Effects of Pharmaceuticals on Aquatic biota - A Review

Current Topics in Toxicology 1: 73-86 (2004)

Effects of Pharmaceuticals on Aquatic biota - A Review Gagné, F., Blaise, C. St. Lawrence Centre, Environment Canada, 105 Mc Gill St., Montreal, Quebec, Canada H2Y 2E7. [email protected].

Manuscript published in Current Topics in Toxicology.

Running title: Effects of drugs on aquatic organisms.

Abstract Pharmaceutical, personal care and veterinary products, which have been found in wastewater and surface water, are likely to contaminate the aquatic environment, groundwater included. The purpose of this review is to examine current and new strategies for assessing the toxicological effects of this special class of xenobiotics on aquatic species. Aquatic sentinel species that bioaccumulate some of these drugs remain to be identified, but studies with mussels and plants have shown that some antibiotics significantly accumulate in tissues. Laboratory tests have been conducted with some success on several aquatic species, including bacteria, plants, invertebrates (molluscs and arthropods) and fish, with commonly found drugs both individually and in mixtures. These toxicity tests generally indicate that acute lethal effects are not likely to occur in the environment but that chronic or long-term effects are possible. In an attempt to measure the effects of pharmaceuticals and personal care products, two types of biomarkers are proposed. The first class, known as integrative biomarkers, consists of measuring ecologically relevant biomarkers that encompass the effects of drugs, such as oxidative stress or DNA damage. Biomarkers that have been shown to predict changes at both the individual and population levels, and that respond to these products, are particularly useful for taking into account the final effects of pollution on feral aquatic organisms. The second class of biomarkers, known as drug target-specific biomarkers, measures the state/integrity of drug targets likely to impede the organism’s health and reproduction. For example, prostaglandin synthase produces prostaglandins necessary to assist spawning in bivalves, and its activity could be blocked by non-steroidal antiinflammatory drugs such as acetylsalicylate and ibuprofen. Finally, two case studies are

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presented to exemplify the use of biomarkers to assess drug target specific interactions and tissue damage in aquatic species. In the first case study, primary cultures of rainbow trout hepatocytes were used to assess the cytotoxicity of carbamazepine (CBZ), a drug commonly found in municipal wastewater, at µg/L range, after exposure for 48 h at 18oC. Results showed that CBZ induced the activity of cytochromes P4503A4 and 2B6 (benzyloxyresorufin as the substrate), known biotransformation enzymes for this drug class (iminostillbene), and was highly correlated with lipid peroxidation and cell viability at environmentally relevant concentrations. Lipid peroxidation and cell viability are considered integrative biomarkers, while cytochrome P4503A4/2B6 activity is a targetspecific biomarker. The second case study concerns feral carp that had survived for four years in an aerated lagoon that treats domestic municipal effluent. Results showed that cytochrome P3A4 activity, as determined by dibenzyloxyfluorescein (another substrate specific for cytochromes P450 3A4, 3A5 and 2C9), was readily induced in the postmitochondrial supernatant of liver homogenates. ATP-dependent dopamine transport activity in synaptosome preparations of brain tissues was shown to be significantly reduced. Increased cytochrome P450-related activities and reduced dopamine uptake suggest the pharmacological effects of opiate-like substances. Preliminary findings thus indicate that some aquatic species are likely to accumulate some drugs and that they are likely to produce harmful effects on fish. Further research is needed to validate such biomarkers and to relate changes in drug targets with their residual levels in tissues.

Key words: Pharmaceutical products, early biological effects, drug ecotoxicology, emerging substances.

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1. Introduction Increasing evidence shows that pharmaceutical and personal care products (PPCPs) enter the aquatic environment via urban wastewater. They also contaminate nearby surface waters and sometimes groundwater, if they persist long enough in the environment [1, 2]. Such observations have raised many concerns about the occurrence, fate and toxicity of these products for aquatic organisms. Moreover, the reported levels of commonly used drug groups, such as antibiotics, are similar to those of pesticides and other organic micro-pollutants [3, 4]. Because these compounds are usually not tested for their environmental effects, their potential ecotoxicological impact on the environment needs to be better understood. PPCPs have been detected in a wide variety of environmental samples in concentrations ranging from parts per trillion (ng/L) to parts per billion (µg/L). It is therefore unlikely that pharmaceuticals will have acute or lethal effects in the environment, provided they do not significantly bio-accumulate. However, the lack of information about the fate and long-term effects of these compounds and/or their metabolites in the aquatic environment makes accurate risk assessment very difficult at present.

Municipal wastewater is the most important source of PPCPs for the aquatic environment. They are alleged to originate from human and animal drug consumption rather than from release by the pharmaceutical industry. In one study, the lipid-lowering agent clofibrate, released in municipal effluent, has been found in groundwater and tap water in Berlin [5]. In another study, trichlosan (antibiotic) and caffeine were found in

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municipal wastewater at concentrations not considered harmful to aquatic life [6]. Moreover, little is known about the potential interactive (i.e., cocktails of PPCPs) and cumulative effects of complex mixtures in aquatic ecosystems. Several common drug classes, such as antibiotics, are found in quantities similar to those of pesticides and research on this latter group could help guide environmental impact studies of PPCPs [3]. Veterinary medicines (antibiotics) persist in soils for days to years, depending on many variables, including temperature, pH and the presence of manure [7]. The persistence of veterinary medicines in soil, manure, slurry and water varies across and within major groups, so generalizing about their respective fates in the environment is not recommended. In wastewater treatment plant effluent, concentrations have been found in the range of 1 µg/L for carbamazepine (CBZ, an anti-convulsant), 0.1 µg/L for clofibrate (lipid-lowering agent), 1 µg/L for diclofenac (anti-inflammatory), 1 µg/L for ibuprofen (analgesic and anti-inflammatory), 0.2 µg/L for ketoprofen (anti-inflammatory) and 3 µg/L for naproxen (anti-inflammatory) [8]. CBZ and clofibrate were fairly persistent in surface lake water, while photo-transformation was identified as the main elimination process for diclofenac. Ibuprofen has a relatively high sorption coefficient toward particles, suggesting that elimination from the aquatic environment occurs by sedimentation. Naproxen and ketoprofen were eliminated mainly by phototransformation and biodegradation. Some

PPCPs, especially those found at

concentrations starting in the µg/L range in municipal wastewater, could be sufficiently persistent to contaminate groundwater. Pharmaceutically active polar compounds, such as iodinated contrast agents, primidone, CBZ and clofibrate, have been found in groundwater samples in Germany [9]. The abiotic photo-degradation of six drugs (CBZ,

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diclofenac, clofibric acid, ofloxacin, sulfamethoxazole and propranolol) has also been studied [10]. In bi-distilled water, CBZ and clofibrate had an estimated half-life (t1/2) of 100 days at the highest latitudes (50 degrees N) in winter, while sulphamethoxazole, diclofenac, ofloxacin and propanolol (a β-blocker drug) degraded with a t1/2 of about 2, 5, 11 and 17 days, respectively. It has also been shown that humic acids act as inner filters of CBZ and diclofenac, but as photo-sensitizers of sulphamethoxazole, clofibrate, oflaxocin and propanolol, thereby increasing their rate of photo-degradation.

These studies indicate that aquatic organisms are being exposed to pharmaceuticals in situ with unknown effects. Because many other types of chemicals are found in aquatic ecosystems (e.g., pesticides, polyaromatic hydrocarbons, heavy metals and surfactants), it is a real challenge to identify effects specifically caused by PPCPs. At the present time, it is more important to identify biomarkers that respond to drugs than to search for biomarkers that are highly specific to exposure to PPCPs. The purpose of this review is to highlight current strategies for assessing the ecotoxic effects of these products, with an emphasis on aquatic biota. First, laboratory-based studies are conducted to identify potentially harmful responses of drug-exposed test organisms and estimate environmental threshold effects for hazard assessment estimates. Second, a toxicological assessment strategy is proposed, based on biomarker measurements of selected organisms, in an attempt to assess the long-term effects of PPCPs on aquatic fauna. A biomarker is defined as one of several biochemical/physical measurements that provide information on a particular (ant)agonistic interaction between cells/tissues and their immediate environment. In this review, two classes of biomarkers, integrative and target-specific,

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are proposed for carrying out toxicological investigations of these products relating to the health status of exposed organisms. This dual-class approach is designed to focus on the toxicity effects of xenobiotics and the state of intracellular targets with which drugs were designed to interact.

Again, the objective of this review is to highlight current and new strategies for determining the effects of PPCPs in the aquatic environment. Particular attention is given to fish and bivalves but other species are also considered, given the paucity of information presently available on environmental impacts of PPCPs. First, laboratorybased toxicity studies are reported that attempt to assess the immediate environmental hazards of these compounds. Second, a biomarker scheme is proposed to assess the biological effects of these drugs on aquatic organisms exposed either in the laboratory or under field conditions in the vicinity of municipal wastewater treatment plants known to be major emitters of PPCPs. Two case studies are provided to highlight the use of integrative and target-specific biomarkers.

2. Bioassays of PPCPs The toxic effects of pharmaceutical products have been examined with many toxicity tests using a wide range of aquatic and terrestrial organisms and effects endpoints (Table 1). The chronic toxicity of the antibiotic metronidazole was tested with Selenastrum capricornutum yielding a 72 h EC10 of 20 mg/L. However, no acutely lethal effect was observed with Acartia tonsa or Brachydanio rerio [11]. The ecotoxic potential of a propanolol β-adrenergic receptor blocking agent was examined in aquatic organisms [12]. This study revealed that reproductive effects on invertebrates (Hyalella azteca,

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Daphnia dubia, Cerodaphnia dubia) occur at much higher concentrations (10 to 125 µg/L) than on medaka (0.5 µg/L). Another study that tested Hydra against 10 commonly prescribed drugs in the U.K. maintained that environmental concentrations are not likely to produce adverse (lethal) effects [13]. The tested drugs (ibuprofen, paracetamol, acetylsalicylic acid, amoxicillin, bendroflumethiazide, furosemide, atenolol, diazepam, digoxin and amlodipine) affected neither survival at concentrations of up to 1 mg/L for seven days, nor feeding and bud formation after 17 days [13]. However, the ability of dissected polyps to regenerate was inhibited by diazepam, digoxin and amlodipine at 10 µg/L. In general, pharmaceutical compounds appear to pose little risk if we rely solely on standard toxicity tests and hazard quotient risk calculation. However, little is known about the long-term effects of psycho-active compounds (i.e., fluoxetine) on reproduction and behaviour that could ultimately disrupt community structure in aquatic organisms [15].

The ecotoxic potential of prescription drugs was tested using a multitrophic approach with the cladoceran Daphnia magna, the chlorophyte Desmodesmus subspicatus and the macrophyte Lemna minor [15]. For most prescription drugs tested (i.e., diclofenac, ibuprofen, naproxen and acetylsalicylic acid), the effective concentration to produce 50% inhibition (i.e., EC50) was in the range of 10 to 100 mg/L, where the macrophyte Lemna minor displayed the most sensitivity. Again, measured toxicities of tested pharmaceuticals showed that acute/short term chronic effects from single substances in the aquatic environment are very unlikely. However, toxicity produced by combinations of various pharmaceuticals was stronger than expected, suggesting that interactive effects

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could occur. The anticonvulsant CBZ, commonly found in surface water and groundwater, was studied using six ecotoxicological model systems with 18 endpoints [16]. The test battery included Vibrio fischeri inhibition of bioluminescence, Daphnia magna immobilization, growth inhibition of the alga Chlorella vulgaris, micronuclei formation and root elongation of Allium cepa. Cell integrity, neutral red uptake, total proteins, MTT dye oxidation, lactate deshydrogenase leakage and glucose-6-phosphate activity were also studied in the salmonid fish cell line RTG-2 and Vero monkey kidney cells. The most sensitive bioassay was that undertaken with Vero monkey cells (EC50 = 19 µM), followed by Chlorella vulgaris, Vibrio fischeri, Daphnia magna, Allium cepa and RTG-2 cells. The authors also concluded that CBZ is not expected to produce acute toxic effects in aquatic biota but recommended that chronic and synergistic effects with other chemicals be examined. Anti-inflammatory drug mixtures appear to have additive effects to the point that mixture toxicity was considerable in Daphnia magna, in which single substances showed no or only very slight effects [17]. It seems that the toxicity of these drugs acts through non-polar narcosis, indicating that the octanol-water partition coefficient (kow) drives toxicity.

Overall, these toxicity studies revealed that chemicals exclusively designed to interact with mammalian receptors could also be toxic to non-target organisms such as algae, plants and invertebrates. However, a biomarker approach is needed to assess the longterm environmental impact of pharmaceuticals in the aquatic environment.

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3. Biomarkers for measuring the effects of drugs in aquatic organisms 3.1 Issues and conceptual considerations Pharmaceuticals are usually designed to be rapidly absorbed and eliminated in tissues while maintaining the desired biological activity. On the one hand, their log kow are usually inferior to three, and hence they tend to be poorly accumulated. On the other hand, aquatic organisms do not have the same drug metabolizing capacity as mammals, and may instead readily retain some PPCPs. Bioaccumulation studies in aquatic organisms are presently lacking for these compounds. However, one study showed that the blue mussel Mytilus edulis readily absorbed two veterinary antibiotics (oxytetracycline and oxolinate) and released the former more slowly than the latter with a t1/2 in the order of four days [18]. The bioaccumulation factor for oxytetracycline and oxolinate was 0.3 and 4.2, respectively, in agreement with the log kow of oxytetracycline (0.3) and oxolinate (1.7). Interestingly, mussel shells were found to accumulate the antibiotics more than tissues, with a t1/2 of 8.3 and 3.85 days for oxytetracycline and oxolinate, respectively. In a 10-day bioaccumulation study with the bryophyte Fontinalis antipyretica, the antibiotics oxolinate, flumequine and oxytetracycline were accumulated in plants with somewhat high bioaccumulation factors ranging between 75 and 450 with mean residence times that could reach 59 days [19]. In another study, the drug ivermectin, which is widely used in animals, including humans, had the ability to cross the blood-brain barrier of the marine teleost sea bream, Sparus aurata [20]. Its levels remained high 24 h after injection in fish. Based on the distribution of log kow values of drug classes, many drugs have log kow values > than 2, such as muscle relaxants, analgesics, corticosteroids, sedatives, antidepressants (including the anticonvulsant CBZ),

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macrolide antibiotics (eryhtromycine), antimalarials, antifungals and antihistamines [21]. This indicates that many drugs could be significantly bioaccumulated.

Thus, bioaccumulation studies are needed to assess exposure in both field and laboratory settings. Moreover, identification of target or sentinel species likely to accumulate significantly and/or be more susceptible to these PPCPs is also lacking. CBZ was proposed as an anthropogenic marker in municipal wastewater and surface water [22] because of its presence and persistence in the environment. This drug requires cytochrome P4503A4 and 2B6 for elimination [23], and animals lacking this system are at risk of accumulating it, perhaps to toxic levels. However, xenobiotic-mediated oxidative metabolism can be deleterious, as it increases the formation of reactive oxygen radicals and species, thereby contributing to oxidative stress.

The greatest challenge for developing parameters of effects for PCPPs resides in finding biomarkers that respond specifically to drugs while maintaining some ecological relevance. Drugs that can have an impact on the reproduction or immune systems and act through specific targets are of interest. For example, non-steroidal anti-inflammatory drugs are cyclooxygenase inhibitors, which prevent formation of prostaglandin precursors from arachidonate. Prostaglandin E2 is known to act on smooth muscle tonic contraction and to stimulate spawning in scallops [24]. The levels of PGF2α and PGE2 in the hemolymph and ovary increased at the end of sexual maturation during the spawning season [25]. However, the addition of an anti-estrogen prevented the increase of these prostaglandins, suggesting the involvement of estrogens in prostaglandin metabolism

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during sexual maturation. Thus, the presence of these drugs acting through the metabolism of arachidonate could have an impact on spawning activity in mussels. Spawning activity is naturally stimulated by serotonin and prostaglandins, and many drugs could act on serotonin/prostaglandin metabolism, such as serotonin reuptake inhibitors, monoamine oxidase inhibitors and serotonin synthesis inhibitors [26]. Serotonin is also involved during sexual differentiation, gametogenesis and spawning, and contaminants in municipal effluent appear to impede these processes [27, 28].

Risk assessment requires the evaluation of exposure from the environment and effects (hazard) characterization in target organisms. A risk factor could then be calculated in terms of the ratio between environmental concentration and a no-observable effect concentration, where a risk is apparent when the ratio exceeds a value of 1. However, a reported environmental concentration could be replaced by a tissue level, which is more closely related to biological effects. This approach is more relevant and practical for field studies, where PPCP residues could be measured in tissues. In this context, the identification of a sentinel species or group of species that are exposed (in the sense of the chemical finding its way in tissues) and accumulate pharmaceuticals would be of value to guide ecotoxicological studies. These studies are aimed at understanding the effects of exposure to this class of compounds using biomarkers or functional toxicity tests (i.e. receptor-based assays). The use of blue mussels as a potential bio-indicator species holds promise for drugs having kow values ≥ 2 [18]. Information on the use of sentinel species (i.e., bioindicator species) is presently lacking, and the possibility exists that aquatic vertebrates (fish and amphibians), which are widely used in aquatic toxicity

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studies, including toxicogenomics, may not be organisms of choice for these compounds. The possibility was raised that lower organisms may be more suited to act as sentinel species for environmental and human health effects required for risk assessment [29]. It was proposed that pesticides, whose polarity more closely matches those of pharmaceuticals than other industrial contaminants, could provide some hints concerning bioaccumulation and toxicological impacts. For example, although mussels and fish accumulated the pesticide p,p’-DDT equally, the proportion of the parent pesticide (DDT) was higher in mussel tissues (40-65%) than in fish tissues (30-45%), indicating that the parent compound resides longer in mussel tissues than in fish [30].

A growing number of receptors, known as orphan receptors, have no known function or ligand in cells but could play a role in environmental signalling in cells. For example, the orphan receptor SXR (steroid/xenobiotic receptor) recognized many classes of chemicals and was able to activate a biological response resulting in the expression of drug metabolizing enzymes [31]. Hence, risk assessment strategies based on three to four short-term toxicity tests on these highly biologically active compounds may not be appropriate [32]. Moreover, long-term effects could occur at much lower concentrations and follow different toxico-dynamics in mammals so that the application of safety factors for deriving no-observed effect concentrations in aquatic fauna may not be suitable. The pharmacological mode of action of these compounds (the internal dose at which the drug produces its effects) and the presence of receptor targets found in anticipated sentinel species should be taken into consideration as a starting point in risk assessment schemes.

3.2 Biomarkers to assess the impact of drugs

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Biomarkers that could delineate the effects of complex mixtures of pharmaceuticals in the environment can be divided into two categories: drug target-specific biomarkers, and integrative biomarkers, such as those affecting reproduction via endocrine disruption (Figure 1). Integrative (and usually non-specific) biomarkers measure the toxicological effects of pharmaceuticals without any assumption of specificity towards that class of xenobiotics. These are destined to take into account the overall effects of xenobiotics and possible impacts at higher levels of biological organization (i.e., populations). In contrast, target-specific biomarkers act through drug-specific modes of action. For example, the drug ethinylestradiol, used in oral contraceptives, can induce vitellogenin in fish that can lead to feminization [33]. Such biomarkers do not respond solely to pharmaceutical agents because of the inherent complexity of mixtures present in the environment. However, target-specific biomarkers have a higher probability of detecting the early biological effects of these drugs because they directly measure the activity /integrity of those drug-targets in exposed organisms.

3.2.1 Integrative biomarkers Xenobiotics (PPCPs reflect a special class of xenobiotics) that impede energy reserves (lipids, proteins and carbohydrates), energy consumption (i.e., oxidative phosphorylation during cell respiration) and oxidative stress are likely to produce significant impacts on health status, which could lead to significant changes at the population level as has been shown in daphnia [34] and zebra mussels [35]. In Daphnia magna, the cellular energy allocation (energy reserves vs. energy consumption) value was significantly correlated with the chronic 21-d effect values based on growth, survival and reproduction.

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Moreover, parameters related to oxidative stress and DNA damage were able to predict long-term effects at higher levels of biological organization, such as growth and net reproductive rate. Current evidence suggests that drugs usually require oxidative metabolism for elimination, frequently involving cytochrome P450 hemoproteins [36]. Thus, drugs have the potential to influence the redox properties in cells or tissues and lead to oxidative stress [37]. The increase in oxidative metabolism in aquatic species imparted by drugs could be taken into account by measuring oxidative stress, such as lipid peroxidation and oxygen-mediated DNA damage. In a study that compared population metrics and effects biomarkers in Mya arenaria clams, animals showing increased metallothionein levels and lipid peroxidation with altered vitellogenin-like proteins and gonado-somatic index were associated with reduced clam density and increased mean age values [38]. These studies suggest that oxidative stress (whether induced directly by contaminants or by increased oxidative metabolism), reproductive integrity, energy reserves and consumption are likely to predict changes at the population level. However, the question remains as to the extent that pharmaceutical products released to the environment could contribute to these integrative responses.

The effects of nine pharmaceutical compounds were examined in non-target aquatic organisms using two fish hepatocyte models: rainbow trout hepatocytes and PLHC-1 fish cell line [39]. The tested substances were clofibrate, fenofibrate, CBZ, fluoxetine, diclofenac, propanolol, sulfamethoxazole, amoxicillin and gadolinium chloride. Clofibrate, fenofibrate and fluoxetine were the most cytotoxic drugs producing oxidative stress before loss of cell viability. The authors of this study also found that basal EROD

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activity, a cytochrome P450 enzyme induced by coplanar aromatic compounds, was inhibited by fibrates, raising concerns about these drugs in the context of exposure to other environmental pollutants. Current evidence suggests that drug metabolism leads to oxidative stress in tissues since many of these drugs require phase 1 biotransformation reactions other than EROD (CYP1A1) activity. Drugs have a wide range of polarity, with octanol water partition coefficients (log kow) ranging from -4 to +4 [21]. Most drugs are oxidized by cytochrome P450s in mammals: 1A2, 2C9, 2C19, 2D6 and 3A4 [40, 41] and conjugated

by

glutathione

S-transferase,

UDPG

glucuronyl-transferase

and

sulfotransferase.

3.2.2 Drug target-specific biomarkers Studies of drug metabolism in fish have found more than 40 drugs in bile at higher concentrations than usually found in tissues or in the surrounding water [42]. This was particularly true for parent substances or their bio-transformed products with molecular weights greater than 400, indicating that bile is a good vector for elimination of heavy drug conjugates. Early studies have also shown that fish and invertebrates possess a cytochrome P450 dependent mono-oxygenase system where exposure to coplanar aromatic hydrocarbons has caused the induction of oxidative metabolism [43]. However, fish lacked the ability to metabolize phenobarbital, which requires cytochrome P4502C and 2B for metabolic oxidation. Considerable species differences are found in levels of activity of both phase 1 and phase II biotransformation enzymes, and it is difficult to predict how a new drug will be metabolized in aquatic organisms. Among the xenobiotic metabolizing cytochromes P450, five forms, CYP1A2, CYP2C9, CYP2C19, CYP2D6 and CYP3A4, appear to be most commonly responsible for the metabolism of drugs in

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humans [40]. Their respective activities could be readily measured by microtiter plate assays. Furthermore, digestive gland microsomes of Mytilus galloprovincialis from a polluted site showed higher levels of CYP1A, CYP2E and CYP4A, as determined by polyclonal antibodies raised against fish species [41]. Immuno-reactivity with at least five cytochrome P450s, CYP1A, CYP2B, CYP2E, CYP3A and CYP4A, in digestive gland microsomes was also present. This indicates that bivalves are equipped to metabolize drugs, with the exception of CYP2C. The lack of cytochrome P4502C in bivalves would limit the elimination of drugs such as barbiturates (phenobarbital), trimipramine (antidepressant), sulfaphenazole and sildenafil (Viagra) [4, 45, 46]. However, other compounds, such as aflatoxin B1, could also modulate cytochrome P450 activity [47] hence, biomarkers are not necessarily entirely specific to pharmaceuticals.

Cells are equipped with a P-glycoprotein located at the cytoplasmic membrane, which extrudes xenobiotics from the cell’s intracellular environment. This transporter protein was first identified as a multi-drug resistance pump in tumour cells that became resistant to anticancer drugs [48]. This pump represents a formidable defence mechanism by which cells could protect themselves by excluding potentially cytotoxic xenobiotics, including PPCPs. Many contaminants, however, induce this multi-xenobiotic resistant (MXR) protein in Mytilus edulis [49] but other contaminants that could block this MXR transporter were found by Smital and Kurelec [50] in sediment extracts at polluted sites, rendering the organism highly sensitive to other pollutants. These authors have shown that exposing mussels to polluted rivers enhanced the accumulation of rhodamine B dye or a carcinogenic aromatic amine in gills, indicating that contaminants could act as

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chemo-sensitizers by enhancing the bioaccumulation of toxic chemicals. Thus, it is theoretically possible that these chemo-sensitizers could enhance the accumulation potential of PPCPs and their toxicological effects in polluted environments. This was shown in another study where membrane vesicles treated with staurosporine, an MXR transporter blocker, enhanced single-strand DNA breaks upon exposure to 2acetylaminofluorene from the gills of the freshwater clam Corbicula fluminea [51]. Organic anion transporting polypeptides are another class of proteins involved in the extrusion of amphipathic organic solutes, such as steroid conjugates, numerous drugs and organic dyes [52, 53].

The discovery of morphine receptors in immunocytes and neurons has suggested a link between the immune system and the response of organisms to stress, infection and malignant transformation of cells [54]. The presence of morphine receptors has been identified in both vertebrates and invertebrates (bivalves), where immunocytes possess a specific receptor (µ3 receptor) capable of interacting with morphine-like substances [55]. The net effect of morphine on motility and phagocytic activity of leucocytes is inhibitory. Indeed, morphine-like and codeine-like substances were found in the pedal ganglia and hemolymph of the blue mussel Mytulis edulis [55]. Both substances were able to antagonize the action of either tumor necrosis factor α or interleukin 1 α on human monocytes and mussel immunocytes to stimulate chemotactic activity, cellular velocity, and adherence. Thus, morphine may have a role in calming or terminating the state of immune alertness. Morphine also blocks inhibition of gill ciliary activity by dopamine (i.e., the net effect of morphine is the maintenance of ciliary activity) and antagonizes dopamine-stimulated adenylate cyclase activity in mussels [56]. Freshwater mussels

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exposed to a primary-treated municipal plume for 60 days showed reduced phagocytosis activity [57]. Moreover, long-term exposure of these mussels to the same municipal effluent plume led to decreased dopamine in visceral ganglia with increased synaptosome uptake, suggesting reduced dopaminergic activity and perhaps increased spawning (serotonergic) activity (i.e, release of oocytes from the gonad and stimulation of ciliary activity in gills) [27]. Although many other chemicals could have played a role in the observed effects, the presence of morphine-like substances (codeine, the cough suppressant

dextrometamorphan,

loperamide

[anti-diarrheal],

morphine

sulphate

[analgesic]) would certainly explain them. The mode of action of loperamide (a morphine-like substance acting through the µ-receptor to reduce intestinal peristaltic action and increases anal sphincter tone. Thus, its presence in the aquatic environment could disrupt the neuro-hormonal control of gametogenesis and spawning activity in aquatic organisms. In another study, the chronic morphine treatment of rats significantly decreased morphine binding to membranes from the anterior basal forebrain without any changes in the affinity of receptors [58].

Mussels appear to possess benzodiazepine receptors that are physiologically linked to GABA (4-amino-butyric acid) and glutamate receptors [59]. Indeed, diazepam, a benzodiazepine receptor agonist, mimicked the effects of GABA to decrease ectodermal body contraction bursts, the number of bursts and endodermal rhythmic potential in Hydra. In another study, a statistically significant increase in the maximal number of binding sites (Bmax) without change in the dissociation constant (kd) in muscle and mantle tissues, was found in mussels from the polluted site [60], suggesting that benzodiazepine-

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binding potential is changed at polluted sites in Mytilus galloprovincialis mussels. However, the exact physiological consequence of this increase in binding activity is not clear. An increased number of binding sites in benzodiazepine could result in either reduced neurotransmitter levels or increased sensitivity towards glutamate or GABAmediated effects. In a study in Mya arenaria clams collected in the vicinity of a site contaminated by tributyltin compounds, gonad estradiol-binding potential was increased, with a significant increase in male/female ratio [61] suggesting that receptors are free by the lack of hormone production. However, other toxicants could block hormone binding to its receptor [62]. Exposure of lead to mitochondria preparations of the marine mollusc Mytilus galloprovincialis decreased the specific binding of isoquinoline carboxamide derivative PK 11195 and adenylate cyclase activity at 10 and 25 µM, respectively. These results suggest that other xenobiotics that are not of pharmaceutical origin, such as lead, produce significant changes in the peripheral benzodiazepine receptor density and cyclic AMP production in the mantle of mussels.

Municipal effluent has been recognized as a major emitter of estrogenic compounds to the aquatic environment. One such compound, 17α-ethinylestradiol (EE2) is often found in municipal wastewater [63]. EE2 is an ingredient in oral contraceptive pills and could interact with estrogen receptors in aquatic organisms, leading to the induction of the eggyolk precursor vitellogenin [33, 64]. In the latter study, juvenile rainbow trout were exposed to EE2 either through injection for seven days or water exposure for 14 days. Exposed trout had increased plasma alkali-labile phosphorus (an indirect measurement for vitellogenins) and plasma calcium, and were highly correlated with vitellogenin, as

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determined by enzyme-linked immunoassay. Long-term exposure to EE2 also leads to significant histological changes [65]. Exposure of new hatches to EE2 for 2-60 days caused a concentration-dependent depression of gametogenesis in both genders with severe kidney pathology, such as glomerular dilation/degeneration, fibrosis and tubule enlargement/necrosis, at a threshold of 10 ng/L EE2 in 60 days post-hatch zebrafish. The effects of EE2 were also examined in other aquatic organisms. In a multi-generation experiment with the amphipod Hyalla azteca, EE2 was found to produce significantly smaller second gnathopods in second generation males (F2) exposed from gametogenesis until adulthood to 0.1 and 0.32 µg EE2/L [66]. Moreover, post-F1-generation males had disturbed maturation of germ cells and spermatogenesis. Chironomid larvae (C. riparius) were resilient to EE2 exposure, as determined by survival and mouth deformity induction [67]. Mouthpart deformities were produced when exposure occurred during the endocrine regulated moulting stage.

4. Case studies Studies dealing with the early biological effects of PPCPs in aquatic animals are rare. In this review, two examples are examined dealing with both laboratory and field exposure settings. In the former, primary cultures of rainbow trout hepatocytes were exposed for 48 h at 18oC to the anti-epileptic drug CBZ, which is often found at the µg/L range in municipal wastewater. Changes in cytochrome P450 activity, oxidative state and cell viability were determined after the exposure period. In the field study, changes in intracellular drug targets were examined in carps after their long-term exposure in aerated lagoons receiving mainly domestic wastewater, hence containing PPCPs.

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4.1 Exposure of rainbow trout hepatocytes to CBZ Exposure of rainbow trout hepatocytes to CBZ, commonly found in surface water and groundwater near urban environments, resulted in significant loss of cell viability, accompanied with increased 7-benzyloxyresorufin (BzRes) dealkylase activity and lipid peroxidation (LPO), as shown in Figure 2. The increased metabolism of this drug leads to lipid peroxidation and reduced viability. Cell viability was significantly correlated with BzRes dealkylase activity (R = - 0.6; p = 0.02) and marginally correlated with oxidative damage as determined by LPO (R = - 0.42; p = 0.1). Moreover, BzRes dealkylase activity was significantly correlated with LPO damage (R = 0.6; p = 0.03), suggesting the role of cytochrome P450 3A4 and 2B6 mediated oxidative stress. The threshold concentration to induce CBZ in trout hepatocytes was 70 µg/L after a 48 h exposure time. Based on the probabilistic bio-concentration factor of CBZ in fish (about 5-90), this corresponds to the presence of environmentally relevant exposure concentrations of 0.7 to 14 µg/L in surface waters. In another study with Hydra, exposure to CBZ leads to the induction of oxidative metabolism, as determined by BzRes dealkylase and heme oxidase activities, leading to the formation of LPO [36]. This study also revealed a significant correlation between BzRes dealkylase activity and LPO in treated Hydra. Hence, these case studies reveal the use of CYP3A4 activity and LPO as integrative and target-specific biomarkers respectively, in that fish and Hydra respond readily to the pharmaceutical CBZ, which is often found in aquatic environments.

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4.2 Evidence of PPCP-related effects in the field Feral carps (Cyprinus carpio) were found in aerated lagoons that receive domestic wastewater from a city of about 15 000 inhabitants. These fish spent four years in these aerated lagoons and offered a unique opportunity to evaluate the state of some intracellular drug targets in the liver and brain (Figure 3). Five cytochrome P450 forms (CYP1A2, CYP2C9, CYP2C19, CYP2D6, and CYP3A4) appear to be most commonly responsible for the metabolism of drugs [40], and CYP3A4 accounts for most drugmetabolizing activity. The experimental hypothesis sought to determine whether exposure to drug residues remaining in the aerated lagoons could induce their expression, especially that of CYP3A4. The activity of CYP2C9, CYP3A4 and CYP3A5 was measured in the liver extracts of carps using dibenzyloxyfluorescein (DBF) as the substrate (Figure 3A). Enzyme activity in carps from the aerated lagoon was significantly induced in comparison to those collected from a reference pond. Results thus indicate the presence of xenobiotics able to induce the expression of the major drug-metabolizing cytochrome P450 (i.e., 3A family) in the aerated lagoons. However, this should be confirmed by chemical analyses because other, perhaps yet unknown, compounds might contribute to this induction. Opiates such as morphine are metabolized mainly by CYP3A4 in fish (N-demethylation of morphine) and are known to increase the effect of dopamine in brain tissue, as determined by inhibition of adenylate cyclase activity in dopamine-treated synaptosomes [68, 56]. ATP-dependent dopamine activity in synaptosomes prepared from carps collected in aerated lagoons was significantly inhibited (Figure 3B), again suggesting the presence of opiate-like substances in these domestic wastewater aeration ponds.

23

5. Concluding remarks More research on PPCPs is clearly needed to validate that current biomarkers respond to PPCPs and to develop more specific biomarkers, such as those measuring the activity or state of drug targets. A new emerging discipline, “safety pharmacology,” is developing, which involves studying the potentially undesirable pharmaco-dynamic effects of exposure to a substance on physiological functions [69]. It combines pharmacology and toxicology, and the issue of PPCPs in the environment is clearly primordial herein. Priority should be placed on reinforcing the links between pharmacologists and toxicologists on the ways to best optimize their cooperative research In this context, biological targets that govern the health status and reproduction of organisms are of particular interest for ecotoxicology.

In this review article, two types of biomarkers are proposed. The first deals with integrative biomarkers that respond to PPCPs as well as to other xenobiotics. For example, glutathione S-transferase, which catalyzes the conjugation reaction of reduced glutathione to many xenobiotics, including drugs, holds promise as an integrative biomarker because this enzyme family is induced by other environmental pollutants such as heavy polyaromatic hydrocarbons (e.g., benzo(a)pyrene). The second type of biomarker determines the state or activity of drug specific-targets that are of concern. For example, inhibition of serotonin reuptake by Prozac, fluvoxamine and paroxetine was shown to stimulate spawning in zebra mussels [26]. In parallel, a primary-treated municipal effluent was shown to possess serotonergic potential that could lead to

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increased ciliary activity and spawning [28]. The case study with carps collected in an aerated lagoon for domestic wastewater treatment, which showed increased DBF activity (CYP3A4) and reduced dopamine transport in brain synaptosome preparations, provides more compelling evidence of the presence of opiate-like substances in domestic wastewater. This warrants searching for opiate-like substances in tissues of exposed organisms. These biomarkers are not meant to replace chemical analysis, but they can be used to detect what might be termed an “effect signature” of PPCPs either acting through drug specific-targets or non-specific toxicity.

Toxicogenomic- and proteomic- based technology are also promising tools for examining the effects of PPCPs because they have the potential to measure changes in gene expression (mRNA and proteins) for hundreds, if not thousands, of selected genes. However, this technology is expensive and data interpretation is complex and time consuming. Moreover, DNA microarrays represent a snapshot of gene transcripts at a given time within a particular tissue while toxicity represents a continuum of effects that varies with respect to exposure conditions such as duration, dose and mode of entry [70]. Thus, validation studies are also needed to relate change(s) in gene expression to adverse effects. It is possible that drugs acting through narcosis (non-specific effect) may not elicit significant changes in gene expression pattern while still producing a biological effect. Increased protein expression could result from modification of the mRNA transcripts, increasing its t1/2 without increasing the levels of mRNA. For example, estradiol-17β was shown to increase both the levels and stability of mRNA transcripts of estrogen receptors and vitellogenin in rainbow trout [71]. Nevertheless, genomics should

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be able to help identify unsuspected effects in aquatic organisms and to better circumscribe biomarkers capable of producing genuine PPCP-based effects in aquatic organisms. Acknowledgements. Authors are grateful to the management of the St. Lawrence Centre for supporting the production of this review on emerging substances in the environment. The assistance of Jacqueline Grekin in the English editing of the text is also appreciated. 6. References [1] Metcalfe, C.D., Miao, X.S., Koenig, B.G., Struger, J. 2003, Environ. Toxicol. Chem., 22, 2881-2889. [2] Boyd, G.R., Reemtsma, H., Grimm, D.A., and Mitra, S. 2003, Sci. Total Environ., 311, 135-149. [3] Jones, O.A., Voulvoulis, N., and Lester, J.N. 2001, Environ. Technol., 22, 1383-1394. [4] Richardson, M.L., and Bowron, J.M. 1985, J. Pharm. Pharmacol., 37, 1-12. [5] Halling-Sorensen, B., Nors, N.S., Lanzky, P.F., Ingerslev, F., Holten-Lutzhoft, H.C., and Jorgensen, S.E., 1998, Chemosphere, 36, 357-393. [6] Kolpin, D.W., Furlong, E.T., Meyer, M.T., Thurman, E.M., Zaugg, S.D., Barber L.B., and Buxton, H.T. 2002, Environ. Sci. Technol., 36, 1202-1211. [7] Boxall, A.B., Fogg, L.A., Blackwell, P.A., Kay, P., Pemberton, E.J., and Croxford, A. 2004, Rev. Environ. Contam. Toxicol., 180, 1-91. [8] Tixier, C., Singer, H.P., Oellers, S., and Muller, S.R. 2003, Environ. Sci. Technol., 37, 1061-1068. [9] Heberer, T. 2002, Toxicol. Lett., 131, 5-17. [10] Andreozzi, R., Raffaele M., and Nicklas, P. 2003, Chemosphere, 50, 1319-1330. [11] Lanzky, P.F., and Halling-Sorensen, B. 1997, Chemosphere, 35, 2553-2561. [12] Huggett, D.B., Brooks, B.W., Peterson, B., Foran, C.M. and Schlenk, D. 2002, Arch. Environ. Contam. Toxicol., 43, 229-235. [13] Pascoe, D., Karntanut, W., and Muller, C.T. 2003, Chemosphere, 51, 521-528.

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[51] Waldmann, P., Pivcevic, B., Muller, W.E., Zahn, R.K., Kurelec, B. 1995, Mutat. Res., 342, 113-123. [52] Hagenbuch, B., and Meier, P.J. 2003, Biochem. Biophys. Acta, 1609, 1-18. [53] Bresler, V., Bissinger, V., Abelson, A., Dizer, H., Sturm, A., Kratke, R., Fishelson, L., Hansen, P.-D. 1999, Helgol Mar. Res., 53, 219-243. [54] Makman, M.H. 1994, Adv. Neuroimmunol. 4, 69-82. [55] Stefano, G. B., Digenis, A., Spector, S., Leung, M. K., Bilfinger, T. V., Makman, M. H., Scharrer, B., and Abumrad, N. N. 1993, Proc. Nat. Acad. Sci. USA 90, 11099-11103. [56] Stefano, G.B., and Leung, M. 1982, Cell. Mol. Neurobiol., 2, 347-352. [57] Blaise, C., Trottier, S., Gagné, F., Lallement, C., and Hansen, P.-D. 2002, Environ. Toxicol., 17, 160-169. [58] Simantov, R. 1993, Neurosci. Lett., 163, 121-124. [59] Kass-Simon, G., Pannaccione, A., and Pierobon, P. 2003, Comp. Biochem. Physiol., 136, 329-342. [60] Betti, L., Giannaccini, G., Nigro, M., Dianda, S., Gremigni, V., and Lucacchnin, A. 2003, Ecotoxicol. Environ. Saf, 54, 36-42. [61] Gagné, F., Blaise, C., Pellerin, J., Pelletier, E., Douville, M., Gauthier-Clerc, S., and Viglino, L. 2002, Comp. Biochem. Physiol. 134C, 189-198. [62] Giannaccini, G., Betti, L., Palego, L., Chelli, B., Gallo, A., Pirone, A., Fabiani, O., Bertellotti, S., and Lucacchini, A. 2004, Comp. Biochem. Physiol., 137C, 197-206. [63] Braun, P., Moeder, M., Schrader, S., Popp, P., Kuschk, P., and Engewald, W. 2003, J. Chromat., 988A, 41-51. [64] Vandenbergh, T., Versonnen, G.F., Arijs, B., and Janssen, C.R. 2002, Comp. Biochem. Physiol., 132C, 483-492. [65] Weber, L.P., Hill, R.L. Jr., and Janz, D.M. 2003, Aquat Toxicol., 63, 431-446. [66] Vandenbergh, G.F., Adriaens, D., Verslycke, T., and Janssen, C.R. 2003, Ecotoxicol. Environ. Saf., 54, 216-222. [67] Meregalli, G., and Ollevier, F. 2001, Sci. Total Environ., 269, 157-161.

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Ultimate effects Non-specific effects (narcosis)

Exposure Absorption/distribution (bioaccumulation) - Metabolism - Elimination Specific effects (functional toxicity)

Molecular

Cellular/tissues

Multixenobiotic Resistance

ImmunoModulation

Oxidative Metabolism and stress

Genotoxicity/ Mutagenicity

Protein/DNA Adducts

Apoptosis (Cell viability)

Receptor (des)activation and endocrine disruption

Individual

Population

Reproduction

-Sex differentiation -Gametogenesis -Fertilisation/ spawning

Behaviour -feeding Energy reserves/ -prey capture/ expenses avoidance

Survival Growth Age structure Reproductive Success (offspring)

Heath status and disease resistance

Figure 1. Exposure and effects of pharmaceutical and personal care products to non-target organisms.

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11000 Mean

BzRes dealkyase activity (fmole Resorufin/min/cell density) or Lipid peroxidation (ng TBARS/cell density)

8500 Stand. error

8000

Cell viability (Fluorescein retention)

7500

7000

6500

6000

5500

5000

4500

4000 0

12,5

25

50

10000

BzRes dealkylase Lipid peroxidation

9000 8000 7000 6000 5000 4000 3000 2000 1000 0

100

0

Carbamazepine (ug/L)

12,5

25

50

100

Carbamazepine (ug/L)

Figure 2. Exposure of rainbow trout hepatocytes to carbamazepine. Rainbow trout hepatocytes were exposed to carbamazepine for 48 h at 18oC. Cell viability (carboxyfluorescein diacetate uptake), benzyloxyresorufin (BzRes) dealkylase activity and lipid peroxidation (thiobarbituric acid reactants) were determined in cells.

32

0,68

A

0,66

Mean Stand. error

Dibenzyloxyfluorescein activity (nmole fluorescein/min/mg proteins)

0,64 0,62 0,60 0,58 0,56 0,54 0,52 0,50 0,48 0,46 0,44 0,42 0,40 0,38 Lagoon

Ref erence Site

B

0,022

Dopamine transport activity (nmole Pi formed/min/mg proteins)

0,024

0,020

Mean Stand. error

0,018 0,016 0,014 0,012 0,010 0,008 0,006 Lagoon

Reference Site

Figure 3. Effects of domestic wastewater exposure on selected drug targets in carps confined to an aerated lagoon for four years. Carps surviving in wastewater aerated lagoons were captured and tested for drug metabolizing activity (CYP3A4) in the liver and dopamine transport activity (ATP dependent) in brain synaptosome preparations.

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Table 1. Organisms selected for toxicity assessment of PPCP1.

Trophic level Decomposers

Primary producers

Test species

Drugs tested

Effects measured

Vibrio fischeri

Carbamazepine

Staphylococcus aureus Selenastrum capricornutum Lemna minor Desmodesmus subspicatus Chlorella vulgaris

Ibuprofen Metronidazole Clofibrinate, ibuprofen carbamazepine, diclofenac, Carbamazepine streptomycin metronidazole Oxolinate, flumequine, oxytetracycline Streptomycin

Inhibition of bioluminescence Growth inhibition Chronic toxicity (72 h) Growth inhibition

Fontinalis antipyritica

Invertebrates

Fish

Scenedesmus oblicus, Chlamydomonas reinhardii Allium cepa Hydra vulgaris

Hydra attenuata

Carbamazepine Ethynylestradiol, ibuprofen, paracetamol, acetylsalicylate, amoxicillin, bendroflumethiazide, furosemide, atenolol, diazepam, digoxin, amlodipine, Carbamazepine

Hyalella azteca

Propanolol

(Cerio)daphia dubia

Clofibrinate, ibuprofen carbamazepine, diclofenac, ethynylestradiol, furazolidone, aminosidine, bacitracin, erythromycin, lincomycin Ethynylestradiol Pharmaceutical waste

Potamopyrgus antipodarum Temora turbinata, Amphitoe valida Mytlilus edulis Acartia tonsa Medaka Rainbow trout (RTG-2 cell line) Fundulus heteroclitus

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Oxolinate, oxytetracycline Metronidozaole Propanolol Carbamazepine Ethynylestradiol

Growth inhibition

Bioaccumulation Growth inhibition Micronuclei (root) Mortality, polyp structure and regeneration, sexual reproduction

Acute toxicity (48 h) Induction of oxidative metabolism and conjugation Reproduction inhibition, survival Survival, immobilization, reproduction

Egg production Reduced growth and egg production Bioaccumulation Acute toxicicty (72 h) Reproduction inhibition Cell integrity and viability Embryotoxicity, morpholical abnormality

…Table 1 continued

Mammals

2

Oryzias latipes

Ethynylestradiol

Danio rerio

Ethynylestradiol

Brachydanio rerio Vero monkey cells

Metronidozaole Carbamazepine

Estrogen receptor and brain aromatase Vitellogenesis Gonad development Growth Sex ratio Acute toxicicty (72 h) Cell integrity and viability

1. This table is not intended to be exhaustive but lists the major studies conducted for the toxicological assessment of Pharmaceuticals and Personal Care Products (PPCPs). References are found in the text (Section 2). 2. Mammals in the context of non-target feral organisms. Those used in clinical settings were not considered here.

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