Effects of acidification on the availability of toxic metals and calcium to wild birds and mammals

Em.'ironmental Pollution 71 (1991) 329-375

Effects of Acidification on the Availability of Toxic Metals and Calcium to Wild Birds and Mammals

A. M. Scheuhammer Environment Canada, Canadian Wildlife Service. 100 Gamelin BIvd, Ottawa, Ontario, Canada K1A OH3 (Received 17 November 1989; accepted 11 May 1990)

ABSTRACT The effects of acidification on wildlife inhabiting aquatic or semi-aquatic environments are reviewed, with particular reference to the possibility .for increased dietary exposure to Hg, Cd, Pb and/or AI, and decreased availability of essential dietary minerals such as Ca. It is concluded that: (1) piscivores risk increased exposure to dietary methyl-Hg in acidified habitats, and Hg concentrations in prey may reach levels known to cause reproductive impairment in birds and mammals; (2) piscivores do not risk increased exposure to dietary Cd, Pb or AI because these metals are either not increased in fish due to acidification, or increases are trivial .from a toxicological perspective; (3) insectivores and omnivores may, under certain conditions, experience increased exposure to toxic metals in some acid(lied environments. Exposure levels are likely to be sufficiently low, however, that significant risks to health or reproduction are unlikely. More importantly, these wildl(l~" species may experience a drastic decrease in the availability of dietao, Ca due to the pH-related extinction of high-Ca aquatic invertebrate taxa (molluscs. crustaceans). Decreased availability of dietao' Ca is known to adversely qffect egg laying and eggshell integrity in birds, and the growth o[ hatchling birds and neonatal mammals. Acidification-related changes in the dietary availabili o, of other essential elements, such as Mg, Se and P, have not been established and require further investigation; (4) herbivores may risk hwreased exposure to A1 and Pb, and perhaps Cd, in acidified environments because certain macrophytes can accumulate high concentrations of these metals under acidic conditions. The relative importance of p H in determining the metal concentrations of major browse species, and the toxieological 329 Environ. Pollut. 0269-7491/91/$03"50 C~ 1991 Elsevier Science Publishers Ltd, England. Printed in Great Britain

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A.M. Scheuhammer consequences for herbivorous wildlife, is not well established and requires further study. A decreased availability of dietary Ca is also likely for herbivores inhabiting acidified environments.

INTRODUCTION Birds and mammals that inhabit aquatic ecosystems are unlike aquatic plants, invertebrates, fish, and amphibians in that they are not directly affected by acidic deposition. They are nonetheless vulnerable to changes in the structure of their habitats (Eriksson, 1987; Ormerod & Tyler, 1987; Schreiber & Newman, 1988; Diamond, 1989). The quantity and quality of their food resources can be of critical importance. ,Reduced fish abundance in low pH environments can have a detrimental influence on the foraging behavior and breeding success of loons (Gavia immer) and other species of piscivorous birds (McNicol et al., 1987a, b; Alvo et al., 1988; Parker, 1988). For non-piscivorous species, particularly those relying on aquatic invertebrates as a food base, food abundance may not be reduced at low pH (Eriksson, 1984; Glooschenko et al., 1986), rather invertebrate diversity is much reduced (McNicol et al., 1987b; Schindler et al., 1989). Insectivores breeding on acidified wetlands are nutritionally dependent on a relatively small number of acid-tolerant invertebrate taxa (McNicol et al., 1987b; McAuley & Longcore, 1988). In such cases, detrimental effects of acidification on the quality, rather than the quantity, of the available food items is of concern. There are two general ways by which food quality may be affected by acidification of environments, and in turn influence the health of wildlife inhabiting these environments: (1) The concentration or availability of some nutritionally essential factor(s) may be significantly decreased in response to increasing acidification. For example, the caloric content of invertebrates may be reduced at low pH (Raddum & Steigen, 1981). It has been argued that the bioavailability of the essential trace element selenium (Se) may be reduced in acidified environments (Mushak, 1985), although empirical evidence is lacking. Many invertebrate species that contain high concentrations of essential calcium (molluscs, crustaceans), and which are eaten by omnivorous and insectivorous birds, are highly sensitive to pH, being among the first taxa to disappear during the acidification of wetlands (Eilers et al., 1984; Okland & Okland, 1986; Schindler et al., 1989); (2) There may be an increase in one or more non-essential, potentially toxic substances as a consequence of acidification of habitats. Studies from Canada (Scheider et al., 1979; Wren & MacCrimmon, 1983; McMurtry et al., 1989; Suns & Hitchin, 1990), the US (Wiener, 1983), and Sweden (Bj6rklund et al., 1984)

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have demonstrated that the Hg content of several fish species is negatively correlated with lake pH. The form of Hg in fish is primarily methylmercury (MeHg), the most readily absorbed and most toxic of the organomercurials. Thus, in acidified habitats that still support fish communities, fish-eating birds and mammals may risk increased dietary MeHg exposure (Wiener, 1987; Spry & Wiener, 1991). Other potentially toxic metals (AI, Cd, Pb) can accumulate in invertebrates (Wren & Stephenson, 1991) and/or aquatic plants (Crowder, 1991) tolerant to acidification. In this review, the potential for increased exposure to metals (Hg, Cd, Pb, Al) in birds and mammals inhabiting acid-sensitive environments is critically evaluated, and an assessment is made regarding increased risk to health or reproductive success. A second section of the review discusses the possibility of substantial reductions in the availability of nutritionally essential minerals, particularly calcium (Ca), due to acidification of environments, and a third section is a case study of metal accumulation in young waterfowl hatched on wetlands demonstrating a range of pH and other water chemistry parameters.

TOXIC EFFECTS OF DIETARY METAL EXPOSURE The toxicology of chronic, dietary metal exposure in birds has been reviewed (Scheuhammer, 1987a), and others have published reviews of the biological effects of toxic metals in a variety of organisms (Doyle, 1977; Demayo et al., 1982; Eisler, 1985, 1987, 1988; Krueger et al., 1985; Wren, 1986). Here, the major effects of dietary exposure to Hg, Cd, Pb, and A1 on birds and mammals are briefly summarized, with the intention of identifying the lowest dietary and tissue concentrations which are known to be associated with significant toxic responses. This is an appropriate approach because, although environmental mobilization of toxic metals certainly occurs in response to acidification (Gilmour & Henry, 1991; Nelson & Campbell, 1991), the subsequent accumulation of metals by potential wildlife food organisms will likely not be of sufficient magnitude to pose a threat of acute toxicity (Crowder, 1991; Spry & Wiener, 1991; Wren & Stephenson, 1991). Since reproductive effects generally occur at lower dietary metal concentrations than those required to produce other signs of impairment in adult birds and mammals, the discussion will focus on reproductive effects, as well as certain other effects that are known to occur at relatively low levels of exposure. The main objectives of this section of this review are: (1) to answer the question 'What are the lowest dietary and tissue concentrations of a particular metal known to have significant toxic effects, and what are the specific impairments produced'?'; and (2) to use these concentrations as

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criteria for assessing tissue and dietary metal levels as actually measured in biota from acidified environments.

Mercury--toxicology Whereas the intestinal absorption of inorganic forms of rig is at most only a small percentage of the ingested dose, absorption of the organomercurials, particularly methylmercury (MeHg), approaches 100% (Berlin et al., 1975). Toxicologically, dietary MeHg concentration is a better indicator of potential health risks than is total Hg concentration (Eaton et al., 1980). For this reason, the present review will consider only the accumulation and toxic effects of MeHg. Ecologically, the feeding habits of the species will determine its relative risk to MeHg exposure. In general, carnivores accumulate more Hg than omnivores, which in turn accumulate more Hg than herbivores. Moreover, predatory species associated with aquatic foodchains accumulate more Hg than those linked primarily to terrestrial foodchains, and fish eaters accumulate more than species feeding mainly on insects or other invertebrates (Hesse et al., 1975; Lindberg & Odsjo, 1983; Doi et al., 1984; Wren, 1986; Braune, 1987). These trends are due largely to differences in the concentrations of readily absorbable Hg (MeHg) in the different food organisms. Fish muscle tends to contain a high proportion of total Hg as MeHg, on average > 50% and often close to 100%, whereas invertebrates such as insects, crustaceans (except crayfish) and molluscs typically contain < 50% MeHg, and plants may contain barely detectable levels of MeHg even in Hg-contaminated environments (Gardner et al., 1978; Hildebrand et al., 1980; Cappon & Smith, 1981). Amphibians have not been well studied with respect to MeHg accumulation (Freda, 1991) but available evidence indicates that, unlike fish, they are not an important link in the transfer of rig along foodchains (Dustman et al., 1972; Terhivuo et al., 1984). Clearly, piscivorous wildlife are at greatest risk from putative pH-related increases in the bioavailability of Hg (Gilmour & Henry, 1991; Spry & Wiener, 1991). These species include loons, mergansers, osprey, kingfishers, herons, eagles, otters and mink. The major effects of overt MeHg intoxication are neurological in nature. Among the earliest signs of dysfunction reported for humans and experimental mammals exposed to MeHg are impaired vision, muscle weakness, and clumsiness (Eaton et al., 1980; Annau & Eccles, 1987). These sublethal impairments could severely compromise the ability of visual predators to obtain food in the wild, resulting in emaciation and an increased susceptibility to disease or other environmental stresses. These secondary effects may thus be important causes of death even at exposure levels insufficiently high in themselves to cause mortality directly.

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Experimental studies have demonstrated that, in birds, the tissue-Hg concentrations which are associated with neurological impairment and death are frequently similar despite differences in species, body size, dietary Hg concentration, or length of time required to produce the effect (Fimreite, 1971; Gardiner, 1972; Stoewsand et al., 1974; Heinz, 1976; Finley et al., 1979; Bhatnagar et al., 1982; Scheuhammer, 1988a). Neurological signs (weakness, difficulty flying or walking, lack of coordination) are typically associated with Hg concentrations of 15 F~g/g (wet wt) or greater in brain, and at least 30/~g/g in liver or kidney in otherwise unstressed adult birds. Death can occur without further increases in brain-Hg levels, although hepatic and renal concentrations tend to be higher (>50/~g/g) in association with outright mortality. Tissue-Hg levels of a similar magnitude have been reported in association with MeHg intoxication in a variety of mammals (Wren, 1986; Wren et al., 1987). The lowest dietary Hg concentration known to produce obvious signs of MeHg intoxication and mortality in adult animals is approximately 5 ~tg/g dry wt (1-1.6 Itg/g wet wt) (Wobeser et al., 1976; Wren et al., 1987; Scheuhammer, 1988a). As a general rule, the dietary concentrations of MeHg that are required to produce significant reproductive impairment are about 1/5 those required to produce overt toxicity in adult birds of the same species. Similarly, tissue-Hg concentrations associated with significant reproductive impairment are typically much lower than those causing overt neurological impairment or death. Liver-Hg concentrations of 2-12/tg/g (wet w t ) i n adult breeding pheasants (Phasianus colchicus) and mallard ducks (Anas plao, rhynchos) were linked to decreased hatchability of eggs and increased hatchling mortality in the absence of toxic signs in the adults (Fimreite, 1971; Heinz, 1976). Overall reproductive success in birds can decrease by 35-50% due to dietary MeHg exposure insufficient to cause obvious signs of intoxication in adults (Heinz, 1974; Scheuhammer, 1987b). Brain-Hg concentrations as low as 3 7 lLg/g can be lethal to newly hatched ducklings, (Heinz & Locke, 1976: Finley & Stendell, 1978), whereas levels at least four times these values are required to cause direct mortality in adults of a variety of species (Stoewsand et al., 1974; Finley et al., 1979; Scheuhammer, 1988a). Table 1 summarizes the reproductive effects of dietary MeHg exposure in birds. Of those studies listed, only that of Barr (1986) was conducted under actual field conditions, and its findings are suggestive of significant behavioral effects of MeHg that may come into play at even lower levels of exposure than those associated with impaired hatching success and hatchling mortality. Loons were observed to lay fewer eggs, and desert nests more frequently when Hg concentrations in prey (fish and crayfish) averaged 0"3-0-4 #g/g wet wt ( ~ 1-2/~g/g dry wt) and there was a severe decline in the occupation of potential territories and almost no eggs were laid when prey Hg concentrations exceeded 0.4/tg/g (Barr, 1986). If these observations can

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TABLE l

Reproductive Effects of Dietary Methylmercury Exposure in Birds

Reference Scheuhammer (1987b)

Species Ring Dove

Hg in food (llg/g dry wt) 4-6

( - ) Fertility ( + ) Hatchling mortality

3

( - ) Egg laying ( - ) Hatched (+) Hatchling mortality

2-3

( - ) Fertility ( + ) Shell-less eggs ( + ) Embryonic mortality

l-2

( - ) Egg laying ( - ) Territory

(Streptopelia risoria)

Heinz (1974)

Mallard

(Anas platyrhychos)

Fimreite (1971)

Pheasant

(Phasianus colchicus)

Barr (1986)

Common Loon

Effectsa

(Gat~ia immer)

USC

a + Indicates significant increase, - indicates significant decrease.

be applied to piscivorous birds in general, then the primary effects of lowlevel, dietary MeHg exposure in free-living piscivores may be mediated primarily through effects on reproductive behavior rather than direct toxic effects on fertility, embryonic development or hatchling viability. Due to the efficient placental transfer of MeHg (Yang et al., 1972; Reynolds & Pitkin, 1975), mammals :also risk reproductive impairment due to dietary MeHg exposure. Hg concentrations in the fetal brain were up to twice as high as maternal brain levels after administration of MeHg to pregnant rodents (Yang et al., 1972). Reproductive defects in mammals span the range from fetolethality to subtle development changes which result in neurological and behavioral effects after birth (Burbacher et aL, 1984; Chang & Annau, 1984; Gunderson et al., 1986; Eccles & Annau, 1987). For adult humans, WHO (1976) has recommended that a Hg concentration in whole blood > 200 ~g/liter be considered as indicative of a significant risk for developing neurological impairment (constriction of visual field, lack of coordination, impaired speech). Due to the greater sensitivity of the unborn fetus to MeHg toxicity, it has been suggested that concentrations of Hg in maternal blood need only exceed 20-30 ~g/liter in order for there to be a

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significant risk of abnormal development (Kjellstr6m et al., 1986; Health and Welfare Canada, 1987). People for whom fish constitutes a major component of the diet should not consume fish containing > 0-2 #g/g (wet wt) Hg, if protection of the fetus is paramount (Health and Welfare Canada, 1987). This concentration is close to the threshold above which Barr (1986) noted significant reproductive effects in free-living loons.

Mercury--risks to wildlife Several studies from different geographical locations have shown that ambient pH and Hg in fish are negatively correlated (Spry & Wiener, 1991). In order to assess the possible effects that an increased Hg burden in fish may have on piscivorous wildlife, it is necessary to know the relationship between ambient pH and Hg accumulation in fish of a species and size range most likely to be utilized by particular species of wildlife. For example, loons prey predominently on fish weighing <300 grams (Barr, 1973, 1986; Parker, 1988). In south-central Ontario, small (20-24cm) walleye (Stizostedion ritreum) from high-alkalinity lakes ( > 1 5 m g CaCO3/liter) had Hg concentrations < 0"2 pg/g, whereas those from low-alkalinity lakes had Hg concentrations in the range of 0.3-0"6#g/g (Scheider et al., 1979). A regression equation presented in Suns and Hitchin (1990) indicates that, even in very small fish (yearling perch [Perca.[tavescens];7-10 cm, 4-10 g) from this same region, Hg concentrations approach 0.2#g/g when lake pH is < 5.5. As these fish grow, their Hg concentrations will likely increase. A Finnish study found a highly significant positive correlation between Hg concentrations in perch of a standardized length (20 cm) and lake-water Al: moreover, A1 was the only water chemistry parameter demonstrating a significant correlation with fish-Hg (MetsfilS, & Rask, 1989). A regression equation presented by Metsfilfi and Rask (1989) indicates that perch accumulate > 0"3/tg Hg/g when lake A1 concentrations exceed 100/tg/liter. Most lakes with such high concentrations of A1 are acidic. In Ontario, A1 is a good indicator of watershed acidification because it occurs naturally at low levels in lakes on the Precambrian Shield, yet is readily mobilized from soil and bedrock by acidic precipitation (Scheider et al., 1979). For northern Ontario lakes studied by the Canadian Wildlife Service, the majority of lakes with [A1] > 150#g/liter have a pH < 6.5 (McNicol et al., 1987a). Sunfish (Lepomis gibosus) from central Ontario lakes generally accumulated 0'2 0"3 ~g/g Hg when pH was < 6"5, but had < 0"2/~g/g when pH was > 6"5 (Wren & MacCrimmon, 1983). Mink collected from this same study area had liver-Hg concentrations comparable to those of mink from the Hgcontaminated English River in Ontario (Wren et al., 1986), the river system where Barr (1986) reported reproductive impairment in loons when Hg

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concentrations in their prey exceeded 0-3/~g/g. Thus, there is evidence linking environmental acidification with Hg levels in fish sufficient to be associated with toxic effects in piscivorous wildlife. Unfortunately, studies designed specifically to determine Hg-related reproductive effects in piscivorous birds and mammals living in acidified habitats have not been undertaken, nor have specific surveys to determine Hg concentrations in prey of an appropriate size and species composition. In addition to more extensive surveys of rig concentrations in appropriate prey items, indicators of exposure to MeHg in loons and other piscivorous wildlife should be incorporated into existing or future acid-rain biomonitoring programs. In order to assess risk, such indicators need to be accurate and dose-related. Feathers and eggs of birds, and the fur of mammals, are good integrators of dietary MeHg exposure, the collection of which does not require the death of the animals being sampled. Feathers may provide a convenient substrate for monitoring for increased dietary exposure to MeHg in birds. Hg is deposited in feathers mainly during their formation, and once bound to the feather matrix, Hg is resistant to leaching (Applequist et al., 1984; Goede & deBruin, 1984). In addition, Hg does not concentrate in the uropygial gland or in the secreted oils, thus contamination of feathers by this route after their formation is probably minimal (Frank et al., 1983). The amount of Hg deposited in feathers, as in other tissues, varies according to the dietary habits of the species in question. In a study of the Hg-contaminated Shiranui Sea by Doi et al. (1984), fish-eating seabirds had the highest feather-Hg levels (--,7/lg/g) and omnivorous waterfowl had higher levels than herbivorous waterfowl (,~ 5/lg/g and --~1/lg/g, respectively). A number of studies have indicated that the normal, background levels of rig in feathers of raptorial birds living in relatively uncontaminated habitats ranges from 1-5/lg/g (Berg et al., 1966; Lindberg & Mearns, 1982; Parrish et al., 1983; Honda et al., 1986a). Variation within this range may be due to geographic variation in the natural background levels of Hg in soils and sediments, variation in the rates of environmental methylation of Hg, and differences in dietary habits. Feather-Hg levels of birds experimentally dosed with dietary MeHg exceeded levels accumulated in other tissues such as liver and kidney by a factor of at least 4 (Heinz, 1976, 1980; Stickel et at, 1977; Finley & Stendell, 1978), indicating that Hg has a high affinity for proteins of the growing feather. In free-living birds, it has been estimated that 60-90% of the total body burden of Hg may be present in the plumage (Braune & Gaskin, 1987; Honda et al., 1986a, b). A similar tendency for Hg to accumulate in growing hair and fur of mammals has been noted (Wren, 1986). The efficiency with which Hg is deposited in feathers, the correlation between feather-Hg and dietary Hg concentrations, the high stability of Hg

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in feathers and the relatively convenient and non-lethal nature of sampling feathers makes them good potential indicators of the bioavailability of Hg to birds. In order to assess whether piscivorous birds breeding in acidified environments are exposed to higher Hg levels than those living in circumneutral habitats, the Hg concentration in feathers of fledglings should be of value. These feathers would develop while the birds were in the natal habitat and should thus reflect the Hg concentration of food items eaten during the growth of the feather. Based on experimental dose-response studies, growing piscivorous birds consuming diets containing Hg concentrations > 1/tg/g (dry wt) as MeHg should accumulate > 20/~g/g (dry wt) total Hg in primary feathers (Finley & Stendell, 1978; Heinz, 1976, 1980; Scheuhammer, 1990). Such concentrations should be considered as indicative of a wetland habitat that may pose a significant threat to the reproductive success of piscivorous wildlife breeding there. However, if loons and other piscivorous birds typically avoid these wetlands, abandon territories, or fail to lay or hatch eggs, then a monitoring scheme based on the Hg content of fledgling birds will be of limited use. Hg is present in the eggs of birds, accumulating particularly in the albumen portion after exposure to MeHg (Backstr6m, 1969; Scheuhammer. 1987b). As is the case for feathers, Hg accumulates in eggs in a dosedependent fashion in response to increasing dietary levels of MeHg (Tejning, 1967; Heinz, 1976; March et al., 1983). Scheuhammer (1987b) noted a biomagnification factor of approximately 5-6 when comparing a range of dietary Hg concentrations with that deposited in albumen of eggs (on a dry weight basis) of ring doves experimentally exposed to MeHg. If this relationship is generally valid, then Hg concentrations > 9/~g/g in dry egg albumen would be indicative of a level of dietary MeHg exposure known to cause reproductive effects in loons (Barr, 1986). On a wet weight basis, whole loon eggs collected from Hg-contaminated sites in which loons demonstrated the lowest reproductive success typically had > 1/~g/g Hg (Barr, 1986), which corresponds to about 4 5 #g/g dry wt. Thus, the Hg content of eggs presents yet another possible indirect measure of dietary Hg levels and may be used to identify lakes and wetlands on which piscivorous wildlife may suffer reduced reproductive success. In order to make the association between Hg in eggs and high-risk habitats, it is essential to have a detailed understanding of the foraging behavior of adult females during and immediately prior to egg formation.

Lead--toxicology Although organolead species are known to be highly toxic (Grandjean & Grandjean, 1984), these forms of Pb, primarily tetraethyl- and tetramethylPb, are almost exclusively of industrial origin. There is little evidence to

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suggest that the alkylation of Pb occurs to any significant extent in natural ecosystems, or that environmental acidification affects production of alkylPb species. The following discussion, therefore, focuses exclusively on the toxicology of inorganic Pb complexes. In adult mammals, the gastrointestinal absorption and retention of soluble inorganic Pb salts from dietary sources is typically 5-10% of the ingested dose (Kehoe, 1961; Van Barneveld & Van den Hamer, 1985) and there is no evidence to suggest that Pb absorption in birds is quantitatively less or greater than in mammals. In immature animals, as well as human children, Pb absorption from food is considerably higher than in adults (30-40% of dose) (WHO, 1986). Absorbed Pb accumulates in greatest concentration in bones of mammals and birds. Skeletal Pb accounts for > 90% of the total body burden (Stowe et al., 1973; Barry, 1975). Of the soft tissues, kidneys accumulate the highest levels (Stowe et al., 1973; Custer et al., 1984). Whereas Pb in bone is comparatively stable, with a half-life of several years, liver and kidney Pb have a half-life of about 3-4 weeks. For this reason, the Pb concentration of blood, liver, or kidney indicates the degree of recent exposure, whereas bone-Pb reflects cumulative exposure. The absorption and retention of Pb from the diet can be greatly affected by nutritional factors. Depressed levels of dietary protein or Ca are particularly effective in increasing the uptake and toxicity of Pb in both birds and mammals (Six & Goyer, 1970; Stowe et al., 1973; Carlson & Nielson, 1985; Van Barneveld & Van den Hamer, 1985; Barton et al., 1988). The major toxic effects of Pb are haematological (Moore & Goldberg, 1985), neurological (Moore et al., 1986; Bondy, 1988), and renal (Goyer, 1985; Bernard & Becker, 1988). Signs of overt Pb poisoning in adult mammals and birds include lethargy, loss of appetite, green watery feces (birds), anaemia, weakness, emaciation, impaired body movement and coordination, tremors, blindness, and death. Typically, the effects of Pb poisoning are associated with tissue-Pb concentrations > 80 #g/dl in blood, > 10#g/g (wet wt) in liver and (or) >20#g/g in kidney (Cook & Trainer, 1966; Goyer & Rhyne, 1973; Stowe et al., 1973; Longcore et al., 1974; Benson et al., 1976; Forbes & Sanderson, 1978; Osweiler & Van Gelder, 1978; Macdonald et al., 1983; Beyer et al., 1988; O'Halloran et al., 1988). Reproductive impairments due to Pb exposure can occur at lower levels of exposure than those required to produce overt intoxication in adults. Of particular interest in recent years has been the elucidation of subtle developmental effects of Pb in fetal and newborn humans and experimental mammals. Major effects include lowered birth weights, hyperactivity, and impaired neurobehavioral development (Cory-Slechta et al., 1985; Yule & Rutter, 1985; Donald et al., 1986a, b; Davis & Svensgaard, 1987; Gilbert & Rice, 1987; Needleman, 1989). Some of the neurobehavioral effects can occur

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at blood-Pb concentrations (10-20#g/dl) long thought to be within the 'normal' range. The relevance for wild mammal (or bird) populations of such subtle neurotoxicological effects has not been addressed. Adults of various avian species require relatively high concentrations of inorganic Pb in the diet (> 800/tg/g wet wt) before mortality is observed (Beyer et al., 1988). Similarly, juvenile precocial birds are relatively resistant to the toxic effects of dietary Pb ingestion. Growing quail (Coturnix coturnix) maintained on diets containing 500 or 1000 ktg/g, and cockerels fed a diet containing 1850#g/g Pb, developed mild anaemia and had slower growth rates than controls, but mortality was not increased (Morgan et al., 1975; Franson & Custer, 1982). Young altricial birds are more sensitive to dietary Pb. Diets containing the equivalent of 125 Itg Pb/g (dry wt) caused significant reductions in growth rate and brain weights in nestling kestrels (Falco sparverius) (Hoffman et al., 1985), and 84-94/~g/g caused anaemia and reduced brain weights in nestling starlings (Sturnus vulgaris) (Grue et al., 1986). Lead--risks to wildlife

There is no evidence to suggest that fish (Spry & Wiener, 1991, amphibians (Freda, 1991), or invertebrates (Wren & Stephenson, 1991) from acidified environments can accumulate Pb to levels sufficient to cause death or obvious signs of Pb intoxication in wildlife using these organisms as food. Should dietary Pb exposure be increased due to environmental acidification, toxic effects would likely be mediated primarily through impairment in the reproductive process. Even fairly subtle reproductive effects require dietary Pb concentrations of at least 100/~g/g (dry wt) in order to become manifest. Breeding kestrels fed 50 ~g/g dietary Pb demonstrated no impairment with respect to egg laying, incubation behavior, fertility, or eggshell thickness (Pattee, 1984). Nestling kestrels required the equivalent of 125~tg/g Pb in the diet in order to suffer reduced growth rates, and 625 #g/g was needed to increase mortality (Hoffman et al., 1985). Wild starlings nesting within highway verges were exposed to dietary Pb concentrations of 84-94/tg/g (dry wt), yet exhibited no decline in clutch size, number of young hatched, or number of young fledged compared to a control colony whose dietary Pb intake was approximately 1/10 that of birds along the highway: however, nestling hematocrits and brain weights were lower in the highway colony (Grue et al., 1986). In experimental rodents, 250/tg/ml in drinking water produced subtle changes in social behavior, but this level of Pb caused no adverse effects on birth rates, birth weights, or overall pup development (Donald et al., 1986b). Such levels of dietary Pb are the lowest concentrations known to cause significant effects in experimental birds and mammals, yet

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Scheuhammer

'

they are still 1-2 orders of magnitude higher than the concentrations of Pb actually measured in most potential wildlife prey items from acidified environments (Spry & Wiener, 1991; Wren & Stephenson, 1991; see also Table 3). Field studies have failed to indicate a potential problem with regard to Pb exposure in wildlife living in acidified habitats. Tissue concentrations of Pb in young waterfowl from acidified lakes in Sweden (Eriksson et al., 1989) were 1-2 orders of magnitude lower than those known to be associated with impaired growth or mortaility in young kestrels (Hoffman et al., 1985), and there was no tendency for tissue-Pb concentrations to be increased in waterfowl from acidified lakes (Eriksson et al., 1989). Likewise, average hepatic Pb concentrations in otter and mink from acidified habitats were toxicologically low ( < 0"5 pg/g wet wt) and were not related to the degree of environmental acidification (Wren et al., 1988). Certain species of submerged aquatic plants can accumulate > 100/~g/g Pb (dry wt) in acidified lakes and ponds (Sprenger & McIntosh, 1989). Dietary levels of this magnitude are capable of exerting toxic effects. It is known from documented outbreaks of Pb poisoning in domestic cattle (Kwatra et al., 1986) and wild swans (Benson et al., 1976) that Pb poisoning can occur in birds and mammals due to chronic ingestion of plant material contaminated with Pb from industrial sources. Thus, although the available evidence from toxicological studies and from field studies comparing acidified and circumneutral environments, indicates that environmental acidification does not increase the risk for Pb-induced health or reproductive effects in most wild birds and mammals, herbivores, particularly those associated with aquatic environments, may be an exception. Unfortunately, empirical data relating to this latter possibility are lacking.

Cadmium--toxicology In response to low, environmentally relevant doses ( < 1 mg/kg body wt), the uptake and retention of Cd from the diet is 0.1-1.0% of the ingested dose in mammals and birds (Engstr6m & Nordberg, 1979; Lehman & Klaassen, 1986; Scheuhammer, 1988b). Injury to the intestinal epithelium, for example by infection or by the direct toxic action of Cd itself, causes increased absorption of Cd (Richardson & Fox, 1974; Bafundo et al., 1984; Scheuhammer, 1988b). Similarly, diets deficient in Ca, Zn or Fe facilitate the intestinal uptake and toxicity of Cd (Hill et al., 1963; Bunn & Matrone, 1966; Banis et al., 1969; Washko & Cousins, 1976, 1977; Flanagan et al., 1978; Koo et aL, 1978; Waalkes, 1986). Most ( > 8 0 % ) of the body burden of Cd is localized in the liver and kidneys where it accumulates through time bound to a low molecular weight

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(6500 daltons), sulfydryl-rich protein, metallothionein (MT) (Hamer, 1986; Dunn et al., 1987). The Cd-MT complex is very stable, resulting in a biological half-life for Cd on the order of several years and a demonstrated tendency for Cd to accumulate with age in a variety of mammal and bird species (Hutton, 1981; Woolfet al., 1982; Eisler, 1984; Norstrom et al., 1986; Blomqvist et al., 1987; Cr~te et al., 1987, 1989; Glooschenko et al., 1988). Since the rate of excretion of biologically incorporated Cd is slow, mammals and birds consistently accumulate Cd in target organs (liver and kidney) in excess of the Cd concentration in their food supplies (Leach et al., 1979: Nomiyama et al., 1979; Hunter & Johnson, 1982; Cain et al., 1983; Scheuhammer, 1987b). The kidney is the critical organ in chronic Cd toxicity. Through time, Cd and MT concentrations of the renal cortex increase. When the Cd concentration reaches 100-200/tg/g (wet wt) in adult humans, other mammals and birds, the ability of MT to protect cells from the toxic influence of Cd begins to decrease and nephropathy characterized by proximal tubular cell necrosis, proteinuria, glycosuria, increased urinary Cd, decreased Cd content in the kidney and the appearance of MT in the plasma occurs (White et al., 1978; Goyer et al., 1984; Kjellstr6m et al., 1984; Kjellstr6m, 1986). The time required before kidney toxicity becomes apparent will vary according to the Cd content of the diet, but even low levels of dietary Cd are able to produce kidney toxicity given that the duration of exposure is sufficiently long. For humans, epidemiological studies indicate that 10% of a population exposed to an average of about 200/~g Cd/day via food is likely to develop renal damage (proteinuria) by age 45 (Kjellstr6m, 1986). This would correspond to an average Cd concentration in food <0.5/~g/g (dry wt). Long-lived species are therefore at greatest risk from chronic, low-level dietary Cd exposure. While a diet containing 200#g Cd/g (dry wt) was required in order to produce kidney lesions in adult mallards within 8.5 weeks (White et al., 1978), ducklings fed diets containing only 20/~g/g developed kidney lesions and anaemia within a similar timeframe (Cain et al., 1983). Adult mallards fed diets with 20/tg/g Cd accumulated less Cd than ducklings fed the same concentration for a comparable length of time (White & Finley, 1978; Cain et al., 1983). Such findings provide evidence that immature birds are more susceptable than adults to Cd accumulation and toxicity, and can develop renal dysfunction at renal-Cd concentrations < 100 #g/g (wet wt). Diets containing 75 #g C d / g produced a variety of toxic effects in young Japanese quail (Richardson et al., 1974). Many of the effects mimicked Fe deficiency (anaemia, bone-marrow hypoplasia, cardiac hypertrophy) or Zn deficiency (testicular hypoplasia). Histologically, extensive damage to the absorptive epithelium of the duodenum was observed, and judged to be the

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primary cytotoxic action of dietary Cd. Disturbances of Fe, Zn, and Ca metabolism by chronic dietary exposure to Cd can occur in the absence of Cd-induced renal dysfunction (Richardson et al., 1974; Watanabe et al., 1986). Dietary levels of Cd of 200 #g/g (dry wt) over a 90 day period caused a decline in egg production in mallards (White & Finley, 1978), as was also the case for chickens fed 60 #g/g Cd (Sell, 1975) or 48/~g/g Cd (Leach et al., 1979). Some eggshell thinning was observed in laying hens fed 48/~g/g (Leach et al., 1979). High dietary exposure to Cd from industrial contamination has been reported in certain wildlife species. For instance, in white-tailed deer (Odocoileus virginianus) living within 8 km of a zinc smelter, fecal Cd concentrations (an indicator of average dietary concentrations) were 2-71/~g/g (dry wt), while corresponding kidney concentrations were 22-400/~g/g (Sileo & Beyer, 1985). Concentrations in the upper end of these ranges are sufficiently high to be of toxicological significance. Even more intriguing are reports of elevated tissue-Cd concentrations which are not associated with obvious anthropogenic activity, yet are still high enough to be of toxicological concern. For example, kidney-Cd concentrations up to almost 100pg/g (wet wt) have been reported in northern fur seals (Callorhinus ursinus) from Alaska (Goldblatt & Anthony, 1983) and concentrations as high as 146 ~g/g have been found in ringed seals (Phoca hispida) from western Greenland (Johansen et al., 1980). Pelagic seabirds generally feed in areas removed from sources of industrial contamination, yet they frequently have higher tissue-Cd concentrations than other aquatic birds. In one study, oceanic great skuas (Catharacta skua) were found to have an average kidney-Cd concentration of 81 #g/g (dry wt) compared with 13 #g/g for coastal herring gulls (Larus argentatus) (Hutton, 1981). Cd in kidney tissue from pelagic fulmars (Fulmarus glacialis) ranged 116-480 #g/g dry wt (approximately 25-110/~g/g wet wt) (Osborn et al., 1979). A similar range of kidney-Cd concentrations was reported for antarctic Adelie penguins (Pygoscelis adeliae) (Honda et al., 1986b). These Cd levels are much higher than concentrations typically reported for wild birds and mammals, which generally average < 12/~g/g dry wt, or < 3 ~g/g wet wt, in kidney (Di Giulio & Scanlon, 1984; Wren, 1984; Leonzio et al., 1986; Blomqvist et al., 1987; Wren et al., 1988). High levels of Cd in tissues of wild vertebrates are almost certainly due to transfer from food items. In the marine environment, various invertebrate species have been shown to accumulate naturally high concentrations of Cd. For example, the hyperiid amphipods, particularly Parathemisto libellula, from marine waters of northern Canada contained an average of about 12/~g/g (dry wt), and may contain up to 40 ~g/g Cd (Macdonald & Sprague,

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1988). These invertebrates are an important food source for arctic seals at certain times of the year. Seaskaters (Halobates micans), wingless oceanic insects that live exclusively on the water surface, were sampled from numerous locations far from sources of industrial Cd in the tropical and subtropical Atlantic, and were found to contain an average of 20-25/lg Cd/g (dry wt) (Bull et al., 1977; Schulz-Baldes, 1989). Dietary Cd concentrations of this magnitude are not trivial, and may well be of toxicological significance, particularly to long-lived predators that have many years to accumulate Cd in target organs. Cadmium--risks to wildlife

Although Cd concentrations can be elevated in certain fish (Spry & Wiener, 1991) and aquatic invertebrates (Wren & Stephenson, 1991) due to acidification, the relationship with pH is neither simple nor strict; other factors are of equal or greater consequence than pH in determining Cd concentrations. For instance, in freshwater benthic a m p h i p o d s (Hyalella azteca), three independent parameters of lake water chemistry accounted for 81% of the variability in Cd concentrations: [Ca 2 +], total Cd in water, and dissolved organic carbon (Stephenson & Mackie, 1988). Although lake-Ca 2+ was strongly related to lake-pH, the other two parameters were not. Furthermore, when pH-related increases in the Cd concentrations of freshwater fish or invertebrates occur, the magnitude of the increase is generally not great enough to be toxic to wildlife that feed on these organisms. For instance, Cd concentrations in whole bluegills from north central Wisconsin increased from < 0-1 to c. 0.5/~g/g (dry wt) as pH declined from 7-5-5"1 (Wiener, 1983), but even the higher concentrations are 1 2 orders of magnitude lower than those known to be toxic. In general, piscivorous wildlife are not at risk for Cd-induced toxicity due to environmental acidification. This is also true for insectivores. Cd in aquatic insects is generally not elevated in acidified environments and concentrations are usually below those known to have toxicological significance (Wren & Stephenson, 1991). Similarly, there is not a clearcut relationship between declining pH and Cd accumulation by aquatic plants, and although concentrations can be high, the relative importance of acidification has not been well established (Crowder, 1991). Cd accumulation in wildlife tissues in relation to environmental pH has not been extensively studied. Published reports that have, in part, addressed this question have found some evidence that Cd accumulation is greater in acid-sensitive habitats than in well-buffered regions (Froslie et al., 1986; CrSte et al., 1987; Glooschenko et al., 1988; Wren et al., 1988). Recently, Cd concentrations in aquatic and terrestrial browse plants of moose and deer

344

A. M. Scheuhammer

have been shown to be higher in a non-buffered, acid sensitive environment than in a well-buffered, circumneutral environment (Hickie et al., 1989; Parker, 1989). However, absolute concentrations of Cd in plants eaten by deer and moose, even in those species considered to be accumulators (certain maples, aspen, and willow) and in fecal pellets, were not toxicologically high ( < 2 #g Cd/g) (Hickie et aL, 1989; Parker, 1989). In addition, other factors, such as proximity to sources of industrial pollution, degree of natural Cd enrichment of soils and bedrock, and forage-plant species composition, may be of equal or greater importance than environmental pH as determinants of Cd concentration in plants (Cr&e et al., 1987; Glooschenko et al., 1988; Kronberg et al., 1989). This does not mean that the possible effects of changes in environmental pH on Cd accumulation by plants should be ignored, but rather that pH effects may be difficult to clarify. Carnivores generally accumulate less Cd in target organs (liver and kidney) than herbivores because of the higher concentration of Cd in plants than in animal flesh. Mink and otter from acid-sensitive areas in Ontario had kidney-Cd concentrations that are not sufficiently high to be of toxicological concern (0-6-2.0/~g/g wet wt; Wren et al., 1988). Much higher concentrations have been reported for adult moose and deer in Ontario and Quebec (Cr~te et al., 1987; Glooschenko et al., 1988), but even the highest concentrations (c. 50 ~tg/g wet wt in moose from Algonquin, Ontario) are not high enough to be toxic to the animals themselves, based on the known toxicology of Cd in humans and experimental mammals for which a renal threshold of > 100 pg Cd/g (wet wt) is required before pathological consequences occur. Rather, the major concern is for increased intake of Cd by human consumers of these tissues. (The public has been warned against consuming liver and (or) kidney from moose, deer and (or) caribou in Ontario, Quebec, Newfoundland, New Brunswick and Manitoba). Similarly, insectivorous goldeneye ducklings collected on acidic lakes in Sweden did not accumulate toxicologically high kidney-Cd concentrations, nor were levels higher than those of ducklings from circumneutral lakes (Eriksson et al., 1989). Due to the inconclusive evidence linking declining pH to increased Cd concentrations in wildlife prey items, the general lack of toxicologically high Cd concentrations in prey and the probable importance of other factors that may be unrelated to changes in environmental pH, it cannot be concluded that the potential for Cd-induced effects on the health or reproduction of wild birds and mammals living in acidified habitats is great.

Aluminum--toxicology The gastrointestinal absorption of all forms of dietary A1 is low and the urinary excretion of excess A1 by animals without kidney dysfunction is

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efficient. Very high oral doses of AI salts are thus required to demonstrate significant deposition of A1 in tissues. For instance, in male and female ring doves (Streptopelia risoria) fed a nutritionally adequate diet supplemented with 0' 1% (1000/~g/g dry wt) A13 +, the only tissue to exhibit a clear elevation in A1 concentration was bone from reproductively active females (Carri6re et al., 1986). For this reason, it is difficult to assess the degree of A1 exposure simply by measuring the metal in those tissues most commonly used for contaminant monitoring (liver, kidney). Concentrations of A1 in bone > 10 kLg/g(dry wt) are suggestive of higher-than-normal AI exposure and/or compromised renal function. However, elevated tissue-A1 concentrations, if they are encountered, cannot be considered as necessarily indicative of AI toxicity. The toxicity of dietary A1 is primarily a function of the ability of A1 to disrupt the dietary absorption and normal metabolism of Ca and P, resulting in bone abnormalities and impaired growth in animals (Deobald & Elvehjem, 1935; Street, 1942; Storer & Nelson, 1968). Similarly, in humans, the elevated ingestion of Al-containing antacids may cause bone pain, weakness, and impaired bone mineralization (osteomalacia) (Insogna et al., 1980). The principal mechanism by which these effects are mediated is the formation of insoluble aluminum phosphate in the gut, resulting in a reduced absorption of P from food, and a condition mimicking P deficiency. Generally, even if dietary P is at a slightly suboptimal level, dietary concentrations of soluble A1 must be at least 50% of dietary P in order for adverse effects to occur. At a dietary P concentration of 0"5%, egg production in chickens declined when the feed was supplemented to 0"3°/,, A1, but not 0"2% Al, (Hussein et al., 1986). Reproductively active ring doves fed a diet containing a high P:AI ratio (0"5% P, 0.1% AI) reproduced normally, and a diet containing 0-15% (dry wt) AI fed to juveniles produced no growth impairments (Carri6re et al., 1986). Young rats exhibited decreased growth rates when the molar ratio of A1: P in the diet was 1: 1 or higher, but not when it was 1:1.5 or smaller (Street, 1942). In rabbits. approximately equal dietary concentrations of A1 and P (0"15%) were required to cause blood and bone P levels to decline (Cox et al., 1931). An interesting study by Sparling (1990), in which growing ducklings were subjected to various regimens of dietary Ca, P and A1, is deserving of special comment. Day-old black and mallard ducklings were randomly assigned to a dietary group: low Ca (0"36%)/low P (0-62%) ILL]; normal Ca (l51%)/normal P (1"35%) INN]; or low Ca (0"36%)/high P (2"1%) [LH]. Each Ca/P group was further subdivided into 3 dietary A1 levels (200 [control], 1000 and 5000/tg/g). Ducklings were provided with food and water adlib, for 10 weeks. For mallards on the NN diet, the addition of dietary AI up to 5000~g/g (0.5%) caused no increase in mortality, nor were there

346

A. M. Scheuhammer

adverse effects on growth and the average food consumption over the 10 week period was high ( > 70 g per duckling per day). Similarly, lowering Ca and elevating P (LH) produced no mortality up to 0.5% AI, and food consumption was high. Lowering both Ca and P (LL) caused no increase in mortality, nor was food consumption depressed, if diets were not supplemented with A1. However, the LL diet with 0.1% A1 caused a 40-50% decline in average food consumption accompanied by 61% mortality; at 0"5% A1, mortality rose to 100% and food intake fell to 10-20% of normal. These data indicate that the appearance of detrimental effects on growing ducklings due tO high dietary A1 are largely dependent upon the corresponding level of dietary P, a finding that is consistent with earlier reports (Cox, 1931; Deobald & Elvehjem, 1935; Street, 1942; Carri6re et al., 1986), and that these effects may be at least partially mediated by inappetance. In the study by Sparling (1990), black ducklings were more sensitive than mallards to manipulations of dietary Ca and P. They suffered 22% mortality when Ca alone was lowered and P elevated (LH) and 50% mortality when both Ca and P were lowered (LL), even though food intake was high. The increased mortality on the LL diet was exacerbated by the addition of A1 to this diet. At 0"1% A1, food consumption dropped by 40%, and mortality increased to 73 %. However, if P was elevated and Ca was reduced (LH), mortality did not increase with increasing dietary A1 concentrations. These results again indicate the importance of low dietary P for the expression of the adverse effects of elevated dietary A1. The LL and LH diets also caused depressed growth rates in surviving black ducklings, an effect that was exacerbated by high dietary A1. Lower food consumption probably contributed substantially to the decreased growth observed in the Sparling (1990) study (Fig. 1), and the very high mortality rates of ducklings 900 800 On

v

700 600

"D

500 400 300 200 40

50

60

70

80

90

Daily Food Intake ( g )

Fig. 1. Positiverelationship betweenaveragedaily food intake of survivingblack ducklings and body weight at 5 weeks of age. Means + SD of 4-10 values/data point are plotted. V, Normal Ca and P; Q, low Ca and P; m, low Ca/elevatedP. See text for actual dietary Ca and P concentrations. Within each dietary Ca/P regimen, the lowest food intake and lowest body weights correspond to the highest dietary AI concentration. Adapted from Sparling (1990).

Acid(fication, metals and wildl(['e

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on the most toxic diets (LL-0.5% A1, N N - I ' 0 % A1) were probably due to starvation. A1 is known to be neurotoxic under certain conditions of exposure, for instance, in relation to dialysis encephalopathy in humans (Parkinson et al., 1981) or following direct intracranial injection of AI salts in experimental animals (Petit et al., 1980). A1 neurotoxicity will not be discussed in the present review, however, because, for wildlife species inhabiting acidified environments, the only significant route of exposure to A1 is through the diet. Neurotoxicity resulting from dietary exposure to A1 has not been demonstrated. Aluminum--risks to wildlife

Assessing the risk to wild animals of possible increased dietary A1 exposure in acidified habitats requires knowledge of both dietary AI and P concentrations. However, in food organisms that contain near-optimal, or at least nutritionally adequate, levels of P ( > 0.4%), it is highly unlikely that A1 concentrations could rise to levels that would pose a significant health threat, based on the known toxicology of dietary A1. The flesh of fish, mammals and birds contains low concentrations ( < 50 #g/g dry wt) of AI (Wren et al., 1983; Spry & Wiener, 1991), therefore piscivorous wildlife and other top predators should not be at risk for increased A1 exposure. On the other hand, available evidence indicates that, as a group, aquatic invertebrates can accumulate higher concentrations of AI (Wren & Stephenson, 1991). Two British studies have reported that several orders of aquatic insects including chironomid midges, caddisflies, stoneflies and mayflies exhibit AI concentrations of 0-1-0-3% (dry wt) (Sadler & Lynam, 1985; Ormerod et al.; 1988). Insectivorous birds that feed in aquatic environments may thus be exposed to dietary AI levels known to cause significant biological effects, given that corresponding dietary P levels are low. Phosphorus concentrations in the insects cited above, however+ averaged at least 5-10 times the A1 content; moreover, no strong relationships were observed between AI concentrations in the insects and environmental [H +] over a pH range of about 5-8 (Sadler & Lynam, 1985; Ormerod et al., 1988). Thus, it appears that A1 concentrations in aquatic invertebrates can be high even under normal, circumneutral conditions and that there is not a dramatic increase in AI accumulation by aquatic invertebrates as pH declines, even though very substantial pH-related increases in dissolved AI concentrations have been documented (Nelson & Campbell, 1991). Indeed, there is evidence to suggest that invertebrateAI is lower under acidic conditions than under circumneutral conditions (e.g., Hall et al., 1988). Available evidence indicates that invertebrate-P

348

A. M. Scheuhammer

concentrations are generally high enough to offset any potential interference by dietary A1 in normal P metabolism. Nyholm and Myhrberg (1977) and Nyholm (1981) reported severe eggshell defects, reduced clutch sizes, and mortality in pied flycatchers (Ficedula hypoleuca) and other insectivorous passerines nesting by the shore of an acid-stressed Swedish lake. The reproductive impairments were not observed in birds whose breeding territories were removed from the lake shore. Of a host of potential contaminants (DDT, PCB, A1, Cd, Cr, Cu, Pb, Hg) measured in various tissues, only A1 in medullary bone samples from nesting females was found to be elevated in affected birds. This result prompted the speculation that high dietary AI intake may have played a causative role in the production of the observed reproductive impairments. Emergent insects used as food by the flycatchers were subsequently found to contain AI concentrations of 0"01--0-12% body weight (Nyholm, 1982). Assuming a normal or even reduced P content (which was not measured), these levels of A1 are not high enough to pose a significant threat to reproductive success, based on the known toxicology of dietary A1. However, unless the dietary Ca and P availability can be determined for the shore-nesting passerines studied by Nyholm, the possibility for A1 toxicity in these birds cannot be completely ruled out. Quail hens laid eggs with thinner shells when on a low Ca/P (0.65%/0.4%) diet, and this effect was exacerbated in the presence of 0"15% A1 (Wolff & Phillips, 1990). Sparling (1990) has reported dramatic increases in mortality of black ducklings exposed to 0.1% dietary A1 in combination with low dietary Ca (0-3%) and P (0-6%). Studies on the reproductive success of kingbirds (Tyrannus tyrannus) in acidified habitats in Canada (Glooschenko et al., 1986) and dippers (Cinclus cinclus) in Britain (Ormerod et al., 1988) have failed to demonstrate the gross abnormalities discussed by Nyholm (1981), even though these bird species prey on invertebrate taxa containing comparable concentrations of A1 ( ~ 0.1-0-2 %). Both the Canadian and British studies reported significant but relatively minor effects of wetland acidity/alkalinity on eggshell thickness, egg mass and/or fledgling growth rates; however, a greater proportion of the total variation in these parameters was accounted for by other factors such as changes between years, differences between nests on the same lake, differences among siblings, or unknown factors (Glooschenko et al., 1986; Ormerod et al., 1988). It is highly improbable that the severe reproductive impairments observed by Nyholm and Myhrberg (1977) can be attributed directly to A1 toxicity. Several researchers .(Carribre et al., 1986; Glooschenko et al., 1986; Blancher & McAuley, 1987; Longcore et al., 1987; Scheuhammer, 1987a, b; Ormerod et al., 1988; Drent & Woldendorp, 1989) have suggested that a scarcity of essential dietary Ca in acidified environments may adversely affect avian eggshell formation, clutch size,

Acidification, metals and wildlife

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and/or hatchling growth and survival, although direct evidence from field studies is scarce. The increased bone A1 accumulation observed by Nyholm (1981) in reproductively compromised female passerines from an acid sensitive habitat might be a secondary consequence of disruptions in normal Ca availability or Ca metabolism (cf. Quarles et al., 1985). A more detailed discussion of the possible influence of acidification on dietary Ca availability is presented in the following section of this paper. For certain wildlife species, the possibility that high A1 concentrations (> 0"1%) in food items may exacerbate the stress of a decreased dietary availability of high-Ca food items in some acidified environments cannot presently be ruled out. Certain aquatic macrophytes, especially submergent species, can accumulate very high concentrations (>4000/~g/g dry wt) of A1 in some acidified lakes and ponds (Sprenger & McIntosh, 1989). These A1 levels are high enough to be of toxicological concern for aquatic herbivores even assuming an adequate dietary intake of Ca and P. However, as is the case for Pb, empirical data relating to possible toxic effects of A1 in aquatic herbivores are lacking.

EFFECTS OF A C I D I F I C A T I O N ON THE AVAILABILITY OF ESSENTIAL CALCIUM Ca is a nutritionally essential element for both mammals and birds. An adequate supply of dietary Ca is particularly crucial during reproduction: in birds, for the proper formation of eggshell and skeletal growth of hatchlings, and in mammals, for skeletal development of the fetus in utero and in growing neonates. Here, the possibility that a reduced dietary availability of nutritionally essential minerals, particularly Ca, may be a significant factor affecting the reproductive success of wildlife breeding in acidified habitats is addressed. Although much of what is discussed will be of relevance to mammals, the discussion will focus on aquatic birds. Many, if not most, species of birds store readily mobilizable Ca in medullary bone. Medullary bone serves as a supply of Ca that can be rapidly transferred to the oviduct when eggshell is being formed, the gastrointestinal absorption of Ca from food alone being insufficient to meet the requirements for Ca (Taylor, 1970). The depletion of medullary Ca during eggshell formation triggers hormonal responses which result in an increased synthesis of intestinal Ca-binding protein, and a subsequent increase in the efficiency of intestinal Ca absorption. This dietary Ca is then used to replenish bone-Ca stores in preparation for the laying down of the next eggshell. It is during this egg laying period that it is critical for female birds to have access to high-Ca food sources. In experimental studies, egg

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A. M. Scheuhammer

production in hen pheasants (Phasianus colchicus) fell to about 1/10 of normal when dietary Ca was lowered to 0"37% from 2.34% and hens on restricted Ca diets were more likely to produce eggs with thin shells (Greely, 1962). Similarly, egg production in pheasants was about 50% lower when dietary Ca was 0"5% versus 3"2%, and hens on the restricted Ca diet lost Ca from bone and developed osteoperosis (Chambers et aL, 1966). Hatchling ducks require diets containing at least 0.5% Ca for optimal growth (Dean et al., 1967). Thus a reduction in the availability of Ca during reproduction can adversely affect clutch size, eggshell quality, the health of the adult female, and the growth and development o f young birds. Commercially available feeds, which reflect the optimal concentrations of various nutrients for maintenance of adult birds, egg laying and breeding, typically contain 1-3% Ca. These levels are an order of magnitude greater than those measured in a variety o f aquatic invertebrates from acidified environments (Hall & Likens, 1981; Sadler & Lynam, 1985; Ormerod et al., 1988). Also, most grains and other plant-type foods used by waterfowl contain very low concentrations of Ca ( < 0.1%) (Krapu & Swanson, 1975). Piscivorous birds can obtain Ca from the bones o f fish, assuming fish are available, but insectivorous and omnivorous birds are obliged to g < X

o

._ c .9

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0 4.5

5.0

i

i

i

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5.5

6.0

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LAKE pH

Fig. 2. Relative diversity of potentially important Ca-rich aquatic invertebrate taxa and their approximate pattern of disappearance from acid-sensitive regions in N. America as a function of pH. Gastropoda (Helisoma anceps, Physa gyrina, Gyraulus), Pelecypoda (Sphaeriidae, Unionidae), Amphipoda(Hyalella azteca, Crangonyx), Crustacea (Copopoda, Orconectes virilis, Carnbarus bartoni, Mysis relicta). Average pH values are plotted where available. All taxa are present at pH 7.0. Based on data from Collins et al., 1981; Eilers et al., 1984; McNicol and Blancher, 1989; and Schindler et al., 1985.

Acid~cation, metals and wildlife

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supplement their diets with high-Ca food sources, particularly during the reproductive period. Indeed, it has been observed that wild female birds actively select high-Ca grit and food items immediately before and during the egg-laying period (MacLean, 1974; Jones, 1976; Ankney & Scott, 1980) and altricial species also feed such items to their growing young (Betts, 1955; Mayoh & Zach, 1986; Blancher et al., 1987). High-Ca food sources of potential importance in aquatic environments are invertebrates which maintain a highly calcified shell (e.g., clams, snails, amphipods, crayfish). As illustrated in Fig. 2, and by Okland and Okland (1986) for similar taxa in Scandinavia, the occurence of these high-Ca aquatic organisms falls rapidly below pH 6"5 and by pH 5 almost all of these organisms are absent. Tree swallows (Tachycineta bicolor) breeding on highly acidic (pH < 5) wetlands in northern Ontario laid smaller eggs, and produced fewer and smaller fledglings than birds breeding near wetlands of higher pH (Blancher & McNicol, 1988). Mollusc shells (snails, clams) were observed in 30-50% of swallow nests near circumneutral (pH > 6) wetlands, but were almost absent in nests near acidic wetlands (Blancher et al., 1987). Researchers from the Netherlands have reported that the eggshell quality of great tits and other hole-nesting birds has fallen dramatically in some locations since 1983, and have related this phenomenon to acidification, declining soil-Ca z+, and increased soil-A13+ (Drent & Woldendorp, 1989). Observations such as these are consistent with the contention that the ability of insectivorous and omnivorous birds to find and use high-Ca food sources is compromised in acidified environments.

Calcium in aquatic invertebrates In order to test the hypothesis that the mineral content of aquatic invertebrates is influenced by wetland acidity or other related parameters of water chemistry, the following study was undertaken. Whirligig beetles (Dineutes nigrior) were sampled during the months of July and August from 16 lakes in northern Ontario for which detailed water chemistry data were available (McNicol et al., 1987a). Whirligigs were chosen for study because they were easily sampled, were abundant on lakes over a wide range of pH, and were occasionally preyed upon by ducklings. Four or five adult female beetles from each lake were freeze-dried, weighed, and digested in nitric acid. Ca and Mg concentrations of the digests were measured by standard flame atomic absorption spectrophotometry. Data were analyzed by the SAS package of statistical programs (SAS Institute, 1988). Initial one-way analyses of variance revealed that the Mg content of the beetles did not differ significantly among lakes (P>0-1), but that Ca concentrations did differ (P < 0"01). Regressions against various parameters of lake water chemistry

A. M. Scheuhammer

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(pH, Ca z+, Mg z+, K +, Na +, AI, Fe, Mn, Zn, [ C a + M g + N a + K ] , [ A I + Fe + Mn], soluble phosphate, dissolved organic carbon [DOC]) revealed that beetle-Ca was negatively correlated with water-A1, - [A1 + Fe + Mn] and - Zn, and positively correlated with pH, - Ca, - Mg, - [Ca + Mg + N a + K ] and DOC (P<0'05). Multiple stepwise regression was then performed for beetle-Ca and its correlates. Water-A1 concentration (ln [A1]) was the first variable added to the regression. This factor alone explained 57% of the variation in beetle-Ca concentration (r = - 0-756, P < 0-001). No other lake chemistry variable explained a significant amount of additional variation. In [A1] was significantly related to lake pH in a negative fashion (r = -0"92). The relationship between the concentration of Ca in whirligig beetles and water-A1 concentrations is depicted graphically in Fig. 3. There was a loss of approximately 30% of the normal beetle-Ca in lakes with the highest water-A1 concentrations. These results indicate that, not only are Ca-rich invertebrate taxa eliminated from wetlands undergoing acidification, but water chemistry parameters associated with acidification negatively influence the Ca concentrations of the remaining, acid-tolerant species. Others have reported a negative influence of ambient pH on Ca in crayfish (France, 1987), mayfly nymphs and blackfly larvae (Hall et al., 1988), stonefly nymphs and caddisfly larvae (Ormerod et al., 1988), and zooplankton (Yan et al., 1989). A convenient supply of Ca-rich foods essential for high reproductive success in non-piscivorous birds is generally unavailable in North American aquatic environments whose pH is less than --~5-3. A similar conclusion regarding the lack of high-Ca invertebrate taxa was reached by Okland and Okland (1986) for aquatic environments in Scandinavia.

-~

1.8

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m

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Fig. 3. Relationship between lakewater Al concentrations and Ca concentrations in whirligig beetles (Dineutes nigrior) collected from lakes in the Wanapitei study site near Sudbury, Ontario. For beetle-Ca concentrations, means for each lake (n = 4 or 5 per lake) are plotted, r = - 0 ' 7 5 6 , P < 0.001, n = 7 6 .

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M E T A L A C C U M U L A T I O N BY W A T E R F O W L IN A C I D I F I E D H A B I T A T S - - A CASE STUDY In order to examine the accumulation of metals by wildlife breeding in acidified environments, prefledgling black duck (Anas rubripes), ring-necked duck (Aythya collaris), c o m m o n goldeneye (Bucephala clangula), hooded merganser (Lophodytes cucullatus) and c o m m o n merganser (Mergus merganser) ducklings were sampled from numerous lakes in two study areas in northwestern Ontario (Fig. 4) for which detailed data on water chemistry, lake morphometry and trophic relationships have been established (Bendell & McNicol, 1987a,b; McNicol et al., 1987a,b). Some basic water chemistry parameters for lakes in the two study sites on which ducklings were collected are summarized in Table 2. Lakes in the Wanapitei study site are within 30-60 km of Sudbury, Ontario, whereas those in the Ranger Lake area are within 40 80 km of Sault St Marie, Ontario. The species of ducks chosen for study have significantly different feeding habits (McNicol et al., 1987a). Young black and ring-necked ducks feed mainly on surface insects in near-shore vegetation; goldeneye and hooded mergansers are pursuit divers that prey on large, mobile invertebrates, although goldeneye take more benthic material, and hooded mergansers probably take some fish and amphibians as well; c o m m o n mergansers are almost exclusively fish-eaters. In acidified habitats (pH < 5'5), the diets of the different insectivorous duck species become increasingly similar due to reliance on relatively few acid-tolerant odonate and nectonic insect taxa (McNicol et al., 1987b). 8

~

LAKE

"'-"-7

[ J^

•

~) !

83

I: I

82

81-

] ~'1 RAN~ER J LAKE

. I I~

I I

N

ONTARIO',

Fig. 4. Location of Ranger Lake and Lake Wanapitei study areas, Ontario, Canada.

A.M. Scheuhammer

354

TABLE 2

Water Chemistry Parameters (pH, Ca 2 + (mg/liter), DOC (mg/liter)) for Lakes in the Wanapitei and Ranger Lake Study Sites from which Ducklings were Collected. Values are Ranges of (n) Lakes

pH Ca DOC

Wanapitei

Ranger

4"3-7-6" (44) 1.2-16-0 (44) 1"5 27-5 (43)

5"2-7"4a (26) 1'9-6.1 (26) 4"3-18.4 (24)

" Significantly different, P < 0-05.

The purposes of this investigation were (1) to assess the tissue concentrations of metals in ducklings from a toxicological perspective; (2) to compare metal accumulation among different waterfowl species with different feeding habits; (3) to compare Ranger Lake and Wanapitei with respect to metal accumulation in ducklings; and (4) to determine whether any of the variability in duckling-metal accumulation could be accounted for by water chemistry parameters. The main water chemistry variables that were thought to be potentially important correlates with tissue metal accumulation in ducklings, and for which data were available, are pH, [Ca 2 +] and DOC. Data for some other potentially significant parameters, such as lakewater concentrations of Cd 2 +, Pb 2 + and Hg 2 +, had not been measured and were thus not available for consideration. Collections were made in 1984-86 during daylight hours from late May to early August each year. Birds were shot with a 12-gauge shotgun and carcasses were frozen until later dissection and removal of kidneys and livers at the National Wildlife Research Centre, Hull, Quebec. Tissues were weighed and sent to Ontario Research Foundation, Mississauga, Ontario for analysis of metals by atomic absorption spectrophotometry. In the case of Hg, only total tissue Hg was measured. The SAS package of statistical programs (SAS Institute, 1988) was used to analyze the resulting data. Only the tissue (liver or kidney) with the highest concentration of each metal of interest was used in the statistical analyses.

Zinc (liver) Hepatic Zn concentrations ranged between 25 and 55 #g/g (wet wt) and were highly species specific (P<0-01), and not influenced by geographical

Acidification, metals and wildlife

355

02)

,~

g,

20 ¸ 10 o

~'~.,~1.6f

(7

1'2f o.a

•-

0.0

,o ""

7

'7,-,

BlockRingneck GoldeneyeMerg. Uerg. Hooded

Common

Fig. 5. Hepatic Zn (top) and Hg (bottom) in prefledging ducklings of various species from lakes in the Ranger Lake or Lake Wanapitei study areas. Values in brackets are sample sizes.

location (study site) (Fig. 5, top). These data are presented in order to illustrate the behavior of a homeostatically controlled, nutritionally essential trace metal.

Mercury (liver) Hepatic Hg concentrations were significantly influenced by both species and location (P < 0"01; Fig. 5, bottom). There was a trend towards increasing Hg concentrations as the predatory tendency of the species increased. The piscivorous c o m m o n merganser had the highest Hg levels, which is an expected result based on the known behavior of Hg in foodchains. For each species, there was a tendency for Hg concentrations to be lower at Wanapitei than at Ranger Lake. Low levels of rig in tissues ofpiscivorous mammals from acid-sensitive environments around Sudbury, Ontario relative to other acid-sensitive area in Ontario have also been reported (Wren et al., 1986). Acidified environments around Sudbury are anomalous

356

A. M. Scheuhammer

with respect to the accumulation of Hg by fish and other biota (Wren & Stokes, 1988) which, in other locations, generally demonstrate significantly elevated Hg levels with declining pH (Spry & Wiener, 1991; Wren & Stephenson, 1991). Some local ameliorating factor, possibly Se which is known to be emitted in high concentrations from Sudbury smelters (Nriagu & Wong, 1983) and to retard the rate of bioaccumulation of Hg by fish, crayfish and haptobenthos (Rudd et al., 1980), is acting to produce generally low tissue-Hg accumulation in biota in the Sudbury region. In Wanapitei, duckling-Hg levels were not related to pH, DOC or Ca except for common mergansers for which a positive relationship was noted between pH and hepatic Hg (r = 0"627, P = 0"022; Fig. 6). This positive correlation is the reverse of the expected trend for piscivores, considering that there is generally a negative relationship between pH and Hg in fish (Spry & Wiener, 1991), thus the anomalous nature of acidified habitats in the Sudbury region with respect to the behavior of Hg is again illustrated. At Ranger Lake, liver-Hg concentrations in ducklings were found not to be significantly related to pH, Ca or DOC; however, the sample size for the piscivorous common merganser was low from this study site (7 samples from 5 lakes). Hepatic Hg concentrations ranged from a low of around 0"1 #g/g in black ducks from Wanapitei to highs of 1.6-1.9/~g/g in some common merganser ducklings from Ranger Lake and Wanapitei. These higher concentrations are not trivial. They are comparable to hepatic Hg concentrations of loon chicks from the Hg-contaminated English-Wabigoon River system where loon reproduction was impaired, and higher than those of loon chicks from adjacent, non-contaminated areas where loon reproductive success was normal (Canadian Wildl. Serv., unpublished). Unfortunately, the sample size for common mergansers from Ranger Lake was low (7 samples from only 5 different lakes) and no sample was collected from a wetland of pH < 6.4. Thus, it was not possible within this study area to test the hypothesis that Hg in tissues of these piscivores increases with decreasing pH of their environment. Data for other Ranger Lake species which, statistically, could be pooled (goldeneye, ring-necked duck, hooded .+.~ • 2.0 •

~ Fig. 6. Relationship between lake pH and hepatic Hg concentration in common merganser ducklings collected at the Lake Wanapitei study lakes; r = 0'627, P < 0"05.

1.5

I . 0

0.5 > o.o •-i 5.0

• sis

8'.o815 L a k e pH

71o

7.s

Acidification, metals and wildliJe

357

merganser) did reveal the expected negative relationship between liver-Hg and pH (r = - 0-263) but the trend was not statistically significant (P = 0.09). These species are not fish-eaters and thus are generally exposed to lower dietary MeHg than c o m m o n mergansers.

Lead (kidney) Due to the presence of some very high Pb values (> 10#g/g wet wt), most likely due to contamination by Pb fragments from shotgun pellets (Frank, 1986), the technique of Grubbs (1969) was used to identify and remove outliers. This procedure was repeated until no outliers remained at P ==0"025. In all, 31 of 142 observations were judged to be outliers. Kidney-Pb was not significantly affected by location (Ranger versus Wanapitei) nor by duckling age. There were significant differences among species, perhaps reflective of different dietary habits. The piscivorous c o m m o n merganser had significantly lower kidney-Pb concentrations than the other species. For those species among which mean renal Pb concentrations were not statistically different (after removal of outliers), and which were then pooled for linear regression analyses, no significant relationships were detected between kidney-Pb and pH, Ca, or DOC at Ranger Lake. At Wanapitei, there was a significant negative relationship between kidney-Pb concentrations and pH (r = - 0 - 3 2 4 , P = 0"048) for a pooled sample of goldeneye, ring-necked and black ducks (Fig. 7, top). Notwithstanding this trend, Pb concentrations were low, ranging from 0"05-0"45/~g/g wet wt. These concentrations are not toxicologically significant.

Cadmium (kidney) Kidney-Cd was influenced by study location, but not by species or duckling age. Kidney-Cd concentrations were significantly higher at Ranger than at

0'5t i . o .... "

Pb

t.

i 1

o 0.01

~_--

-

,

Cd

0.~[, ~ ~ o.2F v • v A~I--'~ ~7 o.1 [ i ~ . . - - - - " ' ~ ' ~ ,~ 0.o'J--A _ [ ~ -~- . •8 ~ . . . . 4.0 s.o 6.0 7.0 Lake pH

I

Fig. 7. Relationship between lake pH and Pb (top) or Cd (bottom) concentrations in kidneys of ducklings collected on lakes in the lake Wanapitei study area. For Pb, r = -0"324, P < 0'05; for Cd, r = 0"366, P < 0"05. Species plotted are goldeneye (Q), ring-necked duck (©), black duck (&), and hooded merganser (V).

358

A. M. Scheuhammer

Wanapitei for black and goldeneye ducklings. At Wanapitei, kidney-Cd concentrations were significantly influenced by pH in a positive fashion for a pooled sample of goldeneye, hooded mergansers and black ducks (r = 0.366, P = 0-019) (Fig. 7, bottom). No other significant relationships were observed. At Ranger, no significant relationships were observed between kidney-Cd and any o f the water chemistry variables. The highest renal Cd concentrations were observed in goldeneye ducklings from Ranger Lake, but even these (0.2q>6/~g/g) were not toxicologically significant. Aluminum (kidney) Kidney-A1 was not significantly affected by study location or duckling age, but was influenced by species. For both sampling sites, kidney-A1 was not related to any of the water chemistry variables tested. Individual-kidney-A1 concentrations ranged from undetectable ( < 0.01 #g/g) to about 5/tg/g. The highest concentrations were in goldeneye from Wanapitei (1-6_+ 1.3 ~g/g; n = 10). Metals in invertebrate prey Aquatic macroinvertebrate taxa were sampled from lakes over a wide pH range in the Wanapitei study site by methods described previously (Bendell & McNicol, 1987a,c; McNicol et al., 1987a). Samples were kept frozen prior to identification. After separation into taxonomic group, individual adult dragonflies (Anisoptera), backswimmers (Notonecta), whirligig beetles (Dineutes) and waterstriders (Metrobates) from each lake were pooled TABLE 3 Concentrations of Cd and Pb (/~g/g wet wt) in Invertebrates from Wanapitei, Ontario. Values are Means (range) from n Lakes, except t, where all Individuals from all Lakes were Pooled Taxa

Anisoptera Dineutes Notonecta Metrobates

Lake pH

n

5.44 7 (4'63-6"33) 5.95 14 (4.35-7.21) 5'86 10 (4.35-7.21) 6.13 8t (5.10-7.10)

nd = Not detectable.

Cd

Pb

< 0.02 0-34 (nd~3-06) (0-06q3.53) 0"06 0-26 (0.013~3.15) (0.17~)'48) 1-5 0.75 (0.2-4-2) (0.05-1'8) 2-8 0.57

Acid!~cation, metals and wildl([i,

359

according to taxa, each pool weighing 0.5-2g twet wt). Cd and Pb concentrations were determined in pooled samples by the Ontario Research Foundation (Mississauga, Ontario) using graphite furnace atomic absorption spectrophotometry. Results are presented in Table 3. Cd and Pb concentrations were highest in waterstriders and backswimmers, and low in whirligigs and dragonflies. Hg concentrations, which were also measured in dragonflies (data not shown) were very low ( < 0"01/~g/g wet wt). From a toxicological perspective, the concentrations of Cd, Pb and Hg measured in these invertebrate taxa are not sufficiently high to be of concern. Adult waterfowl and ducklings should be able to consume these food items without any significant risk to health from metal toxicity.

Case study conclusions The present survey of tissue metal concentrations accumulated by ducklings of various species during their first few months of life indicates that dietary exposure of ducklings to toxicologically relevant levels of Cd, Pb or AI is unlikely to occur in acidified environments. Furthermore, there was a general absence of strong relationships between tissue metal concentrations and pH or other water chemistry parameters. This may in part be due to the fact that ducklings can move among wetlands (D. McNicol, pers. comm.) thus confounding the interpretation of the data. It may be unreasonable to expect extremely close relationships between lake chemistry and tissuemetal concentrations in mobile wildlife species. Better results may be obtained by looking for regional differences (i.e., acidified versus buffered habitats) in metal accumulation. Studies which have done this have generally found increased metal accumulation in wildlife from some acid-sensitive environments (Glooschenko et al., 1986; Wren et al., 1986; Crate et al., 1987). Nevertheless, the accumulation of metals by ducklings in the present study was generally not high from a toxicological perspective. Tissue-Cd concentrations were higher in ducklings collected in higher pH lakes, whereas the opposite trend was observed for tissue-Pb (Fig. 7). The difference in behavior of Cd and Pb follows the predicted biological response of these metals to acidification of aquatic environments (Campbell & Stokes, 1985), and the observed pH-related accumulation of Cd and Pb in some species of aquatic insects (Krantzberg & Stokes, 1988). Although a statistically significant negative relationship was noted between kidney-Pb and lakewater pH, the relationship was weak, and pH accounted for < 10% of the variability in tissue-Pb concentrations. Eriksson et al. (1989) failed to find significantly different concentrations of Hg, Cd, Pb, AI or other metals in tissues of prefledgling goldeneye from acidified lakes compared with circumneutral lakes in Sweden. Also, metal concentrations in invertebrate

360

A . M . Scheuhammer

prey (Table 3) were not sufficiently high to be toxic. The relative risk of metal-induced effects on health or reproductive success of waterfowl inhabiting acidified environments appears to be low. The exception may be Hg exposure in piscivorous waterfowl and, by analogy, other piscivorous wildlife. In the present study, a significant positive correlation was observed between lake pH and liver-Hg in common merganser ducklings collected in the Wanapitei study site. This area, however, is within 30-60 km of the metal smelters of Sudbury, Ontario (Fig. 4), a region in which biota are known to accumulate anomalously low concentrations of Hg, compared with other acid-sensitive environments (Wren & Stokes, 1988; Wren et al., 1986). Hg relationships determined for biota in these habitats should not be considered as typical for acidified environments. Hg concentrations may be influenced more by distance and direction from Sudbury, than by water chemistry parameters like pH. In our other study site, the Ranger Lake area, the sample size for common mergansers was too low for any relationships to be established. Nonetheless, liver-Hg in individual ducklings of this species were high enough to be of toxicological concern. A similar result was reported for individual goldeneye ducklings from acidic lakes in Sweden (Eriksson et aL, 1989). The results of Barr (1986) indicate that fish-eating birds can react to increased Hg in prey by avoiding potential territories, abandoning nests, or not laying eggs. The relationship between acidification, Hg in fish, and possible consequences for fish-eating wildlife requires further investigation.

S U M M A R Y A N D R E C O M M E N D A T I O N S FOR F U T U R E RESEARCH

Mercury Due to the negative relationship between Hg in fish and wetland pH, piscivorous wildlife living in habitats dominated by acidic wetlands risk higher dietary exposure to MeHg than those living in well-buffered, circumneutral environments. Based on the results of controlled feeding studies and a field study, those wetlands in which fish of an appropriate size and species contain Hg concentrations averaging > 0"3/~g/g (wet wt) should be considered as potentially detrimental to the reproductive success of loons and other piscivorous wildlife. Comprehensive surveys of Hg in fish of an appropriate size range and species should be undertaken in order to establish the extent of these 'high-Hg' habitats. Further studies to specifically investigate the feeding and reproductive behavior of, and Hg accumulation by, piscivorous wildlife in such high-risk environments should

Acidification, metals and wildli[e

361

then be undertaken. There is no evidence to suggest that non-piscivorous wildlife are at risk from dietary MeHg exposure in acidified wetlands.

Lead, cadmium, aluminum and calcium There is no convincing evidence to suggest that Pb, Cd or A1 concentrations in prey of most piscivorous or insectivorous wildlife dwelling in acidified environments are sufficiently high to be of general concern with respect to possible health or reproductive impairment. Exposure of wildlife to these potentially toxic, non-essential metals may be of concern in instances where the availability of dietary Ca or P is low and the availability of the toxic metal(s) is uncommonly high. There is, however, some indication that certain macrophytes growing in low pH wetlands can accumulate Pb and/or AI to levels known to be toxic to consumers even in the absence of a restricted availability of dietary Ca and P. Studies should be undertaken to investigate the response of those plant species that are important in the diets of herbivorous wildlife to increasing environmental acidity and results should be assessed in conjunction with data regarding metal accumulation and effects in the appropriate wildlife consumers. Such studies should also consider the response of Ca and P in the same plants and should comment on the possibility for decreased availability of essential minerals to herbivores living in acidified habitats. More generally, the bioavailability of Ca and other essential elements such as P, Mg and Se in acidified environments should be ascertained by surveying a wide variety of fish, invertebrates and plants used as food by wild birds and mammals, and relating results to changes in pH or other water chemistry variables associated with acidification. Present evidence indicates a scarcity of high-Ca aquatic invertebrates in acidified wetlands, and a general decline in plant- and invertebrate-Ca concentrations as ambient acidity and AI 3 ~ concentrations increase. In terrestrial ecosystems, acid precipitation may be responsible for accelerated depletion of Ca from forest soils (Federer et al., 1989). In turn, depletion of essential Ca and increased environmental [AI 3 +] may be contributing to forest decline in some localities (Cronan et al., 1989; Shortle & Smith, 1988) and creating a scarcity of dietary Ca for forest wildlife (Drent & Woldendorp, 1989). Such trends, if unchecked, indicate potentially far-reaching consequences for the reproductive success and health of numerous wildlife species. ACKNOWLEDGEMENTS The author thanks Don McNicol, Barry Bendell and others at CWSOntario Region for the collection, storage and identification of invertebrate

362

A.M. Scheuhammer

and duckling samples and for providing the lake-water chemistry data; Della Bond for providing technical assistance in the laboratory; M a r y G a m b e r g and Laurie Wilson for doing the SAS statistical analyses; Peter Blancher for numerous helpful discussions, and for reviewing an early draft of the manuscript; and the other contributors to this special issue of Environmental Pollution for their comments and suggestions. This work was performed as part o f the C W S - L R T A P (Long Range Transport of Airborne Pollutants) Program.

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Benson, W. W., Brock, D. W., Gabica, J. & Loomis, M. (1976). Swan mortality due to certain heavy metals in the Mission Lake area, Idaho. Bull. Environ. Contain. Toxicol., 15, 171-4. Berg, W., Johnels, A., Sjostrand, B. & Westermark, T. (1966). Mercury content in feathers of Swedish birds from the past 100 years. Oikos, 17, 71-83. Berlin, M., Carlson, J. & Norseth, T. (1975). Dose-dependence of methyl mercury metabolism. Arch. Environ. Health, 30, 307-13. Bernard, B. P. & Becker, C. E. (1988). Environmental lead exposure and the kidney. Clin. Toxicol., 26, 1-34. Betts, M. M. (1955). The food of titmice in oak woodland. J. Animal Ecol., 24, 282 323. Beyer, W. N., Spann, J. W., Sileo, L. & Franson, J. C. (1988). Lead poisoning in six captive avian species. Arch. Environ. Contam. Toxicol., 17, 121-30. Bhatnagar, M. K., Vrablic, O. E. & Yamashiro, S. (1982). Ultrastructural alterations of the liver of Peking ducks fed methyl mercury-containing diets. J. Toxicol. Environ. Health, |0, 981 1003. Bj6rklund, I., Borg, H. & Johansson, K. (1984). Mercury in Swedish lakes Its regional distribution and causes. Ambio, 9, 118 21. Blancher, P. J. & McAuley, D. G. (1987). Influence of wetland acidity on avian breeding success. Trans. N.A. Wildl. Nat. Res. Cot!ll, 52, 628 35. Blancher, P. J. & McNicol, D. K. (1988). Breeding biology of tree swallows in relation to wetland acidity. Can. J. Zool., 66, 842 9. Blancher, P. J., Furlonger, C. L. & McNicol, D. K. (1987). Diet of nestling tree swallows (Tachycineta bicolor) near Sudbury, Ontario, Summer 1986. Can. Wildl. Serv. Tech. Report Ser., 31, 14 pp. Blomqvist, S., Frank, A. & Petersson, L. R. (1987). Metals in liver and kidney tissues of autumn-migrating dunlin Calidris alpina and curlew sandpiper Calidris Jerruginea staging at the Baltic Sea. Mar. Ecol. Prog. Ser., 35, 1- 13. Bondy, S. C. (1988). The neurotoxicity of organic and inorganic lead. in: Metal Neurotoxieity, ed. S. C. Bondy & K. N. Prasad. CRC Press, Boca Raton, FL, pp. 1-17. Braune, B. M. (1987). Comparison of total mercury levels in relation to diet and molt for nine species of marine birds. Arch. Environ. Contam. Toxicol., 16, 217 24. Braune, B. M. & Gaskin, D. E. (1987). Mercury levels in Bonaparte's Gulls (Larus philadelphia) during autumn molt in the Quoddy region, New Brunswick, Canada. Arch. Envir. Contam. Toxicol., 16, 539-49. Bull, K. R., Murton, R. K., Osborn. D., Ward, P. & Chen, G. (1977). High cadmium levels in Atlantic sea birds and seaskaters. Nature, 269, 507-9. Bunn, C. R. & Matrone, G. (1966). In vivo interactions of cadmium, copper, zinc, and iron in the mouse and rat. J. Nutr., 90, 395 9. Burbacher, T. M., Monnett, C., Grant, K. S. & Mottet, N. K. (1984). Methylmercury exposure and reproductive dysfunction in the nonhuman primate. To.vicol. Appl. Pharmacol., 75, 18-24. Cain, B. W., Sileo, L., Franson, J. C. & Moore, J. (1983). Effects of dietary cadmium on mallard ducklings. Environ. Res., 32, 286-97. Campbell, P. G. C. & Stokes, P. M. (1985). Acidification and toxicity of metals to aquatic biota. Can. J. Fish. Aquat. Sci., 42, 2034-49. Cappon, C. J. & Smith, J. C. (1981). Mercury and selenium content and chemical form in fish muscle. Arch. Environ. Contam. Toxicol., 10, 305-9.

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Carlson, B. L. & Nielson, S. W. (1985). Influence of dietary calcium on lead poisoning in mallard ducks (Anas platyrhynchos). Amer. J. Vet. Res., 46, 276-82. Carri+re, D., Fischer, K. L.,Peakail, D. B. & Anghern, P. (1986). Effects of dietary aluminum sulphate on reproductive success and growth of ring doves (Streptopelia risoria). Can. J. Zool., 64, 1500-5. Chambers, G. D., Sadler, K. C. & Breitenbach, R. P. (1966). Effects of dietary calcium levels on egg production and bone structure of pheasants. J. Wildl. Manage., 30, 65-73. Chang, L. W. & Annau, Z. (1984). Developmental neuropathology and behavioral teratology of methyl mercury. In: Neurobehavioural Teratology, ed. J. Yanai. Elsevier, Amsterdam, pp. 405 32. Collins, N. C., Zimmerman, A. P. & Knoechei, R. (1981). Comparisons of benthic infauna and epifauna biomass in acidified and non-acidified Ontario lakes. In: Effects of Acidic Precipitation on Benthos, ed. R. Singer. N. Amer. Benthological Soc., Springfield, Illinois, pp. 35-48. Cook, R. S. & Trainer, D. O. (1966). Experimental lead poisoning of Canada geese. J. Wildl. Manage., 30, 1-8. Cory-Slechta, D. A., Weiss, B. & Cox, C. (1985). Performance and exposure indices of rats exposed to low concentrations of lead. ToxicoL Appl. Pharmacol., 78, 291-9. Cox, G. J., Dodds, M. L., Wigman, H. B. & Murphy, F. J. (1931). The effects of high doses of aluminum and iron on phosphorus metabolism. J. Biol. Chem., 92, 11-12. Cr&e, M., Potvin, F., Walsh, P., Benedetti, J.-L., Lefebvre, M. A., Weber, J.-P., Paillard, G. & Gagnon, J. (1987). Pattern of cadmium contamination in the liver and kidneys of moose and white-tailed deer in Quebec. Sci. Total Environ., 66, 45-53. Cr&e, M., Nault, R., Walsh, P., Benedetti, J.-L., Lefebvre, M. A., Weber, J.-P. & Gagnon, J. (1989). Variation in cadmium content of caribou tissues from northern Quebec. Sci. Total Environ., 80, 103-12. Cronan, C. S., April, R., Bartlett, R. J., Bloom, P. R., Driscoll, C. T., Gherini, S. A., Henderson, G. S., Joslin, J. D., Kelly, J. M., Newton, R. M., Parnell, R. A., Petterson, H. H., Raynal, D. J., Schaedle, M., Schofield, C. L., Sucoff, E. I., Tepper, H. B. & Thornton, F. C. (1989). Aluminum toxicity in forests exposed to acidic deposition: The ALBIOS results. Water Air Soil Pollut., 48, 181-92. Crowder, A. (1991). Acidification, metals and macrophytes. Environ. Pollut., 71(2-4). Custer, T. W., Franson, J. C. & Pattee, O. H. (1984). Tissue lead distribution and hematologic effects in American kestrels (Falco sparverius) fed biologically incorporated lead. J. Wildl. Dis., 20, 39-43. Davis, J. M. & Svendsgaard, D. J. (1987). Lead and child development. Nature, 329, 297 300. Dean, W. F., Scott, M. L., Young, R. J. & Ash, W. J. (1967). Calcium requirements of ducklings. Poult. Sci., 46, 1496-9. Demayo, A., Taylor, M. C., Taylor, K. W. & Hodson, P. V. (1982). Toxic effects of lead and lead compounds on human health, aquatic life, wildlife, plants, and livestock. CRC Crit. Rev. Environ. Control, 12, 257-305. Deobaid, H. J. & Elvehjem, C. A. (1935). The effect of feeding high amounts of soluble iron and aluminum salts. Amer. J. PhysioL, 111, 118-23.

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