by Dr. Siu-Ling Han and Dr. David Stone

Presented by Dr. David Stone

This paper is divided into four components. The first section provides a broad overview of some of the types of effects that may be associated with persistent organic pollutants (POPs). It illustrates the wide range of outcomes that may potentially occur, even as a result of very low levels of exposure. We then briefly draw attention to the role played by persistence, bioaccumulation and biomagnification in determining environmental levels in biota of POPs. Next, we illustrate that despite current low levels in air and water in areas removed from direct local sources, existing environmental burdens of POPs may be approaching critical threshold levels in some wildlife and human populations. The examples we use are from Arctic ecosystems. Finally, we provide some preliminary comment on possible implications to other parts of the world.

1) Effects Overview

The weight of evidence from laboratory, wildlife, and human studies is such that it is generally accepted that POPs are a group of substances that have the potential to cause harm to wildlife and humans (UNECE 1994). However, the available literature on species-specific chronic effects of environmental exposures to individual POPs is far from comprehensive. It is, however, known that the effects caused by a given POP (or congener or metabolite) may vary according to the animal species, age and gender, and the level, extent and duration of exposure (Swain et al. 1992). Timing of exposure relative to the organism's life cycle is also a critical determinant of outcome (Kurzel and Cetrulo 1981; Jacobson et al. 1989; Jacobson et al. 1990; Harada 1976; as cited in Swain et al. 1992).

There may be a considerable delay between POP exposure and onset of effects (Bertazzi 1991). In addition, POP effects are often not manifested in the environmentally exposed adult organism, rather, abnormalities may occur in the second or third generation offspring (Murray et al. 1979; USEPA 1997), and even then perhaps not becoming evident until the offspring reach physical or sexual maturity (Mably et al. 1992; and, Wannemacher et al. 1992; Walker et al. 1994; as cited in Bidleman et al. 1997). While gross abnormalities associated with POP exposure (such as crossed bills, club feet, tumours and lesions etc., ) have been well-reported (Government of Canada 1991), a number of the abnormalities reported to be associated with low level environmental exposure to POPs are subtle and not readily apparent at the individual level, notwithstanding their serious implications at the population level. Examples include immune system deficiencies (Vos and Luster 1989; DeSwart et al. 1995, Ross et al. 1995, Ross et al. 1996; as cited in Bidleman et al. 1997) and impairment of reproductive function and development (Brouwer et al. 1995; Baron et al. 1995, Bosveld and Van de Berg 1994; Peterson et al. 1993 as cited in Bidleman et al. 1997).

A key reason for the heightened concern about the potential effects of POPs is the growing evidence that certain POPs can act as endocrine disruptors, thereby exerting effects by interfering with a variety of biological mechanisms, and resulting in a wide range of different outcomes (USEPA 1997, UNECE 1995, Colborn et al. 1993). The endocrine system plays a critical role in organism development and biochemical function. Disturbances can arise from small or large single exposures as well as from small, cumulative exposures, and often can result in transgenerational damage of a long-lasting nature (Swain et al. 1992).

A recent review by the United States Environmental Protection Agency (USEPA 1997) defines an endocrine disruptor as 'an exogenous agent that interferes with the synthesis, secretion, transport, binding, action, or elimination of natural hormones in the body that are responsible for the maintenance of homeostasis, reproduction, development, and/or behaviour.' The review goes on to state that 'there are potentially several target organ sites at which a given environmental agent could disrupt endocrine function...because of the complexity of the cellular processes involved in hormonal communication, any of these loci could be involve mechanistically in a toxicant's endocrine-related effect.' These statements provide some insight into the breadth of effects that could potentially be induced by exposure to POPs. Among the POP pesticides to which endocrine disrupting effects in wildlife have been attributed are dieldrin/aldrin, DDT, endosulfan, methoxychlor, and toxaphene (USEPA 1997).

Although there have been some results indicative of possible synergistic effects of combinations of POPs (eg. Soto et al. 1994), the evidence remains inconclusive. Reports of synergistic effects in a yeast cell system containing the human estrogen receptor with mixtures of endosulfan, dieldrin, toxaphene, and chlordane by Arnold et al. (1996) have failed to be replicated by others (Ramamoorthy et al. 1997).

The following provides some specific examples of the different effects that have been documented in different species at different levels of exposure to different POPs. It demonstrates the broad range of possible effects, and the low concentrations that may cause such effects. It also illustrates the difficulty in establishing minimum load levels due to the complex relationships between POP exposures and effects.

Immunosuppression

Vos and Luster (1989) suggest that immunosuppressive effects may be one of the most sensitive and relevant environmental threats posed by POPs. Many POPs can cause multiple effects on the immune system, in some cases disrupting the body's ability to produce antibodies or produce T-cells which fight against tumors and viruses or acting directly on the thymus, causing atrophy (Bidleman et al. 1997). Specifically, POP exposures have resulted in reduced antibody production following exposure to a foreign antigen, changes in T-cell populations, suppression of the anti-sheep red blood cell (SRBC) plaque-forming responses due to suppression of T-cell responses, decreased delayed-type hypersensitivity, decreased resistance to virus infections and decreased natural killer cell activity (Tryphonas 1994, Wong et al. 1992, as cited in Bidleman et al. 1997). Experiments with harbour seals fed on Baltic Sea fish show significant differences in immune function measures (e.g. white blood cell counts, lymphocyte function, natural killer cell activity) with seals fed fish from relatively unpolluted Atlantic waters (de Swart et al. 1994). Swain et al. (1992) review the evidence of immune system impairment by POP compounds such as TCDDs, PCBs, HCB, endosulfan, toxaphene, and DDT. Martineau et al. (1990), Muir et al. (1990), Borrel and Aguilar (1991), and Béland et al. (1992) provide examples of correlations between elevated body burdens of certain POPs and disease outcomes in animals from marine mammals die-offs and strandings.

Reproduction and Development

POP exposure can affect reproduction through a number of avenues e.g. by diminishing survival of offspring, or by disrupting reproductive function and reproductive cycles of adult animals. The review by Swain et al. (1992) describes experimental studies that have demonstrated the potential reproductive disruption that may result from exposure to POPs such as TCDD, DDT, PCBs, PBB, HCB, mirex, methoxychlor, dieldrin and HCH. Certain POPs can be embryo- and fetotoxic, or result in higher mortality rates of young born to exposed adults. Abnormalities in estrus cycle and sex hormone levels, reduced sperm production, reduced litter size, and in some cases total reproductive failure have also been reported. There are many well-known examples of bird populations that have been affected by POP exposure, with decreased or retarded egg production, increased embryo mortality, eggshell thinning, embryonic deformities, growth retardation and reduced egg hatchability being among the effects reported. As described in Bidleman et al. 1997, reviews by Brouwer et al. (1995), Barron et al. (1995), Bosveld and Van de Berg (1994), and Peterson et al. (1993) demonstrate the wide range of reproductive and developmental effects that have been associated with POPs. Field observations of population declines that are hypothesized to be linked with POP exposure are consistent with controlled dietary experiments on similar species, with failed implantation, premature births, decreased hatching success, and increased mortality of offspring all having been demonstrated (Aulerich and Ringer, 1977, Heaton et al. 1991, Mac and Edsall 1991, Kihlstrom et al. 1992).

Cancer and Tumour Formation

The importance of POPs relative to cancer effects appears to be greater with respect to promotion, rather than induction, of cancer. Bidleman et al. (1997) state that 'most POPs are regarded as cancer promoters', and support this with examples from the literature on TCDD, other PCDD/PCDF congeners, coplanar PCBs, PCB mixtures. However, notwithstanding the experimental evidence of genotoxic or tumour promoting behaviour of POPs, there have not been, except in the case of PAHs, direct links made between exposure to specific POPs and increased cancer rates in wildlife. PAHs and other POP contaminants have been implicated in the high incidence of tumours and lesions of the skin and organs in wildlife from POP contaminated regions (Martineau et al. 1987, 1988, Muir et al. 1990, Government of Canada 1991, Gilman 1991, Béland et al. 1992, Swain et al. 1992).

Biochemical Markers of Effect

Exposure to xenobiotics can provoke a molecular level response in organisms that can be used as a measure of whether or not a contaminant is acting biologically on the organism, although a quantitative relationship to toxic effects is not well established (Muir et al. 1996). Cytochrome P-450 enzyme induction in liver is the main biomarker system studied in Arctic biota exposed to low environmental levels of contaminants. Induction of cytochrome P-450 has been observed in Arctic fish, beluga, ringed seals and polar bears. Strong statistical correlations of biomarker induction with measured contaminant residue levels have been found (Muir et al. 1996), indicating that the level of induction appears to correlate with degree of contaminant exposure.

Exposure to certain POPs is also known to affect enzymes involved in the biosynthesis of heme, leading to porphyria in cases of long-term exposure to high concentrations. The biochemical changes associated with porphyria occur well before symptoms appear and thus can be used as sensitive biomarkers of POP exposure (AMAP 1997).

Vitamin A metabolism can also be affected by POP exposure, and this can result in increased susceptibility to microbial infections and cancer, reproductive disorders, skin lesions and disruptions in growth and development (Brouwer et al. 1989, AMAP 1997). Experiments with seals (Brouwer et al. 1989, deSwart et al. 1994) and mink (Håkansson et al. 1992) have shown that vitamin A (retinol) levels can be affected by exposure to PCBs.

Certain POPs have also been shown to exert effects on the adrenal gland (Wassserman et al. 1973, Copeland and Cranmer 1974, Lehman et al. 1974, Sanders et al. 1977, Hansen et al. 1979; as cited in Kuiken et al. 1993) and on thyroid hormone levels (Brouwer et al. 1989). A disease complex known as adrenocortical hyperplasia that has been described in Baltic Seals as possibly linked to chlorinated hydrocarbon contamination (Bergman and Olsson 1986, cited in Olsson et al. 1994) although Kuiken et al. (1993) did not find an association between incidence of the syndrome and levels of chlorinated hydrocarbons in porpoise carcasses examined in Great Britain.

2) The significance of persistence, bioaccumulation, and biomagnification in determining environmental levels of POPs

Due to the persistence of POPs, and their tendency to bioaccumulate and biomagnify, almost undetectable concentrations in the abiotic environment have the potential under certain conditions to result in significant exposure levels for organisms at higher trophic levels, e.g. humans and marine mammals (with biomagnification factors greater than 10⁷ in some cases) (Muir et al. 1996).

A number of attempts have been made to establish, evaluate and utilize persistence and bioaccumulation or bioconcentration (not biomagnification) criteria to help guide priority setting for management regimes or policies (Environment Canada, 1995; United Nations Economic Commission for Europe. 1996). Their use in this way has recently been reviewed by Chapman et al (1996) and Chapman (1997). The latter authors point out that no satisfactory methodologies exist for predicting the potential of a POP to biomagnify (successively increase in concentration at higher and higher trophic levels within a given food chain). Consequently, the approach frequently taken is instead to utilize measures of bioaccumulation (the amount taken up by an organism from water and diet). The most commonly used indicator (Biological Concentration Factor - BCF) in fact only estimates uptake from water.

Unfortunately, the bioaccumulation potential of a substance, (or its BCF) is a poor indicator of its biomagnification potential and it must be remembered that it is the latter which is largely responsible for delivering a dose at a potential 'effects level' to upper trophic level organisms. A number of factors are responsible, including the influence of trophic structure and differential metabolic characteristics of different species. A good example of the former has been provided by studies in adjacent Yukon (Canada) lakes where dramatically different levels of POPs in top predator fish species between lakes were shown to be due to differences in the trophic structure of each Lake (Kidd et al. 1995a,b, Schindler et al.1995). Differences in the ability of species to metabolise different POPs or even the different congeners of the same POP frequently also confound attempts to predict biomagnification (Muir et al 1988). For example, Norstrom et al.(1988) have shown that Polar Bears have an ability to metabolize normally recalcitrant PCB congeners and 4,4' DDE. Similarly, Toxaphene concentrations in ringed seals are much lower than in whales or fish from the same vicinity (Bidleman et al. 1993).

3) Environmental Concentrations of POPs and Thresholds for Biological Effects

Marine wildlife feeding at higher levels in the food chain tend to have the highest body burdens of POPs (Muir et al. 1996). As illustrated in Figure 1, Arctic prey species (freshwater and marine fish, seabird eggs, seal) show concentrations of PCBs in ranges that encompass or exceed numerous guidelines (International Joint Commission, US EPA, Environment Canada) for concentrations of PCB in fish set for protection of fish-eating wildlife (AMAP 1997; de March et al. 1997). In some cases, the range of concentrations or even mean concentrations of contaminants in predatory wildlife species approach or exceed No Observed Effect Levels (NOEL), Lowest Observed Effect Levels (LOEL) or Lowest Observed Adverse Effect Levels (LOAEL) reported for similar species. For example, the Arctic Monitoring and Assessment Report (AMAP 1997; de March et al. 1997) shows that eggs from several peregrine falcon species have mean PCB concentrations that exceed LOEL/LOAELs for egg mortality and deformities in cormorant, bald eagle and herring gull and for impaired reproduction and hatching success in the common tern and night heron (Figure 2). With the exception of eider ducks, Arctic seabirds and birds of prey show PCB mean levels or ranges which encompass threshold NOELs and LOEL/LOAELs for hatching success and deformities in chickens. Arctic marine mammals also have body burdens of PCBs in ranges that encompass or exceed known effect levels for mink and otter kit survival; mink litter size, liver and muscle changes; harbor seal and ringed seal reproduction; and rhesus monkey immune effects (AMAP 1997, de March et al. 1997) (Figure 3).

Finally, levels of POPs in the diet of human populations which heavily utilize upper food chain species can (through biomagnification) reach levels which are of concern to public health authorities, even in remote areas with no substantial local emission sources. For example, Figure 4 and Table 1 (from Kuhnlein et al. 1995) illustrate that several traditional foods for Arctic peoples are known to be consumed in quantities that, due to their concentrations of a number of POPs (e.g. Chlordane and Toxaphene), exceed 'Tolerable Daily Intake' levels used by Health Canada (Hansen et al. 1997, AMAP 1997). Chan et al (1997) report that in a community on Baffin Island, the 'high end consumers' have intake levels over 20 times the Tolerable Daily Intake level for chlordane and toxaphene. These levels of dietary exposure have resulted in correspondingly high tissue, blood and lipid levels in the relevant human populations. For example, in the case of marine mammal consumers (predominantly Inuit people), 40-65% of the women who participated in studies conducted in coastal communities of Canada's Northwest Territories have blood levels of PCB above, and up to five times higher, than values used by Health Canada to identify a 'Level of Concern'. In some cases, these levels are at or exceed those that have been associated (through in utero exposure) with intellectual impairment in children in studies conducted in the Great Lakes region (Jacobson and Jacobson 1996).

4) Some possible implications to other parts of the world

Our area of study is confined to the Arctic and we have not been able to undertake a comprehensive review of environmental exposure levels to POPs in other regions of the world. Never-the-less, for the purposes of this presentation, we suggest below that the Arctic is not unique in displaying environmental levels of POPs which approach various effects levels or other criteria used to express concern. However, we do stress the very cursory nature of this survey.

The ubiquitous occurrence of POPs in the global environment has been widely reported in the literature for a number of media. Differences in measurement technique, medium studied, and reporting methodology often make geographical comparisons difficult. One recent global study using a standardized method of tree bark sampling has been reported by Simnich and Hites (1995). They showed high concentrations of organochlorines in tree bark from the United States, Europe, India, the Middle East, Japan, Brazil, Australia, Taiwan, South Korea, and Russia. The highest concentrations were the HCH's, endosulfans, and DDE.

Biomagnification and persistence are key characteristics in the processes which elevate low POP levels in the abiotic environment to concentrations in upper trophic level animals and people which may be in the order of effects levels. Figures 5, 6, and 7 illustrate this for the North Sea, Baltic Sea, and Mediterranean Seas respectively, and are probably broadly representative of any aquatic food chain with several carnivorous trophic levels and an input of POPs (eg Ramesh et al. 1990).

In tropical areas, environmental exposure to POP pesticides will result from both near field and distant sources of these substances. One of the most commonly utilized indicators of human exposure is the presence of POPs in breast milk with the most frequently reported POP being DDT/DDE, for example in remote areas of Papua New Guinea (Spicer and Kereu 1993), and in India (Tanabe et al. 1990). In some cases, investigators have considered levels in terms of Acceptable Daily Intakes (ADI) set by FAO and WHO. For example, Waliszewski et al. (1996) using mean levels in Veracruz (Mexico) have calculated that the exposure of infants to S-DDT may exceed the ADI by 2 times and for HCB by 2.6 times.

Aquatic resources are usually thought of as the main dietary exposure route for humans. However, agricultural products such as cereals may also be a significant source. Thus in Nigeria, Osibanjo and Adeyeye (1995) have reported some degree of exceedence of FAO/WHO maximum residue limits in rice, maize and millet for several POPs including aldrin, dieldrin and S-DDT. Another significant potential exposure route particularly in tropical areas where POP pesticides are used for disease vector control is occupational exposure. Several studies have been conducted in the Veraruz region of Mexico (where DDT is used for Malaria control) to investigate occupational, and dietary exposure to DDT. The authors (Rivero-Rodriguez et al. 1997) reported high adipose tissue levels of S-DDT (Mean 67.41ug/g in the occupational group) and considered that these level represent a public health problem (Rivero-Lopez-Carrillo et al. 1996).

We believe that the above inadequate survey suggests that in many parts of the world, POP levels in aquatic systems and in human dietary sources may be in the general order of effects levels or of other values used by health authorities to express a level of concern.

Acknowledgements

In addition to drawing upon primary literature, we have freely utilized (and in some cases incorporated) segments of reviews completed by others. Our debt to these authors is substantial and we would particularly like to acknowledge the following: Bidleman et al 1997; United States Environmental Protection Agency, 1997; Swain et al, 1992; de March, 1997; and, AMAP 1997.

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