- creases seem to be essential for the successful adaptation of ducklings to saline drinking water. Ducklings given a single oral dose of a crude oil (0.2 ml) at the start of maintenance on saline drinking water did not develop the characteristic rate increases. In addition, high mucosal transfer rates that had been de- veloped in ducklings fed saline water for 3 days ceased 24 h after they received crude oil. Although commercial dispersant (5 ppm or 20 ppm) in fresh or saline drinking water had no effect on ducklings, the presence of dis- persed crude oil (12.5-50.0 ppm) in the water prevented the development of high mucosal transfer rates in the ducklings given saline water. A reduction of the mucosal transfer rates in seawater-adapted ducklings, through the action of ingested crude oil, may limit the amount of free water available to the body (Crocker et al. 1974). Although the high mor- tality among oil-contaminated seabirds may be due to a variety of pathological conditions, dehydration resulting from impairment of mucosal transfer mechanisms may be an im- portant factor contributing to their death. Crude oils from eight different geographical locations reduced the rates of sodium and water transfer across the intestinal mucosa of Pekin ducklings to different degrees (Crocker et al. 1975). Administration of Kuwait crude oil caused the greatest degree of inhibition, and North Slope, Alaska, crude oil caused the smallest. Distillation fractions derived from two chemically different crude oils were adminis- tered to ducklings in volumes that corre- sponded to their relative abundance in the crude oil from which they were derived (Crocker et al. 1975). The greatest inhibitory effect on mucosal transfer was not associated with the same distillation fractions from each oil. A highly naphthenic crude oil from the San Joaquin Valley, California, showed the greatest inhibitory activity in the least abun- dant (2%), low boiling point ( < 245 C) fraction. The most abundant (47%), highest boiling point (>482 C) fraction showed the least in- hibitory activity. In contrast, a highly paraf- finic crude oil from Paradox Basin, Utah, showed the greatest inhibitory effect with the highest boiling point fraction and a minimal effect with the lowest boiling point fraction. The relative abundances of these two frac- tions in the Paradox Basin crude oil repre- sented 27 and 28%. Mucosal transfer inhibition by water-sol- uble extracts of San Joaquin Valley and Para- dox Basin crude oils was roughly proportional to the inhibitory potency of the low boiling point fraction of the oil (Crocker et al. 1975). Weathered samples of these oils showed greater effects than corresponding samples of unweathered oils even though most of the low molecular weight material from both oils was either evaporated or soiubilized in the under- lying water during the 36-h weathering period. Reproduction During the nesting season, small amounts of oil on the plumage of birds can have very serious effects on reproduction. The oil com- pounds that are involved, however, are essen- tially unknown and no extensive tests have been reported.^ Oil washed ashore on a small island in West Germany where terns (chiefly Thalasseus sandvicensis and Sterna hirundo) and Euro- pean oystercatchers (Haematopus ostralegus) were nesting. During copulation, many of the adult terns became dorsally smeared with oil from their mate's oiled feet, but no direct losses among adult terns were attributed to the oil. More than 70% of the young terns ^Reproductive effects have, however, been studied since this manuscript was written (see Szaro 1977). were contaminated with oil and many of them were unable to fly. Some of the eggs laid along the high-tide mark failed to hatch after they became contaminated with oil (Rittinghaus 1956). After ingestion of a relatively nontoxic lu- bricating oil (2 g/kg), one mallard {Anas platy- rhynchos) and two Pekin ducks stopped lay- ing for about 2 weeks. Very small quantities of oil coated on mallard eggs reduced their hatchabihty to 21%, compared with 80% for unoiled eggs. Experimentally oiled mallards continued to incubate their clutches, but their eggs failed to hatch (Hartung 1965). In an experimental application to test the effects of 2,4-D and diesel fuel on eggs of ring- necked pheasants {Phasianus colchicus), there was no adverse effect by the 2,4-D on hatch- ability, but apphcation of the diesel fuel re- duced hatchability to zero (Kopischke 1972). Behavior Exposure to oil causes some obvious changes in behavior patterns of birds because they abandon all activities to attempt to clean the oil from their feathers by preening (Smith 1975). There may be other serious but less readily observed direct effects that influence the birds' ability to locate food, to migrate, or to perform other essential activities. As discussed earlier, small amounts of oil in the water cause significant changes in be- havior of certain marine organisms. Modified behavior among any of the numerous species of animals in the food webs may have serious indirect implications for the welfare of marine birds that depend upon them, Organochlorines By 1971, and perhaps earlier, it became un- likely that any bird dependent upon marine food webs anywhere in the world was free of contamination by the synthetic organo- chlorine compounds that have become ubi- quitous pollutants in the global ecosystems (Sladen et al. 1966; Risebrough and Berger 1971; Bogan and Bourne 1972; Bourne and Bogan 1972; Bennington et al. 1975; Rise- brough 1977; Walker 1977; White and Rise- brough 1977). More information is available on the global distribution patterns of organo- chlorines than for other chemicals in marine birds. Several direct biological effects of or- ganochlorines on marine birds are known. Other relevant information is available on the distribution of these pollutants in estuarine, freshwater, and terrestrial ecosystems, as well as their biological effects on other birds. The most abundant synthetic organo- chlorine compound in tissues and eggs of ma- rine birds is frequently p,p '-DDE, a derivative of p,p'-DDT, which is the principal component of the commercial insecticidal mixture (Rise- brough et al. 1968; Jensen et al. 1969; Koeman et al. 1969). Other DDT compounds fre- quently present in marine birds arep,p'-DDD, p,p'-DDT, and o,p'-DDT (Bennington et al. 1975). Polychlorinated biphenyls (PCB's), or chlorobiphenyls, consist of a mixture of com- pounds differing in chlorine content and the position of chlorine atoms on the parent bi- phenyl molecule. Pentachlorobiphenyls and hexachlorobiphenyls usually constitute the majority of the chlorobiphenyls present in marine birds, but trichlorobiphenyls and tet- rachlorobiphenyls are occasionally present (Risebrough and de Lappe 1972; White and Risebrough 1977). A number of other synthetic organochlorine compounds have been detected in marine birds, but almost always at levels substan- tially lower than those of the DDT and PCB compounds. Hexachlorobenzene (HCB), which has been found in tissues of great cormorants (Phalacrocorax carbo), sandwich terns (Thalasseus sandvicensis), and common eiders (Somateria mollissima) from coastal areas of the Netherlands (Koeman and van Genderen 1972; Koeman et al. 1972a), has been consid- ered a potentially hazardous marine pollutant (National Academy of Sciences 1975b). The HCB is used as a fungicide but may enter the marine environment in significant quantities as a component of the tarry waste products from the manufacture of chlorinated hydro- carbons such as perchloroethylene and carbon tetrachloride that are frequently discharged at sea (Environmental Protection Agency 1973). Chlorinated styrenes were identified by gas chromatography/mass spectrometry in tis- sues of common eiders, sandwich terns, and great cormorants from the Netherlands (Ten Noever de Brauw and Koeman 1972) and in 10 tissues of great blue herons {Ardea herodias) from Lake St. Clair, Michigan (Reichel et al. 1977); their source apparently remains un- known. Chlorinated naphthalenes also were identified in tissues of the great cormorant (Koemanetal. 1973). Mirex was measured in eggs of herons and white ibis {Eudocimus albus) from estuarines of the U.S. Atlantic and Gulf coasts (Ohlen- dorf et al. 1974) and also in the blubber of a seal (Phoca vitulina) from the Netherlands (Ten Noever de Brauw et al. 1973). In addition to its use as an insecticide, mirex, under various trade names, is used as a flame retardant. Dieldrin was found in eggs and tissues of several species of marine birds inhabiting coastal waters of Great Britain (Robinson et al. 1967) and of New Zealand, and also in pelagic species such as the sooty shearwater (Puffinus griseus) breeding in sub-Antarctic islands of New Zealand (Bennington et al. 1975). It is accumulated by ospreys (Pandion haliaetus) and bald eagles {Haliaeetus leuco- cephalus) feeding on coastal marine fish of the eastern United States (Mulhern et al. 1970; Belisle et al. 1972; Cromartie et al. 1975; Wie- meyer et al. 1975). Endrin was detected in brown pelicans (Pelecanus occidentalis) from Florida (Schreiber and Risebrough 1972), the Gulf of California (Risebrough et al. 1968), and Louisiana (J. D. Newsom, personal communi- cation). White pelicans (Pelecanus erythro- rhynchos) from Louisiana also contained en- drin (J. D. Newsom, personal communica- tion). Chlordane compounds, principally oxy- chlordane and cis-chlordane, were found in eggs of herons from the eastern U.S. estuaries (Ohlendorf et al. 1974) and in fish and common terns {Sterna hirundo) from Long Island Sound (R. W. Risebrough and P. Robinson, unpublished data). Heptachlor epoxide, toxaphene, and the several isomers of hexachlorocyclohexane (benzene hexachloride or BHC) are occa- sionally found in estuarine environments. Heptachlor epoxide and BHC isomers have been reported in Antarctic birds breeding in the South Orkneys (Tatton and Ruzicka 1967) but their identification has not been con- firmed (Risebrough 1977). The occurrence, distribution, and effects of organochlorines on wildlife, principally terres- trial, freshwater, and estuarine species, have been summarized in other recent reviews (Prestt and Ratcliffe 1972; L. F. Stickel 1972, 1973; Blus et al. 1977b; W. H. Stickel 1975; Ketchum et al. 1975; L. F. Stickel and F. E. Hester, unpublished manuscript). Earlier studies of the transport of PCB's to the ma- rine environment also have been reviewed (Nisbet and Sarofim 1972; Panel on Hazar- dous Trace Substances 1972). More recently the environmental effects of PCB's (Peakall 1975), their chemical properties (Hutzinger et al. 1974), and the transfer of organochlorine compounds to the marine environment and their incorporation into marine food webs have been reviewed (Risebrough et al. 1976a). Although levels of organochlorine com- pounds other than those of the DDT and PCB groups may occasionally be present at levels deleterious to birds in estuaries, levels in the offshore marine environment are usually well below those considered harmful to marine birds. In the present review, principal em- phasis will therefore be placed on the DDT and PCB compounds. Exposure of Marine Birds to Organochlorines Organochlorine residue data are available from coastal regions, but there have been rela- tively few studies of chlorinated hydrocarbon contamination of marine birds in areas that are far from known pollution sources. All eggs of the Adelie penguin {Pygoscelis adeliae) from widely separated localities in the Antarctic (Risebrough and Carmignani 1972; Risebrough 1977), eggs and tissues of birds from the Aleutians (White and Risebrough 1977) and from sub-Antarctic areas of New Zealand (Bennington et al. 1975), and tissues of birds from the eastern North Atlantic (Bourne and Bogan 1972; Bogan and Bourne 1972) contained residues of DDT compounds. All samples also contained detectable levels of chlorobiphenyl compounds, frequently at levels higher than the total concentration of the DDT group. The DDT and chlorobiphenyl compounds also were detected in all samples obtained from remote terrestrial and fresh- water Arctic ecosystems (Risebrough and Berger 1971; Walker 1977). Although the data are few and sample sizes 11 Table l.Mean PCB and DDE residues (ppm lipid weight) in cormorants (Phalacrocorax spp. ) Percent Locality, date Species N Tissue lipid PCB's DDE PCB/DDE Amchitka, 1971* Red-faced cormorant (P. unle) 1 Yolk 20.0 19.0 3.8 5.0 Amchitka, 1974° Red-faced cormorant 1 Pectoral muscle — 21.0 3.5 6.0 Agattu, 1974° Red-faced cormorant 1 Pectoral muscle 3.8 14.0 2.4 6.0 Amchitka, 1974° Pelagic cormorant (P. pelagicus) 1 Pectoral muscle 4.0 8.0 0.8 10.0 Auckland Islands, Auckland Island 4 Egg lipid 100.0 0.3 0.9 0.3 1972b shag Iceland, 1973'^ Shag (P. aristotelis) 10 Egg 5.0 23.0 3.8 6.0 Great cormorant 13 Egg 4.8 10.0 3.0 3.0 Peru, 1969*^ Guanay (P. bougainuillei) 4 Egg lipid 100.0 15.0 12.2 1.2 Southern Cah- Double-crested 7 Egg lipid 100.0 87.0 754.0 0.1 fornia, 1969e cormorant Greenland, 1972* Great cormorant 3 Body fat - 23.0 9.8 2.3 °White and Risebrough 1977. ^Bennington et al. 1975. '^J. A. Sproul et al., unpublished manuscript. '^R. W. Risebrough et al., unpubhshed manuscript. ^Gressetal. 1973. ^Braestrup et al. 1974. are frequently small, a general picture of global marine contamination by DDT and PCB compounds can be presented. Some of the available data on cormorants have been summarized (Table 1). With the exception of the residue values reported for the double- crested cormorant (Phalacrocorax auritus) in southern Cahfornia, where the birds were ex- posed to industrial contamination from the Los Angeles area (Gress et al. 1973), the samples were obtained from areas reasonably remote from point sources of contamination. The DDE residues in the Auckland Island shags (Phalacrocorax carunculatus) were somewhat lower than in cormorants from Amchitka and Agattu at the equivalent lati- tude in the northern hemisphere. However, PCB values in the southern hemisphere birds were lower by 1-2 orders of magnitude. Other data from biocenotic equivalents in the two areas support the conclusion that DDE levels are slightly lower in the southern than in the northern hemisphere but that PCB values are lower by 1-2 orders of magnitude (Bennington et al. 1975; White and Rise- brough 1977). The DDE levels in an egg of a New Zealand falcon (Falco novae seelandiae) were equivalent to those in eggs of peregrines from Amchitka, but PCB levels were an order of magnitude lower. Comparable differences were found between auklets (Aethia pusilla and Cyclorrhynchus psittacula) and the tufted puffin (Lunda cirrhata) of the Aleutians and the diving petrel (Pelecanoides urinatrix) from the Snares Islands of southern New Zealand. The DDE levels in eggs of the guanay were somewhat higher than those from New Zea- land or the Aleutians, suggesting local sources of DDT contamination in Peru (R. W. Risebrough et al., unpublished manuscript). A comparison of the cormorant samples from the Aleutians and the eggs of two cor- morants (Phalacrocorax carbo and P. aris- totelis) breeding in Iceland suggests that levels of DDE and PCB contamination in the two oceanic areas are similar. In five species of fish obtained from Amchitka in 1974, DDE residues ranged from 1 to 5 ppb; PCB resi- 12 dues ranged from 8 to 20 ppb (White and Risebrough 1977). Residues of DDE in seven species of fish obtained from the coastal waters of Iceland in 1973 ranged from 1 to 9 ppb; PCB levels ranged from 8 to 20 ppb (J. A. Sproul et al., unpublished manuscript). On a lipid basis, PCB residues expressed as tri-, tetra-, or penta-chlorobiphenyls ranged from 0.3 to 2 ppm in the Amchitka fish and from 0.2 to 3 ppm in the Icelandic fish. Body fat of great cormorants from Greenland (Braestrup et al. 1974) contained comparable PCB levels and somewhat higher DDE levels than great cormorants from Iceland. A comparison of DDE and PCB residue levels in black-legged kittiwakes {Rissa tridac- tyla), fulmars (Fulmarus glacialis), and thick- billed murres {Uria lomvia) from Amchitka and Iceland suggests somewhat higher levels in the Icelandic birds, although residues are of the same order of magnitude and with com- parable ratios (J. A. Sproul et al., unpubhshed manuscript; White and Risebrough 1977). The differences may reflect a higher level of con- tamination in those areas of the ocean where the Atlantic birds spend the winter months. Earlier data (Risebrough et al. 1968) sug- gested that DDT compounds were more abun- dant than PCB's in Pacific waters. However, many of the samples were from coastal CaU- fornia waters where DDT contamination was particularly severe. Residue levels and PCB:DDE ratios in the breast muscles of Icelandic birds obtained in 1973 were comparable to those in birds ob- tained earlier from areas north of Britain, in- dicating that no decline in residue concentra- tions in birds had occurred over that short interval (Bourne and Bogan 1972; J. A. Sproul et al., unpublished manuscript). From the Pacific, the visceral fat of 7 black- footed albatrosses {Diomedea nigripes) and 22 Laysan albatrosses (Z). immutablis) from Midway Island contained mean DDE levels of 22 and 8 ppm, and mean PCB levels of 14 and 2 ppm (Fisher 1973). These species are re- stricted to the North Pacific, usually about 20° N and feed primarily on squid. In Hawaii, DDE concentrations in four eggs of the dark- rumped petrel (Pterodroma phaeopygia) ranged from 0.07 to 1.14 ppm (0.6-11.5 ppm, lipid weight) (King and Lincer 1973). PCB values were not reported. From the tropical Atlantic, DDE and PCB levels in the breast muscle of 28 adult sooty terns (Sterna fuscata), breeding on the Dry Tortugas, were 2.5 and 7.8 ppm (lipid weight); mean percentage of hpid was 2.6% (P. G. Connors et al., unpublished manuscript). In these areas of the Atlantic and Pacific, comparatively remote from sources of con- tamination, PCB residue concentrations gen- erally exceeded those of the DDT compounds. In the New Zealand (including the sub-Ant- arctic islands) samples, however, DDT resi- dues were frequently present at higher con- centrations than the sum of PCB's (Benning- ton et al. 1975). In the Antarctic, few eggs of the Adelie pen- guin obtained in 1970 or earlier from widely separated locaUties contained PCB's at de- tectable levels (Risebrough et al. 1976a). Maximum amounts of PCB's in eggs of the Adelie penguin obtained from Cape Crozier in October 1967 were less than one-eighteenth of the concentration of DDT compounds (Rise- brough et al. 1968). Subsequent analysis of some of these eggs revealed the presence of PCB. PCB's were detected also in eggs of the Adelie penguin, chinstrap penguin (Pygo- scelis antarctica), and the gentoo penguin (P. papua) obtained in the Antarctic Peninsula in 1975, although at concentrations less than those of the DDT compounds (Risebrough et al. 1976b). The preponderance of DDT com- pounds in the Antarctic, the most remote area receiving chlorinated hydrocarbons from at- mospheric or oceanic transport, apparently reflects the relative use of these two groups of compounds in the southern hemisphere. In coastal areas local conditions usually de- termine the contamination patterns. For example, liquid chemical wastes discharged by insecticide manufacturing plants in CaU- fornia and the Netherlands subsequently en- tered the sea and caused significant organo- chlorine contamination of coastal birds. High DDT concentrations were found in northern anchovies (Engraulis mordax) from Los An- geles Harbor in 1965 (Risebrough et al. 1967). Subsequent investigations documented ex- ceptionally high levels of DDT compounds in the coastal birds, including the brown peli- cans (Risebrough 1972; Anderson et al. 1975) and double-crested cormorants (Gress et al. 1973). When the company began to dispose of its hquid wastes in a sanitary landfill in 1970, input of DDT compounds into the sea began 13 to decline (Carry and Redner 1970; Redner and Payne 1971; D. R. Young et al., unpub- lished manuscript); residues in fish and in the brown pelicans also began to decline (Ander- son etal. 1975). In 1964, sandwich terns and spoonbills (Platalea leucorodia) were found dying on the island of Texel in the Dutch Wadden Sea. The birds were in tremors and convulsions, signs comparable to those found in other birds poisoned by organochlorine insecticides (Koe- man and van Genderen 1965, 1966). Studies of the distribution of chlorinated hydrocarbons in birds, fish, and mussels (Mytilus edulis) from localities along the Dutch and West German coasts and in the eggs of seabirds of Great Britain indicated a point source of con- tamination by dieldrin, endrin, and telodrin. Telodrin, an insecticide not used in Europe at that time, was being manufactured with diel- drin and endrin in a factory near the mouth of the Rhine River. When it was discovered that these residues were coming from the insecti- cide plant, measures were taken to eliminate discharge; residue levels in the local seabirds began to decline and the sharp decrease in population numbers was halted (Koeman et al. 1968). In addition to these two incidences, coastal contamination from local but diffuse sources has resulted in high levels of organochlorines in birds in Japan, North America, and Europe. Levels of PCB in Japanese birds, in- cluding several species of gulls, were compa- rable to those in industrial areas of North America and Europe (Fujiwara 1974). The PCB residues in breast muscle of eight little egrets (Egretta garzetta) that were found dead or dying in Tokyo Bay ranged from 0.3 to 180 ppm (22-1,600 ppm, lipid basis) with a geometric mean of 9 ppm. Residues of PCB in the breast muscle of eight black-tailed gulls (Larus crassirostris) ranged from 3-39 ppm, with a geometric mean of 13 ppm (Doguchi 1973). In western North America comparatively high levels of DDT and PCB contamination were found in common murres (Gress et al. 1971) and the ashy storm petrels (Oceano- droma homochroa) (Coulter and Risebrough 1973) breeding on the Farallon Islands and in great egrets {Casmerodius albus) and great blue herons breeding at a coastal site (Faber et al. 1972) near local sources of pollution. Most eggs of marine birds from the Strait of Georgia contained more PCB's and DDE and had a higher PCB:DDE ratio than did eggs from the west coast of Vancouver Island and from the Queen Charlotte Islands (Table 2). This comparison within a relatively small re- gion (i.e., the Pacific Coast of British Colum- bia) further illustrates the principle that eggs from birds nesting farther at sea are likely to contain lower levels of organochlorines than those nesting nearer the mainland. Average PCB levels in these samples almost always ex- ceeded those of DDE. Fish, crabs, and shellfish were collected from the lower Eraser River, its estuary, and selected areas of Georgia Strait (Albright et al. 1975). Generally, PCB's were present at higher levels than DDE, and greatest concen- trations of these compounds occurred in biota from waters adjacent to the city of Van- couver. With one exception, animals from Georgia Strait and those away from the im- mediate influence of Eraser River water con- tained no detectable levels of chlorinated hydrocarbons. High levels of DDE and PCB in double- crested cormorants from the Bay of Fundy (Zitko and Choi 1972; Zitko et al. 1972) most likely have resulted from past DDT use in New Brunswick and from diffuse sources of PCB's along the eastern North American coastline. Similarly, contamination levels in ospreys (Wiemeyer et al. 1975; Spitzer et al. 1977) in coastal Connecticut, Massachusetts, New York, and New Jersey most likely were derived from local sources of contamination. Bald eagles found sick or dead in the United States during 1966-72 were analyzed for or- ganochlorines (Mulhern et al. 1970; BeHsle et al. 1972; Cromartie et al. 1975); DDE, DDD, dieldrin, and PCB's were detected in most of the 145 eagle carcasses. Eighteen of the eagles contained possibly lethal levels (greater than 4 ppm) of dieldrin. Since 1964 when data were first collected, 8 of the 17 eagles obtained from Maryland, Virginia, South Carohna, and Florida possibly died from dieldrin poisoning. All four specimens from Maryland and Virginia were from the Chesapeake Bay Tidewater area. In December 1973, eight ruddy ducks (Oxyura jamaicensis) killed in an oil spill on the Delaware River (White and Kaiser 1976), contained DDE (1.1-4.5 ppm) and PCB's (2.8- 10 ppm). Levels of DDT and DDD were below 14 Table 2. Mean PCB and DDE residues (ppm lipid weight) in seabird eggs from the Strait of Georgia, the west coast of Vancouver Island, and the Queen Charlotte Islands in British Col- umbia, 1970 (K. Vermeer, unpublished data). Locality Species Percent N lipid PCB's DDE PCB/DDE Strait of Georgia Mandarte Island Mittlenatch Island Vancouver Island Cleland Island Queen Charlotte Islands Skedans Island Lucy Island Northwest Rocks Double-crested cormorant Pelagic cormorant Glaucous-winged gull {Larus glaucescens) Pelagic cormorant Pigeon guillemot {Cepphus columba) Glaucous-winged gull Leach's petrel (Oceanodroma leucorhoa) Pigeon guillemot Tufted puffin Glaucous-winged gull Fork-tailed petrel (O. furcata) Pigeon guillemot Glaucous-winged gull Rhinoceros auklet (Cerorhinca monocerata) Glaucous-winged gull Glaucous-winged gull 3 6.9 207.0 59.0 3.5 10 5.3 50.0 15.0 3.3 10 6.0 41.5 12.5 3.3 10 4.4 122.0 12.0 10.2 10 10.5 34.0 6.0 5.7 10 8.0 19.0 6.0 3.2 10 12.7 8.5 17.0 0.5 1 10.8 24.0 12.0 2.0 1 10.0 6.5 4.0 1.6 10 8.6 30.0 18.5 1.6 2 10 10 10 3 29.6 11.7 8.3 15.0 9.7 8.7 51.0 3.6 6.0 13.0 6.0 4.2 14.0 1.4 4.0 18.0 3.0 2.6 3.6 2.6 1.5 0.7 2.0 1.6 0.34 ppm in all but one sample. Dieldrin and HCB were present in seven samples, but neither exceeded 0.36 ppm. In a survey of organochlorine residues in 21 aquatic bird species at 31 locations in Alberta, Saskatchewan, and Manitoba, DDE and diel- drin levels were higher in eggs of larids and fish-eating birds than in those of geese and ducks, presumably reflecting different trophic levels between those two groups of birds (Ver- meer and Reynolds 1970). On the Niagara Peninsula, an area of On- tario that is intensively developed for agricul- ture and heavy industry and has a large urban population, eggs were collected in 1972 from 20 species of birds having a variety of feeding habits (Frank et al. 1975). Representative species were obtained from both the terres- trial and aquatic food chains. Highest total DDT residues were in the eggs of aquatic car- nivores, including common tern (22.4 ppm), herring gull (10.4 ppm), black-crowned night heron (Nycticorax nycticorax; 7.8 ppm), and black tern (Chlidonias niger; 7.6 ppm). Herbi- vores and insectivores contained lower total DDT residues regardless of the environment in which they fed. The highest mean residues of PCB's also were in carnivores in the aquatic food chain, including herring gulls (74 ppm), common terns (42 ppm), and black-crowned night herons (27 ppm). Eggs of anhingas {Anhinga anhinga), herons, and ibises were collected during the 1972 nesting season at coastal and inland lo- caUties from Florida to New Jersey (Ohlen- dorf et al. 1974). Measurable residues of DDE 15 occurred in all 209 eggs. The highest mean value (4.0 ppm) was found in great egrets from New Jersey. Among the coastal locali- ties, levels of DDE as well as total DDT pro- gressively declined toward the south. The PCB's occurred second most frequently and also reached their highest mean level (4.2 ppm) in the great egret eggs from New Jersey. Other pollutants occurred less fre- quently and at lower levels. In Great Britain, where the presence of or- ganochlorine pollutants in seabird eggs was first demonstrated (Moore and Tatton 1965), organochlorine residues in seabird eggs from a number of colonies have been monitored. Populations of common puffin (Fratercula arc- tica) in Great Britain declined (Flegg 1971, 1972), but those birds analyzed have not shown excessively high contamination levels. Birds from Saint Kilda contained 7.6 ppm of PCB (61 ppm in fat), but seven other puffins contained lower concentrations. Five eggs ob- tained from Saint Kilda in 1969 contained lower residues of PCB's, DDE, and dieldrin than did eggs of either the common murre or the razorbill (Parslow et al. 1972). figgs of the murre from Lundy, Skomer, and Berry Head contained lower PCB levels than did eggs of the kittiwakes from the same location, but DDE levels were lower in the kittiwakes (Parslow 1973). In some localities on the Brit- ish coast, eggs of murres contained levels of PCB's that were as high as those reported from California and the Baltic, but DDE levels were lowest in Britain. Biological Effects of Organochlorines on Marine Birds Although there is a considerable amount of information on residue concentration and re- productive effects of organochlorines in ma- rine birds, there is relatively little information on toxicology, physiology, and pathology in these species. Therefore, it is particularly rele- vant to consider also such effects in the more frequently studied terrestrial species. Toxicology, Physiology, and Pathology Evidence is substantial that PCB's may have contributed to the mortality of contami- nated birds. Great cormorants found dead in the Netherlands may have died of PCB poisoning (Koeman et al. 1973). Residues in the brain and liver were equivalent to those in birds poisoned through feeding of the PCB preparation Clophen A60. Chlorinated diben- zofurans, however, were present in the com- mercial PCB mixture (Vos et al. 1970) and may have contributed to the mortality of the experimental birds. Therefore, the residue levels in tissues may not be equivalent in the toxicological sense. The occasional "wrecks" of seabirds, par- ticularly of common murres, are usually asso- ciated with storms. In 1970 more than 100,000 murres died in Bristol Bay, Alaska, following stormy weather (Bailey and Daven- port 1972). There had been no oil spills in the area. The birds were emaciated and appar- ently had starved as a result of an inability to find food during the prolonged storm, but they were not analyzed for organochlorines. In Great Britain, PCB concentrations in the livers of gannets {Moms bassanus) that died during large-scale mortality incidents in 1972 ranged from 3,300-9,600 ppm, lipid weight; DDE concentrations ranged from 260- 520 ppm, lipid weight (Parslow et al. 1973). Organochlorine concentrations of this magni- tude might contribute to the death of marine birds either through direct poisoning follow- ing mobilization of fat or through more subtle sublethal effects on the birds at a time of en- vironmental stress. Because concentrations of chemicals in the body are greatly affected by weight gains and losses, it is sometimes more useful to compare total body loads. Estimated body contents of PCB's and DDE in five murres found dead during a 1969 wreck in the Irish Sea were 2,700 /ig (range 800-8,900) and 673 /ig (314- 1,535) (Holdgate 1971). Five birds that were shot in the same general area had 3,500 (ig (800-7,200) of PCB's and 1,484 ng (468-3,211) of DDE. However, eight other murres that died in the wreck had an average estimated body burden of 4,660 ng of PCB, twice as much as in nine other apparently healthy birds that were collected (Parslow and Jef- feries 1973). Depletion of body fat during times of hunger could be expected to mobilize chlorinated hydrocarbons, providing addi- tional stress when the birds are poorly equipped to cope with it. The overall contribu- tion of chlorinated hydrocarbons, particularly 16 PCB's, to such mortality remains to be determined. A glaucous gull {Larus hyperboreus) found in convulsions on Bear Island in the Arctic contained 311 ppm of PCB's and 67 ppm DDE in the liver (Bogan and Bourne 1972); its weakened condition and abnormal coordina- tion were attributed to these high levels. The glaucous gulls in this colony were feeding on the eggs of other seabirds. Necropsy findings and the high level of DDD (200 ppm) in the brain of a common loon {Gavia immer) found in a soybean field in Madison County, Mississippi, indicate that the bird died of DDD poisoning (Prouty et al. 1975). Experiments have been conducted with cap- tive birds to determine which tissue might contain chemical residues that are diagnostic of organochlorine poisoning (L. F. Stickel et al. 1966; W. H. Stickel et al. 1969, 1970, 1973; Stickel and Stickel 1970). There is little doubt that many closely related compounds have a lethal additive effect in the nervous system (Ludke 1976; J. L. Ludke and W. H. Stickel, personal communication). Chemical residues in the brain, in association with pathological conditions of the body, may reveal that the compounds caused death. Lethal ranges have been established for DDT, DDD, DDE, diel- drin, and mirex. Suggestions also have been made for weighing and summing brain resi- dues of DDT, DDD, and DDE for interpreta- tion of field specimens. Recent studies of the induction of hepatic enzymes by PCB's have been reviewed (Peakall 1975). Levels of PCB in many sea- birds may be assumed to be sufficient to in- crease the activity of various mixed function oxidase enzymes. Elevated activity levels of these enzymes also enhance steroid metab- oUsm and degrade nonpolar compounds of foreign origin. The biological consequences of increased steroid metabolism are unknown but birds may compensate for the higher level of steroid metabolism by increasing levels of synthesis. Teratogenic effects observed in experi,- mental feeding studies with PCB's have in- cluded malformations of the eye, legs, and beaks (Carlson and Duby 1973; Tumasonis et al. 1973; Cecil et al. 1974). Such abnormalities may have been caused by contaminant di- benzofurans in the PCB mixtures (Vos et al. 1970; Bowes et al. 1975). Similar abnormali- ties have been found in common terns breed- ing in Long Island Sound (Hays and Rise- brough 1972) but the cause remains unknown; a Unk with PCB or chlorinated dibenzofurans has not yet been proven. Diets containing 10 and 30 ppm (dry weight) DDE were fed to black ducks (Anas rubripes), and diets containing 1, 5, and 10 ppm (dry weight) DDE were fed to mal- lards (Longcore et al. 1971a). Among the results were the following changes in black duck eggshell composition: (1) significant in- crease in the percentage of magnesium, (2) sig- nificant decreases in barium and strontium, (3) increases (which approached significance) in average percentage of eggshell sodium and copper, (4) a decrease in shell calcium that ap- proached significance, (5) patterns of mineral correlations that in some instances were dis- tinct to dosage groups, and (6) inverse correla- tions in the control group between eggshell thickness, magnesium, and sodium. Changes in mallard eggshells were: (1) signi- ficant increase in percentage of magnesium at 5 and 10 ppm DDE, (2) significant decrease in aluminum at 5 and 10 ppm DDE, (3) a signifi- cant decrease of calcium in the 10 ppm DDE group, and (4) an increase in average per- centage of sodium in eggshells from DDE- dosed ducks that approached significance. Blood samples were taken for 2 successive years from canvasback ducks [Ay thy a valis- ineria) trapped in the Chesapeake Bay (Dieter et al. 1976). The first winter (1972-73), five plasma enzymes known to respond to organo- chlorine poisoning were examined. Altera- tions in enzyme activity indicated tissue dam- age (specifically in membrane permeability) at the cellular level. Abnormal enzyme eleva- tions suggested that 20% of the population sampled (23 of 115 ducks) might contain ele- vated levels of organochlorine contaminants, but no residue analyses were performed. The second winter (1973-74), two of the same en- zymes, aspartate aminotransferase and lac- tate dehydrogenase, were assayed in 95 blood samples. The PCB concentrations in represen- tative blood samples were significantly (P < 0.05) correlated with plasma aspartate aminotransferase activity. Male coturnix quail {Coturnix cotumix) were fed diets containing graded levels of DDE, PCB (Aroclor 1254), malathion, and 17 mercuric chloride (Dieter 1974). At 12 weeks, increases in each of the activities of five plasma enzymes (creatine kinase, aspartate aminotransferase, cholinesterase, fructose-di- phosphate aldolase, and lactate dehydroge- nase) of birds were proportional to the log dose of the respective agents. In addition, the pattern of enzyme responses in the experi- mental groups had changed, and was illustra- tive of the specific type of substance that had been fed. The data suggest that qualitative and quantitative identification of environ- mental contaminants in birds, and perhaps a variety of wild animals, may be possible by utilization of multiple plasma enzyme assays. Residue analyses after 12 weeks of feeding showed that DDE accumulated in carcasses and Uvers at concentrations up to fourfold higher than those in the diets. In contrast, residues of Aroclor 1254 attained in carcasses were identical to, and in livers one-half of, the concentration in the feed. Wild-trapped starlings {Sturnus vulgaris) were fed concentrations of DDE or Aroclor 1254 (5, 25, and 100 ppm, dry weight) that were found to be sublethal when fed to pen- reared coturnix quail for 12 weeks (Dieter 1975). Although the experimental design had been to compare plasma enzyme responses at 3, 7, and 12 weeks, reliable measurements could only be made through 7 weeks of the ex- periment because of unexpected mortality. Variations in enzyme response were greater in wild than in pen-reared birds, but not enough to mask the toxicant-induced changes in en- zyme activity. Cholinesterase, lactate dehy- drogenase, creatine kinase, and aspartate aminotransferase activities increased in those fed the organochlorine compounds. Evalua- tion of enzymatic profiles appears to be a po- tentially valuable technique to monitor the presence of toxicants in wild populations, es- pecially if used to complement standard chemical residue analyses. After feeding for 7 weeks, liver residues of either organo- chlorine compound were about threefold higher than the concentrations fed daily. However, 4 times as much DDE as Aroclor 1254 had accumulated in the carcasses. Dietary DDE at levels from 10 to 1,000 ppm (dry weight) inhibited nasal gland secretion in mallards maintained in fresh water (Friend et al. 1973). However, in subse- quent studies on the effects of dietary DDE (10-250 ppm, dry weight) on osmoregulation and nasal gland function in mallards, Pekin ducks, black guillemots (Cepphus grylle), and common puffins, DDE had minimal effects on plasma electrolyte levels and total nasal gland Na,K-ATPase activities in each of these species (Miller et al. 1976). Liver DDE levels in experimental ducks and guillemots were comparable with those reported for seabirds found dead after kills; levels in starved puffins were much higher. Therefore, DDE at envi- ronmental levels may not affect osmoregula- tion of nasal gland Na,K-ATPase in ducks or in these two species of marine birds. Coturnix quail were fed 1 ppm (dry weight) dieldrin, 2 ppm DDE, or the two chemicals to- gether (Ludke 1974). When fed alone, both dieldrin and DDE reached their highest con- centrations in the birds' livers after 28 days on treatment, followed by a slight decrease after 56 days. In whole-body samples (carcass minus liver), dieldrin residues increased stead- ily throughout the treatment period. Dieldrin residues in the birds exposed to dieldrin alone were always similar to residues in birds that were exposed to dieldrin in combination with DDE. In birds fed DDE, either alone or in combination with dieldrin, DDE residues in the carcass increased similarly for 28 days. After 56 days, DDE residues were signifi- cantly greater in the birds fed the dieldrin and DDE mixture. The continued increase of DDE residues when both DDE and dieldrin were fed suggests an interaction in which dieldrin promotes an increased uptake or retention of DDE. No weight loss or mortality occurred among bobwhite (Colinus virginianus) fed a control diet or those fed chlordane (10 ppm, dry weight) alone. However, birds that were fed endrin (10 ppm, dry weight) or a combina- tion of chlordane and endrin lost weight and died within a few days (Ludke 1976). Mori- bund individuals had lost considerable body weight and contained much less body fat than did individuals that were not exhibiting signs of intoxication when sacrificed. Birds that died from intoxication averaged weight losses of 32.2% (endrin-treated) and 31.4% (chlor- dane + endrin-treated) when compared with the control group. Individuals that survived exposure had significantly lower brain resi- dues than those that died. Residues of endrin were significantly lower (by 38%) in brains of 18 birds that died from endrin plus chlordane than in those dying from endrin alone. These data indicate that closely related toxicants may have an accumulative effect at the site of action. Two of 14 male American kestrels {Falco sparuerius) died after 14 and 16 months on a diet containing 2.8 ppm DDE (Porter and Wiemeyer 1972). The brains of the two birds contained DDE residues of 213 and 301 ppm compared with an average of 14.9 ppm (range, 4.5-26.6 ppm) for 11 of the adult males that were sacrificed after 12 to 16 months on dosage. Each of the two birds that died had lost about one-third of its weight since treat- ment began and necropsy revealed typical characteristics (reduced pectoral muscle and badly depleted fat reserves) of organochlorine poisoning. Endrin was consistently the most toxic of 89 pesticidal chemicals that were tested for their lethal dietary toxicity to young bob- whites, coturnix quail, ring-necked pheasants, and mallards (Heath et al. 1972a). Aldrin and dieldrin were among the six most toxic chemi- cals of those tested on all species, and toxa- phene was the only other organochlorine that was particularly toxic to mallards. Major species differences in vulnerability to various chemicals such as were demonstrated in this study must be considered whenever toxicity of particular chemicals to avian species is un- known. Further testing made this point in- creasingly clear (Hill et al. 1975). Among the more toxic organochlorine compounds, nearly all are alicyclic hydrocarbons. Of these chemi- cals tested, most of the aromatic chlorinated hydrocarbons are among the less toxic. Toxicities of six PCB compounds (Aroclor 1232, 1242, 1248, 1254, 1260, and 1262) to penned mallards, pheasants, bobwhite, and coturnix quail were generally less than that of DDT (Heath et al. 1972b). Aroclor toxicity was positively correlated with chlorine per- centage (last two digits of Aroclor number) for the 2-week-old birds that were fed treated diets for 5 days. The joint toxicity of Aroclor 1254 and DDE on coturnix was additive, not synergistic. When 18 chemicals (including 8 organochlorines) were fed in 13 pairs to co- turnix quail and ring-necked pheasant, the effects of the organochlorines also were addi- tive rather than synergistic (Kreitzer and Spann 1973). To learn if the percentage of chlorine in a mixture of PCB's alone determines toxicity, Hill et al. (1974) fed coturnix quail diets con- taining Aroclor 1248, 1254, or 1260 at levels that added equal amounts of chlorine to the feed. Sublethal concentrations produced no detectable effects. Lethal concentrations with equal chlorine showed Aroclor 1248 to be the least toxic of the three compounds at the highest chlorine concentrations. At lower con- centrations, Aroclor 1254 was the most toxic Aroclor. Although chlorine percentage of a PCB is positively correlated with its avian toxicity, PCB toxicity is apparently not simply a function of chlorination. Toxicity also is related to the positions the chlorine atoms occupy on the benzene rings. Toxicity of hexachlorobiphenyl mixtures to bird em- bryos has been shown to be correlated with their dibenzofuran content (Vos and Koeman 1970; Vosetal. 1970). Experiments with coturnix quail were used to simulate the stresses on wild birds of breed- ing condition and of weight loss due to migra- tion (Gish and Chura 1970). Light conditions in the laboratory were manipulated to stimu- late reproductive development in one group of birds and suppress development in another group. Within each of these groups, some birds were partially starved before dosage and some were fully fed. Birds were then fed dietary levels of 0, 700, 922, 1,214, or 1,600 ppm (dry weight) of DDT for a period of 20 days or until death. Birds partially starved before dosage were more susceptible to DDT intoxication than nonstarved ones. Similarly, males died earlier than females, and the lighter birds died earlier than the heavier ones. The heavier birds of each sex not only survived longer than lighter individuals re- ceiving the same treatments, but they also lost a greater proportion of their weight before death. During the early portion of the dosage period, females in breeding condition were less sensitive to DDT than were non- breeding females and males. After 10 days on dosage, however, the cumulative mortality of females in breeding condition rapidly ap- proached that of males and of females not in breeding condition. Reproduction Field and experimental evidence indicates that declines in eggshell thickness observed in 19 certain species in North America and Great Britain since the mid- 1 940 's have been largely caused by residues of p,p '-DDE or other com- pounds or metabolites of the DDT group (Cooke 1973). At moderate or high levels of DDE, shell thinning is severe and eggs may break during incubation. High DDE levels have been recorded in California; species af- fected there have included brown pelicans (Risebrough et al. 1971), double-crested cor- morants (Gress et al. 1973), great egrets, and great blue herons (Faber et al. 1972). As indi- cated previously, much of the DDE probably originated from an insecticide manufacturing plant in southern California. DDE levels asso- ciated with the shell thinning of eggs of the common murres (Gress et al. 1971) and ashy storm petrels (Coulter and Risebrough 1973) on the Farallon Islands in central California may also have originated in part from this particular source. Eggshell thinning has occurred in several other species that occur in freshwater or es- tuarine habitats or that nest on coastal islands. In 1967, shell thickness in herring gull eggs from five States decreased with in- creases in chlorinated hydrocarbon residues (Hickey and Anderson 1968). Comparison of eggshells taken before 1946 with those taken since then reveals that several species includ- ing the peregrine falcon, brown pelican, double-crested cormorant, black-crowned night heron, bald eagle, and osprey have sustained shell-thickness and shell-weight decreases of 20% or more, at least for brief periods (Ander- son and Hickey 1972). In some of these, re- gional population declines are known. How- ever, in seabird species that depend upon ma- rine food chains in Iceland, there was no evi- dence of shell thinning in 1973 (J. A. Sproul et al., unpublished manuscript). Shell thickness was significantly and in- versely correlated with the concentration of DDE in 40 great blue heron eggs from Alberta (Vermeer and Reynolds 1970; Vermeer and Risebrough 1972). In the Upper Great Lakes States, 9 of 13 species of fish-eating birds were found in 1969-70 to have sustained statistically signifi- cant decreases in eggshell thickness since 1946 (Faber and Hickey 1973). Maximum changes in a thickness index occurred in great blue herons (-25%), red-breasted mergansers (Mergus serrator; -15%), and double-crested cormorants (-15%). Heron eggs taken in Lou- isiana generally displayed a smaller post- 1946 change than herons in the Middle West. Al- though DDE was a prominent factor for most groups, especially herons, in relation to the eggshell thinning observed, dieldrin and PCB's also were associated with thinning in some species. This relationship, however, may have been due to correlation in concentrations of these chemicals and concentrations of DDE. The thinning of eggshells of the brown peli- can has proven to be related to the concentra- tions of DDE in the eggs (Blus et al. 1971; Blus et al. 1972a, 1972b). Nearly all brown pelican eggs collected from 13 colonies in South Carolina, Florida, and California in 1969 and from 17 colonies in South Carolina and Florida in 1970 exhibited eggshell thinning (Blus 1970; Blus et al. 1974a). Of the 100 eggs analyzed for residues of pollutants, all eggs contained measurable quantities of DDE; most eggs contained measurable quan- tities of DDD, DDT, dieldrin, or PCB's. DDE appears to have been responsible for virtually all the eggshell thinning. Nest success of brown pelicans in South Carolina was related to residues of DDE and dieldrin in sample eggs (Blus et al. 1974b). Residues of DDE seemed primarily respon- sible for nest failure; however, deleterious ef- fects of this pollutant on nest success was not satisfactorily separated from those induced by dieldrin. Significant intercorrelation of all five organochlorine residues identified in the eggs complicated the relationship of residues to nest success. Maximum DDE residues in an egg from a successful nest were 2.4 ppm and in an egg from an unsuccessful nest, 8.5 ppm. Comparable maximum residues for dieldrin in sample eggs were 0.54 ppm (suc- cessful) and 0.99 ppm (unsuccessful). Resi- dues of DDD, DDT, or PCB's in sample eggs were not significantly related to nest success. Reproductive success in the brown pelican colony was subnormal in the 2 years of study (1971 and 1972) but reproductive success was normal in those nests in which the sample egg contained either 2.5 ppm or less of DDE, or 0.54 ppm or less of dieldrin. Residues of DDE, DDD, DDT, dieldrin, and PCB's exhibited a significant decline in South Carolina brown pelican eggs from 1969 through 1973 (Blus et al. 1977a), but the de- 20 crease in DDD was greatest. In 1973, the peli- cans experienced excellent reproductive success for the first time in many years, and the decline in residues was related to this im- provement. DDE was implicated as the agent responsible for most pollutant-induced nest failure; residues above 3.7 ppm in the sample egg were associated with total failure of those eggs remaining in the nest. The improvement in reproductive success was not associated with an increase in average eggshell thick- ness. The peregrine falcon appears to be affected by shell thinning in all areas of its nearly glo- bal range thus far examined, including areas in the Aleutians (Peakall et al. 1975), Green- land (Walker et al. 1973), and coastal Chile (Walker et al. 1973) where they depend on ma- rine food chains. On the Auckland Island in the sub- Antarctic, one egg of the New Zealand falcon contained DDE residues that were similar to those associated with shell thinning in the closely related peregrine (Bennington et al. 1975). Peregrine falcons that breed along the coast of Scotland feed largely on seabirds, and these populations have declined in numbers at a time when populations that were preying on land birds in the interior remained stable (Rat- cliffe 1972). A decline in reproductive success of the white-tailed eagle {Haliaeetus albicilla) in Germany has most likely been caused by DDE (Koeman et al. 1972b). In the Baltic, where white-tailed eagle populations declined during this century (Henriksson et al. 1966), very high concentrations of PCB and DDT compounds have been measured in eagles that were found dead (Jensen et al. 1972). During an early study, the population of the Bermuda petrels (Pterodroma cahow) was undergoing an unexplained decline that was attributed to the presence of DDT (Wurster and Wingate 1968), but reproductive success subsequently improved. Reexamination of the tissues that had been analyzed for DDT, and analysis of dead chicks and unhatched eggs obtained subsequently, showed no changes in either DDT or PCB concentrations during the periods of poor reproductive success and sub- sequent recovery. Moreover, residues were comparatively low when related to those of other species of petrels in more contaminated areas (D. Wingate and R. W. Risebrough, un- published data). Shell thinning of eggs of the osprey in the northeastern United States where reproduc- tion has been low and where population numbers have declined is also related to DDE concentrations (Spitzer et al. 1977). Dieldrin and PCB's also may have contributed to the rapid population decline in the affected areas in the Northeast, principally Connecticut (Wiemeyer et al. 1975). In the Northeast, shell thinning has been documented in eggs of the gannets breeding on Bonaventure Island (J. A. Keith, personal communication). The breeding population of gannets, after increasing over the previous 80 years, declined by 16% between 1969 and 1973 (Nettleship 1975). In the recent past, DDT was extensively used in forest spray op- erations in adjacent areas of New Brunswick. Patterns of reproductive failure in declining populations of several European and North American raptorial species were duplicated experimentally with captive American kes- trels that were given a diet containing dieldrin and DDT, two commonly used organochlorine insecticides (Porter and Wiemeyer 1969). Major effects on reproduction were increased egg disappearance, increased egg destruction by parent birds, and reduced eggshell thickness. In other experimental studies, DDE has caused significant eggshell thinning in cap- tive screech owls (Otus asio) (McLane and Hall 1972) and American kestrels (Wiemeyer and Porter 1970). The levels of DDE found in the kestrel eggs in the second reproductive season of that study are similar to those found in British peregrine falcon eggs (Rat- cliffe 1967). Bald eagle eggs collected in 1968 from nests in Wisconsin, Maine, and Florida all con- tained residues of DDE, DDD, dieldrin, hep- tachlor epoxide, and PCB's (Krantz et al. 1970). Many also contained traces of DDT. Eggs from five nonproductive nests in Maine contained much higher residues than did eggs collected from either productive or nonpro- ductive nests in Wisconsin and Florida. Twenty-three bald eagle eggs collected in Alaska, Maine, Michigan, Minnesota, and Florida during 1969 and 1970 were analyzed for organochlorines and mercury (Wiemeyer et al. 1972). All eggs contained residues of DDE, dieldrin, PCB's, and mercury. Average residue concentrations were lowest in eggs 21 from Alaska. Significant eggshell thinning has occurred among eggs in samples from most major areas. Some eggs contained DDE residues of the same magnitude as those that produced shell thinning in experimental species. High dieldrin residues in some eggs could have an adverse effect on reproductive success. Egg failure was the major cause of poor re- productive success of ospreys on the Potomac River during 1970 (Wiemeyer 1971). Many eggs disappeared between visits to the nests; some were found broken or damaged in the nests, and others failed to hatch. Osprey eggs were exchanged between Con- necticut and Maryland nests in 1968 and 1969 to determine which environmental factors might have contributed to the decline in re- productive success of Connecticut ospreys (Wiemeyer et al. 1975). Incubation of 30 Con- necticut osprey eggs by Maryland ospreys did not improve the hatching rate. Forty-five Maryland osprey eggs incubated by Connecti- cut ospreys hatched at their normal rate. The results of the egg exchanges and associated observations indicated that the most prob- able cause of the poor reproduction of Connec- ticut ospreys was related to contamination of the birds and their eggs. Residues of DDT and its metabolites, dieldrin, and PCB's were gen- erally higher in fish from Connecticut than from Maryland. There were no major changes in residue content of Connecticut eggs col- lected in 1968-69 compared with those col- lected in 1964. One Connecticut osprey had a concentration of dieldrin in its brain that was in the lethal range. The average shell thick- ness of recently collected osprey eggs from Connecticut had declined 18%, and those from Maryland had declined 10% from pre- 1947 norms. Dieldrin, DDE, and PCB's are three environmental pollutants that have most likely been important factors in the greatly reduced reproductive success and rapid population decline of Connecticut ospreys. All black duck eggs that were collected in 1971 from the northeastern United States and Canada contained DDE residues (Longcore and Mulhern 1973). Means for States and Provinces ranged from 0.09 to 5.94 ppm, with mean concentrations exceeding 1.0 ppm in eggs from Maine, New York, New Jersey, and Delaware. The highest DDE concentration (14.0 ppm) was in an egg from Delaware. The DDD and DDT residues averaged < 0.5 ppm for each collection area. No mirex residues and only trace amounts of dieldrin and hepta- chlor epoxide were detected. Of the 61 eggs, 57 contained PCB's; means ranged from < 0.05 ppm in samples from Nova Scotia to 3.30 ppm in those from Massachusetts, with trace amounts occurring in nearly half the samples. Mean organochlorine pesticide resi- dues were lower in the 1971 samples than in those analyzed in an earlier study in 1964. Average shell thickness of eggs collected in 1964 (0.321 mm) was significantly less {P < 0.01) than that of eggs collected before 1940 (0.348 mm) or in 1971 (0.343 mm). Eggs of captive black ducks fed diets con- taining DDE at 10 and 30 ppm (dry weight; approximately 3 and 9 ppm wet weight) ex- perienced significant shell thinning and an in- crease in shell cracking when compared with eggs of untreated black ducks (Longcore et al. 1971b). Survival of ducklings from dosed parents in terms of "percentage of 21- day ducklings of embryonated eggs" was 40- 76% lower than survival of ducklings from un- dosed parents. Average DDE residues in eggs from hens fed 10 and 30 ppm DDE were 46 ppm and 144 ppm. In another experiment, black duck hens fed 10 ppm (dry weight) of DDE in the diet laid eggs with shells 22% thinner at the equator, 30% thinner at the cap, and 33% thinner at the apex than those of controls (Longcore and Samson 1973). Natural incubation increased shell cracking more than fourfold as compared with mechanical incubation. Hens removed cracked eggs from nests, and one hen termi- nated incubation. Hens fed DDE produced one-fifth as many ducklings as did the con- trols. The DDE in eggs of dosed hens aver- aged 64.9 ppm. Concentrations of 10 and 40 ppm DDE (dry weight) in the feed of penned mallard ducks caused significant eggshell thinning and cracking and a marked increase in embryo mortality (Heath et al. 1969). In other studies, eggshell thinning also occurred in mallards fed DDE (Haegele and Hudson 1974), DDT (Tucker and Haegele 1970; Davison and Sell 1974), or dieldrin (Lehner and Egbert 1969; MuUer and Lockman 1972; Davison and Sell 1974), but low dietary levels (25 and 50 ppm) of Aroclor 1254 produced no measurable re- productive effects (Heath et al. 1972b). 22 Diets containing various levels of DDT (at 20 ppm, dry weight, or greater), or dieldrin (at 10 ppm, dry weight) caused significant reduc- tion in eggshell thickness, weight, and cal- cium in mallard ducks (Davison and Sell 1974). The reduction in eggshell thickness was linear with increasing dose of dieldrin through all levels studied. Mallards were fed untreated feed or feed containing 40 ppm (dry weight) DDE, 40 ppm PCB, or 40 ppm DDE + PCB beginning a month before laying (Risebrough and Ander- son 1975). Mean shell thickness indices were similar in the control and PCB groups, but they were reduced by 17% in the DDE group and 19% in the DDE + PCB group. The con- tents of 12 eggs randomly selected from the DDE group contained 373 ppm DDE (lipid basis), and 13 eggs from the DDE + PCB group contained mean residues of 344 ppm DDE + 364 ppm PCB (lipid basis). Egg pro- duction was similar in all groups for about the first 7 weeks, then it dropped markedly in the DDE + PCB group. Part, but not all, of this group's lower production of intact eggs was caused by egg eating. This behavior ac- counted for 18 of 282 eggs observed lost in the DDE + PCB group, 6 of 394 eggs in the PCB group, and none in the control and DDE groups. Although there was no significant change in shell thinning or DDE residues when PCB was added to the diet, the reduc- tion in the number of intact eggs produced by the DDE + PCB group suggests that the two compounds may nevertheless interact to in- fluence reproductive success. Behavior In England, gray herons {Ardea cinerea) have been observed breaking their own eggs, and others dropped their live young from the nest (Milstein et al. 1970; Prestt 1970). Such aberrant behavior may be related to sublethal organochlorine residues in the birds, as these authors suggested. The birds did not eat the eggshells, but tossed even the fragments from the nest. Therefore, the alternative possibility of calcium "hunger" does not seem to be true in herons. Mallard ducks fed a diet containing 3 ppm DDE (dry weight; equal to about 0.6 ppm in a natural succulent diet) laid eggs that con- tained an average of 5.8 ppm DDE; duckUngs that hatched from these eggs differed from controls in behavioral tests designed to measure responses to a maternal call and to a frightening stimulus (Heinz 1976b). In re- sponse to the maternal call, ducklings from parents fed DDE were hyper-responsive; com- pared with controls, a greater percentage ap- proached the call and a greater percentage of those that approached remained near" the call for the remainder of the test. In a test of avoidance behavior, ducklings whose parents were fed DDE traveled shorter distances from the frightening stimulus than did controls. Coturnix quail chicks were given sublethal amounts of chlordane, dieldrin, endrin, DDE, or Aroclor 1254 in their feed, beginning at 7 days of age, and their avoidance response to a moving silhouette was measured daily for 14 days (Kreitzer and Heinz 1974). The birds were on dosage for 8 days, and on untreated feed for 6 days immediately thereafter. Group avoidance response was significantly sup- pressed {P from 0.01 to < 0.001) by chlordane, dieldrin, endrin, and Aroclor 1254, but no effect of DDE on the birds' behavior could be detected. The behavior of the endrin-treated birds returned to normal after 2 days on un- treated feed. The data indicated partial recov- ery for birds treated with dieldrin and chlor- dane, but none for those treated with Aroclor 1254. Heavy Metals The sources, occurrence, food web transfer, and toxicology of heavy metals and other trace elements must be understood to eval- uate the significance of these chemicals to ma- rine birds. These more general aspects have received considerable attention in recent sym- posia and reviews (Larsson 1970; Nelson et al. 1971; Gavis and Ferguson 1972; Eisler 1973; National Research Council of Canada 1974; Leland et al. 1975). Consequently, our discus- sion will be restricted to the more specific aspects of the exposure of aquatic birds to these chemicals, but will include some inter- pretive information relative to terrestrial avian species. Most techniques that are used for measur- ing mercury residues in environmental samples determine levels of total mercury, re- gardless of the chemical form in which it 23 occurs. The various forms of mercury, how- ever, differ widely in their toxicities. Unless otherwise specified, mercury concentrations presented here represent concentrations of total mercury. Exposure of Marine Birds to Heavy Metals Animals acquire heavy metals from the foods they eat, from the water that surrounds them, and possibly from the air they breathe. Quantities accumulated differ greatly among organisms, depending upon exposure and physiology (White and Stickel 1975). Mercury in tissues of living organisms is often primarily in the more toxic methyl mer- cury form (Westoo 1967; Fimreite 1974), and methyl mercury is readily incorporated into the bodies of aquatic organisms (Leland et al. 1975). Most of the mercury in fish is in the form of methyl mercury (Koeman et al. 1975), but the high mercury concentrations dis- covered in the livers of six dead great cormor- ants and in hvers of three others that were col- lected in the Netherlands were not primarily methyl mercury (Koeman et al. 1973). Mer- cury concentrations, primarily in forms other than methyl mercury, increased with age in some marine mammals and were correlated with concentrations of selenium and bromine (Koeman et al. 1975; Martin et al. 1976). Per- haps, like some marine mammals, cormorants may be able to detoxify methyl mercury by a chemical mechanism in which selenium and bromine are involved. However, mercury and selenium concentrations in livers of common murres and of a razorbill (Koeman et al. 1975) and in liver and breast muscle of sooty terns of known age were not correlated (P. G. Connors et al., unpublished manuscript). Inor- ganic and organic mercury from industrial sources may be converted into methyl mer- cury by some organisms, including birds (Jensen and Jernelov 1969; Kiwimae et al. 1969). Mercury concentration increases with body weight, or age, in fish (Bache et al. 1971; Fim- reite et al. 1971), crayfish (Vermeer 1972), and herons (Hoffman 1974). The concentration in- creases at higher trophic levels in fish, other aquatic organisms, fish-eating birds, or ducks (de Goeij 1971; Fimreite et al. 1971; Fimreite 1974; Hoffman 1974; Kleinert and DeGurse 1972; Vermeer et al. 1973; Baskett 1975). Mercury concentrations in various tissues of the body are correlated with each other (Fimreite 1971; Koeman et al. 1971; Vermeer and Armstrong 1972a; Fimreite 1974; Heinz 1974, 1976a; Hoffman 1974). Eggs normally contain between a fifth and a ninth of the mer- cury concentration in the liver of the female (Fimreite et al. 1970; M. T. Finley, personal communication). Mercury in the liver of fe- male California gulls (Larus californicus) aver- aged 5.5 times that in their eggs (Vermeer 1971a). High mercury residues in aquatic organisms and in the related avifauna are often related to discharges from chlor-alkali plants, pulp mills, or other industrial plants that use mercury (Fimreite 1970; Fimreite et al. 1971; Nelson et al. 1971; Vermeer 1971a). Ospreys and great crested grebes (Podiceps cristatus) now have about 3 times as much mercury in some industrially contaminated areas as in uncontaminated areas (Larsson 1970). In a survey of aquatic birds at 33 locations in Alberta, Saskatchewan, and Manitoba, mercury levels were generally higher in gulls {Larus spp.) and fish-eating birds than in ducks and geese (Vermeer 1971a). The highest mercury levels were found in herring gulls, possibly related to their scavenging and fish- eating habits. Elevated mercury levels were found in livers of common mergansers {Mergus mer- ganser; up to 86 ppm), common loons (90ppm), and great blue herons (128 ppm) from Ontario (Fimreite 1974). Lower concen- trations were found in mallards (12.5 ppm) and pintails (Anas acuta; 6.2 ppm). Mercury levels were higher in adults than in imma- tures. A chlorine plant about 80 km upstream from the collecting locahty was believed to be the source of mercury found in the birds. Mercury was present in spotted sandpiper {Actitis macularia) eggs collected upstream from Edmonton, Alberta, at lower concentra- tions (0.09 ppm) than in those eggs collected downstream (0.28 ppm), suggesting municipal or industrial contamination originating at Edmonton (Vermeer 1971b). During another survey in Canada, highest concentrations of mercury in livers of fish-eat- ing birds collected near sites of industrial con- tamination were in red-necked grebes 24 (Podiceps grisegena; Fimreite et al. 1971). Four common tern eggs averaged 0.58 ppm and two red-breasted merganser eggs aver- aged 0.81 ppm. Aquatic bird eggs from the upper Great Lakes States contained higher mercury levels than those from Louisiana, although species represented from the two areas were not iden- tical (Faber and Hickey 1973). Highest mean residues were in three species of mergansers (up to 1.6 ppm; red-breasted merganser). For those species with all eggs containing less than 0.25 ppm of mercury, the residues were considered to represent background levels. Mercury exceeded 1 ppm in one or more eggs of black-crowned night heron, hooded mer- ganser (Lophodytes cucullatus), common mer- ganser, and red-breasted merganser. Highest levels (up to 1.9 ppm) were in addled eggs of red-breasted mergansers. Many birds dependent upon aquatic areas in the Lake St. Clair, Michigan, region have high residues of mercury in their tissues (Dustman et al. 1972). In 1970, carcasses, livers, and eggs were collected and analyzed. Mercury levels in great blue herons (up to 175 ppm in the liver; 23 ppm in the carcass) and common terns (up to 39 ppm in the liver; 7.5 ppm in the carcass) far exceeded those in any other species. The levels are comparable to those in birds in Sweden that died under ex- perimental dosage with methyl mercury and in birds that died under field conditions in sev- eral Scandinavian countries with signs of mer- cury poisoning (Henriksson et al. 1966; Borg et al. 1969; Holt 1969). Mercury residues in eggs of all of the five common terns (up to 6.2 ppm), five of nine mallards (up to 2.7 ppm), three of the five black-crowned night herons (up to 1.1 ppm), and the single egg of a pied- billed grebe (Podilymbus podiceps; 4.0 ppm) were in the range of residues (0.5-3.1 ppm) in eggs of ring-necked pheasants whose repro- ductivity was reduced by mercury in experi- mental studies (Borg. et al. 1969; Fimreite 1971; Spannetal. 1972). In 1973, eggs of some of these species were again collected at Lake St. Clair, following re- strictions on industrial discharges of mercury into the St. Clair River (Stendell et al. 1976). Mercury levels in the eggs were appreciably lower than were found in these species in 1970. Common terns contained the highest residues (up to 1.3 ppm). Mallard eggs con- tained relatively low residue levels (<0.05 to 0.26 ppm). Black-crowned night heron eggs (up to 0.76 ppm) and great egret eggs (up to 0.45 ppm) contained intermediate amounts. Mercury levels generally are low in most species of ducks and geese but higher levels have been found in those species that con- sume a greater proportion of animal material in their diet (Kleinert and DeGurse 1972; Krapu et al. 1973; Fimreite 1974; Heath and Hill 1974). Among North American waterfowl species, the highest levels have been found in mergansers. Common mergansers from On- terio had up to 86 ppm mercury in their livers (Fimreite 1974). Hooded mergansers from Clay Lake, Ontario, contained up to 12.3 ppm and common goldeneyes (Bucephala clangula) up to 7.8 ppm in their breast muscle (Vermeer et al. 1973). Food items were also analyzed and crayfish {Oronectes virilis), which the hooded mergansers eat, contained the highest average concentration of mercury (7.1 ppm). Mercury has been found in the visceral fat of black-footed albatrosses and Laysan alba- trosses from Midway Atoll, North Pacific Ocean (Fisher 1973). Average residue levels in the Laysan albatrosses were 0.104 ppm and those in the black-footed were 0.075 ppm. Mercury has been found in the hvers of birds collected around the British coast (Dale et al. 1973). The highest concentration (26 ppm; converted from 122 ppm dry weight, see Holdgate 1971) was in a red-breasted mer- ganser. Common eiders, which feed on mussels that are known to accumulate mer- cury, also had high concentrations (10 ppm). All of the more pelagic species, including black-legged kittiwakes, fulmars, and auks {Alca torda and Alle alle) had less than 2.2 ppm. Three gannets had slightly higher levels (up to 2.9 ppm). Herring gulls from oceanic islands contained relatively low mer- cury residues (up to 2.6 ppm) like the pelagic birds, but those from near shore had higher residues. Common puffins collected around the coast of Britain contained up to 7.7 ppm (Parslow et al. 1972), and eiders from the Tay region had up to 0.45 ppm mercury in their livers (Jones etal. 1972). Elevated levels of mercury have been found in birds of the Baltic region (Jensen et al. 1972). In 1969, mercury content of common murre secondaries had doubled the levels from 1906-1325. Mercury levels in murre eggs 25 were approximately the same in 1968 and 1969, averaging about 0.52 ppm, with the upper extreme concentration of 0.67 ppm. Even higher levels of mercury (3.7 ppm) were found in muscle tissue of great cormorants than in muscle of murres (0.9 ppm) or black guillemots (1.8 ppm) from the Baltic. Mercury in feathers (up to 51 ppm), muscle (up to 26 ppm), and brains (up to 14 ppm) of white- tailed eagles exceeded the levels in other species. Mercury concentrations in the kid- neys (48-123 ppm) and muscle tissue (1.9- 8.5 ppm) of other white-tailed eagles from the same area further indicate that the species may have serious mercury pollution problems (Henriksson et al. 1966). Bald eagles in the United States also occasionally contain high levels of mercury (up to 43 ppm) in their car- casses (Belisle et al. 1972). Mercury levels (figures not specifically stated) in the muscle of eiders and sandwich terns of the Dutch Wadden Sea appear 3 to 5 times higher than the levels considered rep- resentative of natural background (de Goeij 1971). Analyses were also made of various organs of three common murres and one razorbill that were found as oiled birds along the Dutch coast (Koeman et al. 1975): mer- cury in the livers did not exceed 2.5 ppm; sele- nium in the liver of one common murre was 4.6 ppm, but the levels of these metals were not correlated with each other. There were no significant geographical or species differences in two essential heavy metals (copper and zinc) in Antarctic and North American petrels (Anderlini et al. 1972). Silver, cobalt, and lead were difficult to detect at the low levels that were found, but there were no detectable differences in their concentrations. Cadmium, chromium, nickel, and mercury levels in petrels suggested a cor- relation of increasing concentration with in- creased exposure to industrialized areas. Higher concentrations of these metals in ashy petrels are probably the result of their feeding in the proximity of San Francisco Bay. Livers from ruddy ducks killed by an oil spill on the Delaware River contained detect- able levels of lead, cadmium, and mercury (White and Kaiser 1976). Lead ranged from 0.19 to 0.61 ppm, cadmium from 0.27 to 1.60 ppm, and mercury from 0.06 to 0.74 ppm. Residues of these metals were simi- lar to those found in canvasbacks from the Chesapeake Bay region (D. H. White and R. C. Stendell, unpublished manuscript). Mercury residues in the livers of six gannets from the Irish Sea (4 ppm; 18.4 ppm dry weight) averaged higher than in two gannets from eastern Scotland (1.6 ppm; 7.3 ppm dry weight) that died during unrelated large-scale mortality incidents (Parslow et al. 1973). Average levels of copper (7.4 ppm; 34 ppm dry weight) and zinc (64.8 ppm; 298 ppm dry weight) in the livers of gannets from the Irish Sea also were higher than in two others from eastern Scotland (2.8 ppm copper; 26.3 ppm zinc). The differences in the metal concentra- tions between the two groups were considered the result of the differences in liver sizes. Al- though the cause of the gannet deaths could not be established, heavy metal concentra- tions in the birds apparently were responsible for the death of only one individual with high mercury levels (22 ppm; 98 ppm dry weight). Lead and cadmium concentrations were below the limits of detection in all of these birds, but another gannet that died in an earlier incident had measurable residues of lead (0.2 ppm) and cadmium (2.0 ppm). Mercury concentrations in the livers and kidneys of common murres that died in the seabird wreck in the Irish Sea during autumn 1969 did not exceed 5 ppm (23 ppm dry weight) (Holdgate 1971). Some of the birds showed relatively high levels of particular metals and in some the highest concentra- tions were above the level at which poisoning may have occurred. However, in general, the range of mercury levels in the casualties of the incident and in the healthy birds shot for com- parison overlap. The levels of mercury (up to 5 ppm), lead (8.7 ppm), cadmium (2.8 ppm), and arsenic (8.3 ppm) in the livers and kid- neys of some birds appeared elevated. Biological Effects of Heavy Metals on Marine Birds Toxicology, Physiology, and Pathology In 1953 a severe neurological disorder caused by mercury poisoning was first recog- nized among people living in the vicinity of Minamata Bay, Japan (Kurland et al. 1960). Toxic effects and similar histopathological changes have been reported for fish, birds, and mammals that died as a result of mercury 26 poisoning, but no particular studies have been made on the toxicity of heavy metals to sea- birds (Parslow et al. 1973). However, in certain terrestrial species, symptoms of poi- soning might be expected when mercury con- centrations in liver or kidney tissues reach about 30ppm (W. H. Stickel 1971). By con- trast, normal levels are less than 1 ppm. Although death may not have been caused by mercury poisoning, mercury in the livers of adult great egrets found dead in California ranged between 2 and 9.5 ppm (Faber et al. 1972). Mercury in the liver (22 ppm; 98 ppm dry weight) of a gannet from the Irish Sea could have caused the bird's death (Parslow et al. 1973). Female mallards fed 3 ppm mercury (dry weight) as methyl mercury in their diet had average mercury residues of 11.1 ppm in their livers, 14.7 ppm in their kidneys, 5.0 ppm in their muscles, 4.6 ppm in their brains, and 5.5 to 7.4 ppm in their eggs (Heinz, 1976a). Males had higher residues, and many of the duckhngs from these parents died within 1 week after hatching (Heinz 1974, 1976a). The ducklings also had high levels of mercury in their tissues. In short-term tests of lethal dietary toxicity of pesticidal chemicals, Ceresan M, a fungi- cide containing ethyl mercury, was relatively more toxic to young mallards than were 37 other compounds (Heath et al. 1972a). Only endrin and Dasanit were more toxic. In simi- lar subsequent tests, Morsodren, another organomercurial fungicide, was also highly toxic to young mallards (Hill et al. 1975). Mercury potentiated the toxicity and biochemical effects of parathion in coturnix quail fed a sublethal concentration of Morso- dren (4 ppm dry weight as methyl mercury) for 18 weeks (Dieter and Ludke 1975). Mean residue concentrations in these birds were 21 ppm of mercury in the liver and 8.4 ppm in the carcass. The computed LD50 of parathion was 5.86 mg/kg in birds not fed Morsodren and 4.24 in those fed the heavy metal. When challenged with a sublethal oral dose of para- thion (1.0 mg/kg), Morsodren-fed birds ex- hibited significantly greater inhibition of plasma and brain cholinesterase activity than controls. After administration of various mercury compounds to domestic chickens (Gallus gallus), the methyl mercury compounds were rather evenly distributed among the organs, whereas the other mercury compounds, or- ganic and inorganic, gave very high concen- trations in the liver and kidneys compared with other organs (Kiwimae et al. 1969). Dif- ferences in the proportion of methyl mercury compounds to total mercury occurred in the white and the yolk of the eggs from these hens. Although the proportion in the white was similar to that in the blood and the muscles, the proportion in the yolk was simi- lar to that in the liver and the kidneys. The al- bumen contained mainly methyl mercury compounds in concentrations that varied with the compound given to the hens. The methyl mercury concentration in albumen was always much lower when other compounds were administered than when the hens were given the same quantity of methyl mercury hydroxide. In another study, mercury was not detect- able in the albumen but was present at high levels in the yolk following intravenous injec- tion of mercuric nitrate into laying coturnix quail (Nishimura et al. 1971). Evaluation of enzymatic profiles appears to be a potentially valuable technique for moni- toring the presence of toxicants in wild popu- lations, especially if used to complement standard chemical residue analysis (Dieter 1975). Lactate dehydrogenase activity in- creased twofold and cholinesterase activity decreased in birds fed Morsodren. After feed- ing for 3 weeks, mercury in starling carcasses reflected the concentrations fed daily, whereas the concentration in the livers was 2 to 4 times that in the diet. A decrease in cholinesterase activity oc- curred in male coturnix quail that were fed diets for 12 weeks containing graded levels of mercuric chloride (Dieter 1974). At 12 weeks the decrease was proportional to the log dose received, although this was not true after 2 and 4 weeks on the treated diet. Mercury residues attained in the tissues were 5% or less of those in the feed. There was a marked sexual difference in rates of mercury loss in coturnix quails (Back- strom 1969). Males lost little of the mercury in their bodies, especially from the brain and muscle, in 30 days, but females had a marked loss in this period, largely because of excre- tion in eggs. Ring-necked pheasants lost 33% to 50% of the mercury from their livers and 27 kidneys in 2 months, and approximately 99% was lost in 6 months (Borg et al. 1969). Ospreys apparently have a similar loss rate of mercury (Johnels et al. 1968). When methyl mercury dicyandiamide was fed to mallard ducks at a concentration of 3 ppm mercury (dry weight), mercury accumu- lated in the eggs to an average of 7.2 and 5.5 ppm in 2 successive years (Heinz and Locke 1976). Mercury in the eggs caused brain lesions in ducklings. Lesions included demye- lination, neuron degeneration, necrosis, and hemorrhage in the meninges overlying the cerebellum. Brains of dead ducklings con- tained an average of 6.2 and 5.2 ppm mercury in the 2 successive years. Upon necropsy, ring-necked pheasants that were killed after receiving 4.2 ppm mercury (dry weight) in their diet for 350 days ap- peared normal, but those that received greater concentrations (12.5, 37.4, or 112 ppm, dry weight) died during the experi- ment and showed variable amounts of subcu- taneous edema and decreasing amounts of subcutaneous and abdominal adipose tissue as survival time on the treated diet increased (Spann et al. 1972), Birds that died on the higher dosages showed signs of neurological disturbance, including ataxia and torticollis, before death. Lead poisoning has long been recognized as a serious problem for waterfowl (Wetmore 1919; Jordan and Bellrose 1951; Bellrose 1959). Histopathological changes occur in the kidneys of mallards as a result of lead shot in- gestion (Locke et al. 1966, 1967). In addition, significant changes in activity of three en- zymes often used to assess hepatic damage occurred in mallard ducks following oral ad- ministration of lead shot (Rozman et al. 1974). The ingestion of one number 4 lead shot by each of 80 pen-reared mallards that were fed whole-kernel corn caused 19% mortality within an average of 20 days (Longcore et al. 1974a). Coating or alloying lead with other metals only delayed mortality among dosed ducks. Disintegrable lead shot with water- soluble binder and lead-containing biochemi- cal additives were as toxic to mallards as com- mercial lead shot. Lead levels in brains, tibiae, and breast muscle of mallard ducks that died and in tibiae of those that were sacrificed increased significantly from dosage with one number 4 lead shot (about 1.4 g) until death (Longcore et al. 1974b). In mallard ducks, lead levels ex- ceeding 3 ppm in the brain, 6 to 20 ppm in the kidney or liver, or 10 ppm in clotted blood from the heart indicated acute exposure to lead. One month after dosage, mean lead levels in mallards given one number 4 all-lead shot were about twice those in tissues of mallards given one number 4 lead-iron shot that con- tained about 50% lead (Finley et al. 1976a). Necropsy of sacrificed ducks failed to reveal any of the tissue lesions usually associated with lead poisoning in waterfowl. Lead in the blood of ducks dosed with all-lead shot aver- aged 0.64 ppm, and 0.28 ppm in ducks given lead-iron shot. Lead residues in livers and kidneys of females given all-lead shot were significantly higher than in males. In both dosed groups, lead levels in wingbones of the females were about 10 times those in males, and were significantly correlated with the number of eggs laid after dosage. It appeared that after the laying hens ingested sublethal amounts of lead shot, high lead deposition in the bone occurred as a result of mobilization of calcium from the bone during eggshell for- mation. Lead levels in contents and shells of eggs laid by hens dosed with all-lead shot were about twice those in eggs laid by hens dosed with lead-iron shot. Lead levels in eggshells best reflected levels of lead in the blood. The inverse correlation between delta- aminolevulinic acid dehydratase (ALAD) ac- tivity and blood lead concentrations was highly significant in canvasback ducks from the Chesapeake Bay (Dieter et al. 1976). ALAD is an important enzyme in hemoglobin synthesis. The activity of this enzyme in the blood provides a sensitive and precise esti- mate of lead contamination in waterfowl. In mallards, lead concentrations in blood were strongly correlated with erythrocyte ALAD activity, suggesting that biochemical re- sponse to two types of lead shot (one all-lead, the other containing 50% lead) depends upon the quantity of lead present (Finley et al. 1976b). Reproduction Mercury levels (3.5 to 1 1 ppm) in the eggs of Swedish white-tailed eagles that failed to hatch indicate that the decline in reproduction of this species could be attributed to mercury poisoning (Borg et al. 1969). A corresponding 28 decline in this species in Finland also was as- sociated with mercury contamination (Hen- riksson et al. 1966). However, as discussed earlier, organochlorines also may be partially responsible for the observed decline. There were apparently no young produced by common loons in Clay and Ball Lakes, On- tario, in 1970 and 1971 (Fimreite 1974). (Both lakes receive effluent from a chlorine plant.) Fledging success of common terns at Ball Lake was 10% of normal, but fledging was normal at nearby Wabigoon Lake, where birds contained lower residues. Average total mercury in the eggs was 3.6 and 1.0 ppm; average methyl mercury was 2.4 and 0.8 ppm in the two colonies. There are considerable differences between species in susceptibility to mercury pollu- tants. Mercury concentrations as high as 16 ppm in western Ontario herring gull eggs apparently did not affect their hatchability (Vermeer et al. 1973), but 0.5 to 1.5 ppm mer- cury in ring-necked pheasant eggs reduced hatchability, reduced egg weight and produc- tion, and produced a large number of eggs without shells (Fimreite 1971). Concentrations of mercury found in the livers of abnormal young terns (Sterna hirundo and S. dougallii) from Great Gull Island (in Long Island Sound) ranged from 0.2 to 1.2 ppm, but were not thought to have caused the abnormalities (Hays and Rise- brough 1972). Livers of normal young terns were not analyzed. Hatchability in the Great Gull Island colony has consistently been greater than 90%, but hatchability of com- mon tern eggs in Lake Ontario colonies has been low. Concentrations of heavy metals in common terns were studied to determine the reason for the difference in hatchability. Con- centrations of cadmium, chromium, cobalt, copper, lead, mercury, nickel, silver, and zinc in bone, liver, breast muscle, and kidneys of adult birds from the two locations were simi- lar (Conners et al. 1975). Therefore, these metals apparently were not responsible for the differences in hatchability. Although the reproductive effects of mer- cury in other species are largely unknown, mercury residues of 0.5 ppm were associated with poor reproductive success in an experi- mental study with ring-necked pheasants (Fimreite 1971). Average mercury residues in field-collected eggs of four species of aquatic- related birds on the Niagara peninsula, On- tario, were between 0.5 and 1 ppm (Frank et al. 1975). These included red-winged blackbird {Agelaius phoeniceus; 0.68 ppm), herring gull (0.74 ppm), black-crowned night heron (0.64 ppm), and common tern (0.83 ppm). Mercury was found in measurable quanti- ties in all of the 100 brown pelican eggs from 13 colonies in South Carolina, Florida, and CaUfornia (Blus et al. 1974a). Six of the 21 pelican eggs from South Carohna contained 0.5 ppm or more of mercury. Sixteen of the 49 pelican eggs from Florida contained 0.5 ppm or more of mercury, and 1 on the verge of hatching contained 1.43 ppm. Methyl mercury at low dietary levels (0.5 or 3.0 ppm, dry weight, equal to about 0.1 or 0.6 ppm mercury on the basis of a natural suc- culent diet) caused lowered reproductive success in experimental mallards and black ducks. Mallards fed 3 ppm mercury in their diet during one reproductive season showed reproductive impairment, but none was evident among birds fed 0.5 ppm (Heinz 1974). Adverse effects in the group fed 3 ppm included a decrease in egg laying, an increase in embryonic mortality, and reduced duckling survival. These effects resulted in the produc- tion of less than half (46.5%) as many 1-week- old ducklings as the controls. Levels of mer- cury reached about 1 ppm in eggs of the birds fed 0.5 ppm mercury and between 6 and 9 ppm in the eggs from ducks fed 3 ppm mercury. The hens from the first reproductive season were kept on diets containing mercury into a second season (Heinz 1976a). During the sec- ond season, levels of mercury in eggs from hens on these diets averaged 0.79 and 5.46 ppm. On a dry-weight basis, the concen- tration of mercury in eggs was about 5 times that in the feed. There were no significant dif- ferences in egg production or hatching suc- cess among control birds and those fed mer- cury. However, duckling survival decreased: ducklings from hens fed 3 ppm mercury during the two reproductive seasons were less likely to survive to 1 week of age than were controls or ducklings from parents fed 0.5 ppm mercury. Mallards whose parents were fed a diet con- taining 0.5 ppm mercury (dry weight) were themselves fed a diet containing 0.5 ppm mer- cury (dry weight) from 9 days of age through their first reproductive season (Heinz 1976c). Mercury in the eggs of these hens fed mercury 29 averaged 0.86 ppm. Hens fed mercury made less efficient use of feed and laid a greater per- centage of their eggs outside their nest boxes compared with controls. They also produced fewer 1-week-old ducklings than did controls, although there had been no difference in duck- ling production by their parents fed 0.5 ppm mercury in the preceding years. The ducklings from dosed parents did not grow as fast as did those from controls. Black ducks given a diet containing 3 ppm mercury (dry weight) as methyl mercury hatched fewer eggs than did controls and fewer of their ducklings survived (Finley and Stendell 1978). Average mercury residues in brain, Uver, and muscle of ducklings that died (3.7, 9.4, and 4.9 ppm) were about twice those in tissues of ducklings sacrificed at 4 weeks of age (1.6, 5.7, and 2.1 ppm). Mercury residues in pheasant eggs were 0.9 to 3.1 ppm following administration of 4.2 ppm mercury (dry weight) in their diet (Spann et al. 1972). The birds exhibited greatly reduced egg production and increased embryo mortality in the few eggs laid. Mercury residues in bobwhite eggs from birds fed a dietary concentration of 1.7 ppm (dry weight; administered as ethyl mercury p-toluene sulfonanilide) averaged 1.6 ppm (J. W. Spann and R. G. Heath, unpublished manuscript). There was a significantly greater mortality among young whose parents re- ceived mercury in the diet. The principal period of increased mortality included the last 5 days of incubation and the first day arfter hatching. Behavior The behavior of mallard ducks whose par- ents were fed a control diet or a diet contain- ing 0.5 or 3.0 ppm mercury (dry weight) as methyl mercury was studied (Heinz 1975). There was no significant difference among controls and ducklings from mercury-treated parents in the percentage of ducklings that approached the tape-recorded maternal call. However, control ducklings moved back and forth toward the call more than ducklings from mercury-treated parents and also spent more time in the end of the runway near the loudspeaker than ducklings whose parents were fed a diet containing 0.5 ppm mercury. Compared to control ducklings, those from parents fed a diet containing either mercury concentration were hyper-responsive in avoid- ance behavior tests. Among mallard ducklings produced in the 2nd year of the study in which hens were fed a control diet or a diet that contained 0.5 or 3 ppm mercury (dry weight), the findings were similar (Heinz 1976a). There were no signifi- cant differences among controls and groups fed mercury in approach responses toward a recorded maternal call and ducklings from mercury-treated parents were hyper-respon- sive compared with controls in avoidance behavior. In the second generation, there were no sig- nificant differences between controls and ducklings from parents fed 0.5 ppm mercury in approach responses to tape-recorded maternal calls, in avoidance of a frightening stimulus, or in open-field behavior (Heinz 1976c). Plastic and Other Artifacts Small plastic beads and irregularly shaped particles up to 0.5 cm in diameter are com- monly found in plankton samples from widely separated oceanic areas, including the north- western Atlantic, Sargasso Sea, Bristol Channel (United Kingdom), and the coastal waters of southern New England (Carpenter and Smith 1972; Carpenter et al. 1972; Morris and Hamilton 1974; Colton et al. 1974). The particles are primarily composed of polysty- rene or polyethylene compounds and have about the same density as seawater. Their various colors include white, green, brown, blue, red, or clear (Carpenter and Smith 1972; Morris and Hamilton 1974; Colton et al. 1974). The polystyrene spherules evidently are of industrial origin, because they have been found in the effluents from manufacture of polystyrene (Hayes and Cormons 1974; Morris and Hamilton 1974). Their abundance in the British Channel water was lowest near the seaward end and greatest in the inner part of the Channel, near the Holm Islands. Ben- thic sediments near the Holm Islands con- tained as many as 20,000 beads/m^ (Morris and Hamilton 1974). Small fish ingest the beads and particles (Carpenter et al. 1972; Kartar et al. 1973). 30 Plastic particles have been found in the stomachs of fork-tailed petrels, horned puffins (Fratercula corniculata), and parakeet auklets (Cyclorrhynchus psittacula) from the Aleutians (G. J. Divoky and C. M. White, per- sonal communication), as well as in the stomachs of adult and nestling Leach's petrels from Newfoundland and New Bruns- wick (Rothstein 1973). Gulls and terns regurgitate indigestible parts of their food, such as bits of shell and fish bones. Polystyrene particles have also been found in these pellets (Hays and Cor- mons 1974). It is not known whether the birds ingest the plastic particles directly, but petrels appar- ently do. Other marine birds may acquire par- ticles in their stomachs by consuming fish that have previously ingested the plastic particles. Evidence of harmful effects of plastic par- ticles to any species is lacking, except for the possibility of intestinal blockage in smaller fish (Carpenter et al. 1972). However, they do accumulate in the environment, are eaten by fish, and are found in the stomachs of marine birds. It has been suggested that the plastics industry develop products that are degrad- able, but the most likely outcome of such an effort would be introduction of finished prod- ucts that would disintegrate into smaller particles similar to those described here (Hays and Cormons 1974). Rubber thread cuttings may represent a hazard to marine birds. Common puffins, in particular, appear to mistake them for fish and swallow them. These elastic threads form knots and the tangled mass may remain in the stomach. In one case the entangled elastic was tightly packed into the gizzard exit; in an- other it had formed a ball of rubber in the gizzard itself. Although the rubber threads may not kill the birds, there is a possibility that they make them less able to withstand other stresses (Par slow and Jefferies 1972). Although other artifacts, such as trash scattered on beaches or jetsam washed ashore, may contribute significantly to the mortality of certain species of marine birds (Gochfeld 1973), in other circumstances, such debris may enhance the habitability of an area. An apparent increase in the number of black guillemots breeding in the Barrow, Alaska, area appears to be associated with the local increase in man-made debris. The birds tjT)ically nest in cavities in rock cliffs and crevices in talus slopes. Because such nest sites are absent in the Barrow area, guille- mots have nested in an empty oil drum, under a collapsed building, and under other types of man-made debris (Divoky et al. 1974). No explanation has been found for the ap- pearance along the Northumberland (United Kingdom) coast of severely debilitated common murres whose plumage has been ex- tensively abraded. Fluoride, discharged by a nearby aluminum smelter, was considered a possible cause because the birds had a strong odor resembling chlorine, another closely re- lated halogen compound. There also were similarities between the signs observed in the affected birds and those observed in cases of acute or chronic fluorosis in other animals. The implication of fluoride was dismissed, however, in part, on the basis of low fluoride residue levels in bone, skin, internal organs, and digestive tract of the affected birds. Fur- ther, normal murre feathers were not dam- aged by soaking in various fluorine com- pounds, in samples of smelter effluent, and in undiluted scrubber liquid (Croxall 1972). Recommendations The levels of any pollutant, or combination of pollutants, in the marine environment should remain below a level that damages the viability of any population or species of ma- rine bird. Thus the global use of organo- chlorine compounds must be regulated, if nec- essary, to restrict input into the sea. The undersea exploitation of petroleum, the marine transport of petroleum, and the ac- tivities of coastal refining and petrochemical industries must also be regulated to prevent harm to local populations of marine birds. Much remains to be learned about the expo- sure of marine birds to environmental pollu- tants in northern North America. The most critical areas for study include the effects of chronic sublethal exposure to petroleum hy- drocarbons, certain organochlorines, and mer- cury. The possible synergistic effects of these compounds in marine birds should also be in- tensively studied. A long-term program to monitor increasing or decreasing levels of any particular poUu- 31 tant in the marine environment, with particu- lar reference to the levels that affect the most sensitive species of marine bird, is necessary. A portion of this program might be carried out by using the eggs of marine birds, because colonies of some species are large and eggs may be obtained on a regular basis. The variance of pollutant distributions and the mathematical nature of these distributions are imperfectly known and the statistics of sampling have not yet been adequately formu- lated. Moreover, it would be desirable to carry out such programs in conjunction with other programs that examine changes in pollutant levels in the marine environment Uke the "Mussel Watch" (Goldberg 1975), which is following changes in the levels of plutonium isotopes, petroleum compounds, chlorinated hydrocarbons, and selected metals in mussels from U.S. coastal localities. Priorities in future research might be given to more intensive studies within local areas to obtain a better understanding of the dynamics of pollutant accumulation by birds. Of primary concern is the need to determine whether petroleum compounds are accumu- lated in food webs, including marine birds, and whether such compounds exert dele- terious physiological effects. Because pe- troleum compounds seem to have longer-last- ing effects in colder water, the impending ex- ploitation of oil resources in the offshore and North Slope areas accentuates the urgent need for information on the environmental consequences of chronic as well as acute exposure. The environmental effects of small plastic particles that are commonly found in oceanic areas, including northern North America, should be investigated. The relationships between chronic exposure to environmental pollutants and other envi- ronmental stresses are relatively unknown, as are relationships and effects of pollutants on many of the essential organisms in the food webs upon which marine and estuarine birds depend. An annual symposium on the marine birds of northern North America should be held to serve as a forum for presentation of new infor- mation. 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GPO 845 - 756 As the Nation's principal conservation agency, the Department of the Interior has responsibility for most of our nationally owned public lands and natural res )urces. This includes fostering the wisest use of our land and water resources, protecting our fish and wildlife, preserv- ing the environmental and cultural values of our national parks and historical places, and providing for the enjoyment of life through out- door recreation. The Department assesses our energy and mineral resources and works to assure that their development is in the best interests of all our people. The Department also has a major rcsiMinsi- bility for American Indian reservation communities and for people who live in island territories under U.S. administration. UNITED STATES DEPARTMENT OF THE INTERIOR FISH AND WILDLIFE SERVICE EDITORIAL OFFICE AYLESWORTH HALL, C8U FORT COLLINS. COLORADO 80523 POSTAGE AND FEES PAID U.S. DEPARTMENT OF THE INTERIOR INT 423 NOTE: Mailing lists are computerized. 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