Search

Prediction of mitochondrial targeting, or prediction of exclusion from mitochondria, of small-molecule xenobiotics (biocides, drugs, probes, toxins) can be achieved using an algorithm derived from QSAR modeling. Application of the algorithm requires knowing the chemical structures of all ionic species of the xenobiotic compound in question, and for certain numerical structure parameters (AI, CBN, log P, pK a, and Z) to be obtained for all such species. Procedures for specification of the chemical structures; estimation of the structure parameters; and application of the algorithm are described in an explicit protocol.

The developing uterus is highly sensitive to a brief exposure to different substances, in particular those with endocrine-disrupting activity. Thus, exposure to environmental, nutritional, chemical, and other xenobiotic factors affecting signaling events during critical organizational periods can alter the normal course of uterine development with lasting consequences. In this chapter, we provide an experimental protocol to evaluate the development of the rat uterus as a toxicity biomarker at two different developmental time points: (1) the neonatal period, on postnatal day (PND) 8, and (2) the prepubertal period, on PND21. In this experimental approach, we propose to assess: (1) uterine morphology and cytodifferentiation, (2) uterine cell proliferation, and (3) the expression of proteins involved in uterine organogenetic differentiation. All these morphological and molecular markers are useful tools to determine the consequences of exposure to toxicants with the potential to disrupt the uterine development.

Intravital microscopy (IVM) is an essential experimental approach for evaluating, in real time, cell interactions in the blood and rheological parameters in the microcirculation of the living animals. Different tissues are surgically exposed to the visualization of the microvascular network in optical microscopies connected to video cameras and image software. By evaluating in situ microcirculatory network, IVM allows the visualization and quantification of physiological and pathological processes in the blood or in the adjacent tissues considering the whole system. Therefore, IVM has been used to evaluate the effects and mechanisms of actions in the microvascular network caused by pharmacological or toxic chemical agents. In this chapter, different experimental approaches are described to study the toxic effects and mechanisms of xenobiotics in the microcirculatory network.

Toxicology is a broad field that requires the translation of biochemical responses to xenobiotic exposures into useable information to ensure the safety of the public. Modern techniques are improving rapidly, both quantitatively and qualitatively, to provide the tools necessary to expand available toxicological datasets and refine our ability to translate that data into relevant information via bioinformatics. These new techniques can, and do, impact many of the current critical roles in toxicology, including the environmental, forensic, preclinical/clinical, and regulatory realms. One area of rapid expansion is our understanding of bioenergetics, or the study of the transformation of energy in living organisms, and new mathematical approaches are needed to interpret these large datasets. As bioenergetics are intimately involved in the regulation of how and when a cell responds to xenobiotics, monitoring these changes (i.e., metabolic fluctuations) in cells/tissues post-exposure provides an approach to define the temporal scale of pharmacodynamic responses, which can be used to guide additional toxicological techniques (e.g., "omics"). This chapter will summarize important in vitro assays and in vivo imaging techniques to take real-time measurements. Using this information, our laboratory has utilized bioenergetics to identify significant time points of pharmacodynamic relevance as well as forecast the cell's eventual fate.

Metabolism is the process of chemical transformation within a biological context. Metabolites comprise all the small-molecule (

The body is exposed to a variety of chemicals everyday in the form of pharmaceutical agents, household chemicals, dietary supplements, and environmental contaminants, many of which are extremely toxic. The primary defense mechanisms against xenobiotics in the body are the drug metabolizing enzymes (DMEs) involved in metabolism and excretion of xenobiotics [1]. Liver is the primary organ involved in the metabolism of xenobiotics including chemicals and pharmaceutical agents. Other organs including kidney and intestine have minor drug metabolism capabilities but liver is the site of metabolism for a vast majority of drugs and chemicals [2]. The chemical reactions involved in drug metabolism generally convert chemicals into more water-soluble metabolites rendering them easier for eventual excretion in the urine. Whereas many reactions catalyzed by the DMEs result in more water-soluble products, which are relatively less harmful, a number of the DME-mediated reactions result in the production of metabolic intermediates, which are highly reactive and induce tissue damage [3]. Therefore, the detoxification function of the liver is not homogenous and metabolism of each chemical in the liver should be investigated on a case-by-case basis. Nevertheless, there are some general principles that apply to the entire drug metabolism process.

Transgenic animal models are important tools, which have been extensively used to gain insight into the molecular mechanisms of human diseases. We have created two transgenic consomic rat strains, in which malignant hypertension can be induced by administration of a nontoxic xenobiotic. Studies combining quantitative trait locus (QTL) analysis, pathological examination and molecular biological investigation have demonstrated that the renal renin–angiotensin system (RAS), especially angiotensin converting enzyme (Ace), is a key factor affecting the degree of hypertensive kidney damage in these strains.

In the broadest sense, biologic markers are substances present in biologic specimens that when measured, help to determine the relationship between xenobiotic exposures and human diseases. The purposes served by biomarkers are twofold: firstly, to achieve the earliest identification of health impairment resulting from a xenobiotic exposure; and secondly, to gain insight into the mechanism(s) responsible for any adverse impact of such exposure on the health of individuals or specific populations at risk. The National Research Council through its Committee on Biological Markers has developed a model for conducting environmental health research which defines the progression from xenobiotic exposure to clinical disease and identifies four stages during that process [1]. These stages mark the course by considering the magnitude of the internal dose, its relationship to the biologically effective dose, the presence of early biologic effects and eventually on alterations in the structure and/or function of the target organ.

Over the last decade, great strides have been made in understanding the roles that the nuclear receptors PXR, CAR, PPARα, and AHR play in the induction of CYP genes. The ability of xenobiotics to bind and activate NRs to induce the expression of the CYP enzymes involved in their metabolism provides a mechanism by which an organism can mount an adaptive response to its changing chemical environment. The identification of endogenous ligands for some NRs indicates that these receptors play important roles in regulating CYP levels during physiological processes as well. It has become clear that the expression of many CYP genes is dependent on more than one NR. Recent studies have demonstrated that NRs often share xenobiotic ligands, response elements, and target CYP genes. The existence of multiple xenobiotic receptors with broad and sometimes overlapping functions likely increases the ability of an organism to detect and respond to a wide range of chemicals. The challenge for the future will be to understand how the NRs participate in a complex network to regulate CYP gene expression and to mediate the physiological response to xenobiotics.

Pseudomonads are ubiquitous γ-proteobacteria with a remarkable degree of physiological and genetic adaptability. Members of the genus Pseudomonas are found in large numbers in different natural environments (soil, freshwater, marine) as well as in association with plants and animals. These bacteria are involved in important metabolic activities in the environment, being element cycling and degradation of biogenic and xenobiotic pollutants some of their major tasks56, 64, 90, 96. The metabolic versatility of Pseudomonas strains has been used for biotechnological applications, mainly to degrade waste (bioremediation) and to synthesize specialty chemicals (biocatalysis)69, 99.

Chlorinated compounds are ubiquitous environmental pollutants due to their extensive use in industry, agriculture, and private households. Despite efforts to replace chlorinated chemicals by compounds that are of lesser environmental concern, large amounts of chlorinated chemicals are still produced for a variety of applications (44). In the early 1970s, the potential danger that chloroorganic chemicals pose to the environment and human health was recognized and consequently, production and handling operations were optimized and restricted to minimize human exposure, as well as to improve environmental quality. At the time, it was believed that the chemically synthesized chloroorganic compounds had no naturally occurring counterparts, and were therefore termed “xenobiotics” (from the greek, xenos = foreign, and bios = life). Research over the last two decades, however, has demonstrated that most, if not all, chloroorganic pollutants are also produced naturally, and are not “foreign to life”. Hence, the term xenobiotic is misleading, and should only be used to indicate that anthropogenic (from the greek, anthropos=people, and genesis = creation) activity has changed the concentration of a chloroorganic compound in certain environments. Significant amounts of chloroorganic compounds are of biogenic and geogenic origin (7, 28, 59-62, 76, 77, 146). Marine organisms, in particular seaweeds, sponges, ascidians, soft corals, and algae, commonly produce a variety of secondary metabolites, many of which are halogenated compounds (62, 72, 73). The function of these compounds for the producing organisms is largely unknown, however, in some cases haloorganic compounds may play a role as feeding deterrents (72,73). Abrahamsson et al. (1) demonstrated the biogenic formation of tetrachloroethene (PCE) and trichloroethene (TCE) by different species of temperate, subtropical, and tropical marine microand macroalgae. Other algal species have been shown to produce halogenated propanes (63), and cyanobacterial mats were implied to be a major source of a variety of halogenated aliphatic compounds (161). The natural global production of

The genetic polymorphism of drug-metabolizing enzymes has a major influence on the fate of xenobiotic substances, whether as drugs or absorbed from the environment. Our understanding of this polymorphism is important in order to evaluate the genetic predisposition for exposure-related risks and, in the future, to develop individualized drug therapy. Cytochrome P450 enzymes play a central role in the metabolism of many drugs, chemicals, and carcinogens. Differences in the activity of these enzymes are responsible for the inter-individual variability in drug response and toxicity (Bertilsson, 1995). Of the cytochrome P450 enzymes, the isoform CYP2C19 is of particular interest because of its high inter-individual and inter-racial differences (Goldstein et al., 1997; Kaneko et al., 1999; Griese et al., 2001).

Adaptive Metabolism: As early as the 1950s, it was observed that exposure to certain polycyclic aromatic hydrocarbons (PAHs) influenced the metabolism of structurally related xenobiotic chemicals. For example, administration of 3-methylcholanthrene (3MC) to rats led to a marked increase in hepatic microsomal 3-methyl-4-monomethylazobenzene demethylase and 4-dimethylaminoazobenzene reductase activities (1). The observation that the inducing PAHs were commonly substrates of the upregulated enzymes gave rise to the idea that this was an adaptive metabolic response. This early work also gave rise to important nomenclature that still colors this field. Most important in this regard is the term aryl hydrocarbon hydroxylase or AHH (2). In the early literature, this term referred to the enzymatic activities of a number of cytochrome P450-dependent monooxygenases that convert benzo[a]pyrene to hydroxylated metabolites (reviewed in (3, 4). Collectively, data from this early period demonstrated the existence of a mechanism for the adaptive metabolism of PAHs. This pathway appeared very similar to the adaptive metabolic pathways that had been previously described in simpler organisms and thus is commonly referred to as induction (5).

Efforts during the past 50 or more years to improve agricultural productivity and manufacturing processes have led to the introduction of numerous man-made chemicals into the environment. In the course of this chapter, these chemicals, generally, will be referred to as environmental toxicants or xenobiotics. It is now known that many of these chemicals have unexpected effects on the animal populations of the planet and one expects that such effects have occurred or will occur in the human population. The untoward effects that are of particular interest to us in the context of reproductive biology are effects that interfere with the normal development and function of the male and female reproductive systems and which lead to reduced fertility, infertility, or sterility. Although the mechanisms by which environmental toxicants cause their disruptive effects on reproduction were at first elusive, it is now clear that many act by interfering with the physiological regulation of reproductive processes by sex-steroid hormones. Other environmental toxicants act through other mechanisms, such as the thyroid hormone receptor and through the aryl hydrocarbon receptor (1, 2). This chapter will concentrate exclusively on xenobiotics that act through the sex-steroid hormone pathway since these hormones are known to be intimately involved in regulating reproductive processes.

There are only a few testing strategies directed specifically at female reproductive toxicity. (Generoso et al, 1971; Generoso and Cosgove, 1973; Bishop et al., 1997). There are some elegant molecular studies of female germ cell and reproductive biology (Eichenlaub-Ritter, et al., 1988; Albertini, et al., 2001; Hunt et al, 1995) but few, if any, of these investigations involve toxicological perspectives. Future research efforts in female, as well as male, reproductive toxicology would benefit greatly by enhanced integration of knowledge and methodology from molecular developmental and cell biology. In this chapter, I will review some basic reproductive biology, emphasizing differences between females and males, describe tests for female reproductive toxicology used by the National Toxicology Program (NTP) at NIEHS, summarize NTP test data and discuss some of the dogmas of germ cell mutagenesis that have recently been dispelled by these data.

The widespread use of synthetic chemicals by man has led to the accumulation of many persistent compounds in the marine environment. This contamination of the ocean, and the subsequent uptake of foreign compounds (xenobiotics) by marine biota, has led to a vast amount of research and related literature on their concentration in, and effect on, the marine environment at all levels of the food chain (e.g. Harding et al., 1997; Tatsukawa, 1992).

Organic anion transporters were first identified in intestine, liver, and kidney based on extensive studies of membrane transport phenomena mainly of xenobiotics. Recent advances in membrane transport studies with the use of molecular biological approaches have confirmed the participation of such carrier-mediated transport in these tissues and the pharmacological and/or pharmacokinetic relevance of these transporters to the disposition of xenobiotics. Furthermore, the tissue distributions of these transporters suggest the physiological significance of carrier- mediated transport of nutrients, endogenous substrates, and xenobiotics in many tissues in addition to intestine, kidney, and liver. Since at least some of the organic anion transporters probably have physiologically crucial roles, manipulation of drug disposition by utilization of these transporters is expected to be useful for regulating the pharmacokinetics of anionic drugs not only in intestinal absorption and renal and hepatic elimination, but also in distribution into pharmacological target tissues.

Although the levels of "naturally occurring" free radicals, such as NO and flavins, semiquinones and the like are present at quite low levels they are, in fact, essential for life. We are generally unaware of other free radical metabolism that occurs in biology with the exception of reactive oxygen species (ROS) such as Superoxide, hydroxyl radical and a variety of lipoxy radicals, which are quite short-lived and result from severe circumstances such as ischemia-reperfusion or hyperoxia. While the consequences of a burst of ROS are well appreciated and are the focus of an intense effort in antioxidant research, there may be several classes of more stable free radicals lurking in tissues and cells.

Biotransformation enzymes modify xenobiotic compounds into more hydrophilic products (metabolites) which can be efficiently excreted from the body. Phase 1 enzymes, of which the cytochromes P450 are the most important, perform functionalization reactions creating a reactive centre in the molecule through oxidation or reduction. Phase 2 enzymes, such as glutathione S-transferases, UDP-glucuronyltransferases, N-acetyltransferases and sulphotransferases couple an endogenous molecule to a functional moiety in the molecule or metabolite from the phase 1 reaction, generally leading to more hydrophilic metabolites that are easily excreted with urine or bile.

Over the past two decades it has become increasingly apparent that xenobiotic agents can directly influence transcriptional regulation of gene expression without producing detectable changes in the structure of DNA. The response of a cell to its environment is a dynamic process in which differentiation-driven paradigms for gene expression can be modulated by both endogenous and exogenous extracellular agents. Typical outcomes include: altered cell function, changes in proliferation and/or regulation of cell growth or cell death. Altered gene expression is often the result of a physiologically programmed interaction between the cell and a regulatory hormone or cytokine. Alternatively, altered gene expression can result from an interaction of the cell with a xenobiotic molecule. In many cases the consequences of these apparently disparate events, as well as their biochemistry, do not appear to be remarkably different. Our laboratory has been investigating the role of altered transcriptional regulation in the response of hematopoietic and immune cells to xenobiotic agents as one approach to understanding molecular mechanisms of toxicity and also with a view toward exploring whether altered patterns of transcriptional regulation can be used as a tool for predicting the target organ toxicity of xenobiotic agents. Results of these studies suggest that the effects of biological reactive intermediates on transcriptional regulation of gene expression are cell-type specific and that the response of individual target cells is at least in part determined by differentiation-dependent patterns of signal transduction and gene expression.

In humans and experimental animals, carcinogenesis is a complex process in which normal cell growth is modified. Carcinogenesis is divided into three main stages: initiation, promotion and progression. Chemopreventive interaction in carcinogenesis offers two major strategies. The first strategy will inhibit or at least slow down carcinogenesis by blocking its progress. This might occurr at all stages of carcinogenesis. This strategy includes the scavenging of bioreactive intermediates (BRIs), induction or inhibition of enzymes of the metabolism of xenobiotics which create or detoxify the BRIs, of enzymes of DNA-repair or of other enzymes. A second strategy aims at reversing the process of tumour formation either by redifferentiation of transformed cells or the elimination of precarcinogenic clones. Both strategies cover one or several steps of carcinogenesis and approach the subject matter in more general terms without consideration of individual susceptibility to particular cancers (Kelloff et al., 1999).

Research on human milk encompasses a wide range of topics including the motivation of mothers who choose to breastfeed, milk storage,milk banking and methods of analysis for important milk components.The occurrence of xenobiotic contaminants and medications in human milk is an important consideration in our industrial society. A new, but extremely complex area of research in human milk involves the relationship between modes of feeding, intestinal microflora and disease in infants.The following section contains examples of rich,emerging areas that offer expanding opportunities for further research in human milk and lactation.

The tripeptide glutathione (L-γ-glutamyl-L-cysteinylglycine) is found ubiquitous in microorganisms, plants and animals. In mammalian cells, where the tripeptide fulfils numerous functions, concentrations range from 0.5 to 10 mM (Meister & Tate, 1976; Meister, 1984; Redegeld et al., 1990). Glutahione is involved, for example, in the synthesis of proteins and DNA, in the regulation of enzyme activity, in the transport and reservoir of amino acids. A very important function of glutathione is the protection of cells, for instance as an antioxidant or as a co-factor in the conjugation of xenobiotics (Meister & Anderson, 1983; Redegeld et al., 1990).

The pulmonary immune system is composed of innate and acquired components. Innate (natural) immunity includes physical barriers, cells, and blood-borne macromolecules. These elements are present prior to exposure to a foreign substance (xenobiotic) and do not discriminate among (recognize) xenobiotics. The inflammatory response represents one form of natural immunity.

Trifluralin (TF) is a lipophilic, pre-emergent herbicide widely used in agriculture and known to bioconcentrate in fish. We have characterized the accumulation of TF in rainbow trout under a variety of experimental conditions. Our approach has been to use static water exposure systems and intravascular dosing in combination with clearance-volume pharmacokinetic (CV-PK) models to obtain quantitative estimates of uptake clearance, apparent volume of distribution and elimination due to xenobiotic metabolism. This paper will briefly discuss pertinent physicochemical data for TF and review the toxicokinetics of TF in rainbow trout. Emphasis will be placed on physiological interpretations of TF model parameters and practical aspects of modeling TF toxicokinetics with CV-PK models.

Pesticides, like other xenobiotics, are metabolized in both target and non-target spe cies by a number of different enzymes, including several isoforms of cytochrome P450 (P450) and the flavin-containing monooxygenase (FMO).1-3 While P450 and FMO can both as activation as well as detoxication enzymes for xenobiotics, P450 is the most im portant activating enzyme, producing electrophilic metabolites that react with endogenous cellular nucleophiles, including substituents on both proteins and nucleic acids.

Excretion of abundant compounds is an essential property of living organisms. Each day, a vast amount of material is taken up by the body that is of no nutritional value. Up take of such xenobiotics occurs mainly with the food but also by inhalation or transder-mally. Accumulation of these compounds would result in an enormous body burden. Thus, efficient mechanisms for the excretion of such compounds developed that have their roots very early in the evolution of life.

Biotransformation of relatively insoluble organic chemicals to more water-soluble compounds is a prerequisite for their detoxification and excretion. The first step in biotransformation is usually the oxidative step, catalyzed by the microsomal cytochrome P450 dependent mixed function oxidase (MFO) system. This “Phase I” metabolism is usu ally followed by “Phase II” in which oxygenated groups of xenobiotics are conjugated with glucuronate, sulfate, acetyl, or glutathione by different families of transferase en zymes. Thus, resulting polar and water-soluble end product can be excreted from the or ganism through bile or urine.1

Pharmacokinetic studies of xenobiotics are useful for quantitative characterization of their accumulation and elimination, bioavailability, bioconcentration and/or bioaccumulation, residue depletion, tissue distribution, and toxicity. Aspects of the study design that should be addressed in advance include identification of the desired information, test species (and size), route of administration, tissues and fluids to sample, and analytical methods for the test compound and metabolites. Small fish limit studies to a determination of the whole body concentration-time profile during and after bath or dietary administration. Whole body concentration-time and plasma concentration-time profiles may be determined in intermediate-size fish, and in large fish plasma and individual tissue concentration-time profiles are feasible using a variety of routes of administration. The principal pharmacokinetic parameters are clearance constants that characterize transfer of parent compound and metabolites among compartments, and the apparent volume of distribution of each compartment. Parameters derived from clearances and volumes of distribution include half-lives, bioavailability, and bioconcentration or bioaccumulation factor.

To investigate the involvement of prostaglandin H synthase (PHS) and lipoxygenase (LPO) in the activation of xenobiotics in platelets, platelet sonicates were preincubated with α-naphthol. Protein covalent binding of α-naphthol was measured following addition of arachidonic acid. Protein covalent binding was increased in a dose-dependent manner until it plateaued at 500 μM arachidonic acid. Pretreatment by two inhibitors of PHS, aspirin and indomethacin, resulted in a dose-dependent inhibition of α-naphthol-induced covalent binding, confirming PHS involvement. In addition, pretreatment by a LPO inhibitor, nordihydroguaiaretic acid (NDGA), also prevented covalent binding substantially, showing that LPO may be an alternative pathway for xenobiotic activation in platelets. Furthermore, combined treatment of aspirin and NDGA almost abolished the increase of aα-naphthol-induced covalent binding, suggesting that PHS and LPO are both major pathways for xenobiotic activation in platelets.

In the terrestrial environment, soils are the most important sinks for many anthropogenic xenobiotics. Airborne or otherwise introduced hydrophobic organic compounds such as polycyclic aromatic hydrocarbons (PAH) or polychlorinated biphenyls (PCB) accumulate mainly in the organic topsoils due to their high affinity for organic matter and their low water solubility. The further fate of these substances is dependent on their chemical structure and may include volatilization as the major pathway for the loss of lower chlorinated biphenyls from sewage sludge-treated soils (Alcock et al., 1993). Well documented is the biochemical degradation of low molecular weight PAHs in soils by autochthonous micro flora (Wild et al., 1990) or by introduced specialists such as certain genera of white rot fungi (Barr and Aust, 1994, Kastner and Mahro, 1996). However, 5- and 6-ring PAHs and PCBs seem to be largely resistant against biochemical attacks in the soil environment and are therefore regarded as highly persistent (Wild et al., 1990, Alcock et al., 1993). Furthermore, the bioavailability of organic chemicals, which is a prerequisite for degradation, can be greatly reduced due to incorporation into the macromolecular structures of humus molecules and the formation of so-called bound residues.

Xenobiotics are metabolized by many enzymes, including a number of isoforms of cytochrome P450 (P450), the flavin-containing monooxygenases (FMO), prostaglandin synthetase, alcohol and aldehyde dehydrogenases, molybdenum hydroxylases, esterases, as well as a number of transferases, particularly the glutathione S-transferases, the glucuronyl transferases and the sulfotransferases.1,2 Both activation and detoxication reac tions can be catalyzed by any of these enzymes but P450 is the most important with regard to activation of toxicants. While many of the reactions carried out by the FMO and P450 are similar there are differences between them. The most significant of these differences are that FMO does not, so far as is known, carry out oxidations at carbon atoms,3 prefering soft nucleophiles as substrates and carrying out oxidations at heteroatoms, such as nitro gen, sulfur, selenium and phosphorus, in organic molecules.

Comamonas testosteroni is a Gram-negative bacterium capable of growing on C19 to C27 steroids in the absence of other suitable carbon and energy sources (Marcus and Talala1956). Several steroid metabolizing enzymes, as well as steroid binding and transporting activities have been described and characterized (Pousette et al., 1986; Pousette and Catrom, 1984; Thomas et al., 1989; Watanabe and Watanabe, 1974; Me Donald Francis et al.,, 1985). In addition to these steroid catabolic features several other non-steroid metaboing activities such as polychlorinated biphenyl or aromatic hydrocarbon degradation were reported (Sondossi et al., 1992; Ahmad et al., 1990; Hollender et al., 1994). Both, transport system and degradative enzymes for either steroids or xenobiotic compounds are induced by the presence of these substances in the growth medium (Mobus et al., 1; Oppermann and Maser, 1996). At present, there is no information available on the genetic organization of the genes encoding steroid degradative enzymes.

Pharmacokinetic models used for estimating and interpreting the bioconcentration and bioavailability of xenobiotics in fish and other aquatic organisms are critically reviewed. Equilibrium partitioning, the simplest model for determining and interpreting bioconcentration,is based on the concept that xenobiotic distribution is thermodynamically driven. Although a number of factors can result in deviation from the model, the model works well for a number of lipophilic xenobiotics that are not metabolized. The accelerated bioconcentration model was developed as a method to determine bioconcentration factor for xenobiotics that require a long time to reach equilibrium. Model limitations include differing units for uptake and elimination, and representation of the fish as a single compartment. Clearance-volume models of bioconcentration are interpretable in terms of animal physiology. These models measure uptake and elimination in the same units and allow prediction of their relative importance to bioconcentration. Physiologically based pharmacokinetic models are complex models of xenobiotic distribution based on the anatomy and physiology of the fish. These models are most useful in their predictive value for different conditions and in interpretation of processes controlling accumulation of xenobiotic. Several alternate methods are available to determine bioconcentration factor without exposure to a constant concentration of a xenobiotic.

According to the present state of knowledge mutagenicity or genotoxicity of the ulti mate genotoxic agents ethylene oxide or styrene oxide cannot be increased by further me tabolism. However, in the present study we demonstrate that mutagenicity of several ultimate genotoxic substances is increased by human erythrocytes. For instance mu tagenicity of mafosfamide, N-nitroso-N-methylurea, ethylene oxide, and styrene oxide to Salmonella typhimurium TA 1535 was increased 5.5-, 5.1-, 2.7-, and 2.3-fold, respectively, by addition of human erythrocyte homogenate to the preincubation mixture in the Ames test. On the other hand, the mutagenicity of cumene hydroperoxide, benzo[a]pyrene-4,5-oxide, and 1-methyl-3-nitro-1-nitrosoguanidine was decreased approximately 3-, 5-, and 1000-fold by human erythrocytes, respectively.

Mineralization of organic molecules by microbes is essential for the carbon cycle to operate. The massive deposition of organic compounds in the environment, and the introduction of xenobiotics into the biosphere, lead to unidirectional fluxes, which result in the persistence of a number of chemicals in the biosphere, and thus constitute a source of pollution. Molecular biology offers tools to optimize the biodegradative capaCity of microorganisms and accelerate the evolution of “new” activities (Ramos et al., 1994). In several cases, genetically modified microorganims (GMMs) have been designed for the bioremediation of specific xenobiotic compounds and their deliberate release into the environment is expected (Ramos et al., 1995).

Xenobiotics are chemicals that are taken up into the body from exogenous sources, either ingested with the food we eat or imbibed with the water we drink, but may also be contained in the air we breathe and absorbed through our skin and lungs. Most of these chemicals are lipophilic, i.e., they have a greater solubility in lipid than in aqueous media. As a result they pose a potential problem in elimination, because the peritubular cells lin ing the renal tubules all have cell membranes composed of phospholipid and proteins. As the glomerular filtrate passes through the renal tubules, solutes contained in it are either selectively removed along with the water or become concentrated. Compounds with a suf ficiently high lipid/water partition coefficient are absorbed down the concentration gradi ent, dissolving into the cell membranes. From there they pass through the cells back into the body. In this simplistic model the only elimination of the compound would be that pre sent in the excreted water, the urine and this would be at the concentration present in the plasma and glomerular filtrate.

Dihydrodiol dehydrogenase (DD) [EC 1. 3. 1. 20] catalyzes the NADP+-linked oxidation of /trans-dihydrodiols of polycyclic aromatic hydrocarbons to corresponding catechols, and controls the formation of both their carcinogenic dihydrodiol epoxides (Oesch, et al.,, 1984) and cytotoxic o-quinones through autoxidation of the catechol metabolites (Flowers, et al.,., 1996). In addition, DD in mammalian liver is implicated in the metabolism of xenobiotic carbonyl compounds, steroids and prostaglandins because of its broad substrate specificity (Penning, et al.,, 1986; Hara, et al.,, 1986; Ohara, et al.,, 1994; 1995).

Metabolic gene polymorphisms encode for enzymes which are involved in both metabolism and conjugation of environmental as well endogenous compounds. Some of the products of the metabolic process are carcinogens. Two main categories of metabolic genes are known: Phase I genes, which include CYP1A1, CYP2E1, CYP2D6, and Phase II genes, such as GSTM1, GSTT1, NAT2. Polymorphisms have been described in these genes, with different frequencies according to ethnicity and geographic area.1 Several case-control studies have been conducted to study the association between metabolic gene polymorphisms and cancer of various sites, with special focus on lung and bladder cancer. It has been suggested that these genes play a role in cancer risk only when they interact with environmental exposure, since the substrates of their gene products are xenobiotic chemicals or their metabolites.2 This form of gene-environment interaction (GEI) has been described as “Type 2” GEI by Khoury3 and Ottman.4 According to this model, the presence or absence of the genetic risk factor is irrelevant for disease causation, if there is no exposure to an environmental agent. When the dose of environmental exposure (such as smoking) is analyzed with respect to metabolic susceptibility gene polymorphisms, two patterns are seen. In one case, a low exposure-gene (LEG) effect is observed, in which a decreasing degree of interaction occurs as a function of increasing exposure dose. A high exposure-gene (HEG) effect is observed when there is an increased degree of interaction as a function of exposure dose.2

The Fate of Xenobiotics in the Body Enzymes of Metabolism J.B. Schenkman. Protein-Protein Interactions in the P450 Monooxygenase System J.B. Schenkman, et al. Structures of Mitochondrial P450 System Proteins I. Hanukoglu. Biochemical Aspects of Flavin-Containing Monooxygenases (FMOs) E. Hodgson, et al. Expression and Regulation of Flavin-Containing Monooxygenases R.M. Philpot, et al. Correlations between in Vivo and in Vitro Studies in Human Drug Metabolism M. Eichelbaum. Pharmacogenetics: Polymorphisms in Xenobiotic Metabolism F.J. Gonzalez. Clinical Aspects of Polymorphic Drug Metabolism in Humans M. Eichelbaum. Genetic Polymorphisms of Cytochromes P450 1A1 and 2E1 and Glutathione S-Transferase M1 and Cancer Susceptibility in the Human M. Watanabe. Cytochrome P450 Isoforms: Insecticide Metabolism in Insects and Mammals and Role in Insecticide Resistance E. Hodgson, et al. Inhibitors of CYP51 as Antifungal Agents and Resistance to Azole Antifungals S.L. Kelly, et al. The Role of Oxidative Drug Metabolizing Enzymes in Liver and Lung Specific Toxicity R.M. Philpot. Importance of Drug Metabolism in Drug Discovery and Development R. Gasser. 7 additonal articles. Index.

Any study of the effects of pollutants on the endocrine system must begin with some knowledge of the likely concentrations to which a particular tissue is exposed in fish in polluted habitats. This is not as simple as might first appear, since uptake of the pollutant, its distribution between different organs and its clearance rates differ not only between species and pollutants, but also between sexes and seasons and is influenced by a multitude of other factors. It will also depend upon whether it is administered via the holding water, or in the diet. These will all affect the biomagnification factor, the ratio of the concentration in tissue to that in the holding water. Exposing a top carnivore to the water concentration prevalent in its habitat is unlikely to mimic the effects in wild fish held in such waters since they consume a diet of smaller fish or invertebrates which in their turn have already accumulated the pollutant through their diet. It might be expected that concentrations of pollutants would be higher in predator than in prey fish, but while a number of studies (Tables 3.1–3.5) have examined concentrations in several species from the same locality, they have rarely provided any information on the diets of the different species which could explain the differences in xenobiotic uptake. Any exposure regime should therefore start with a knowledge about the basic biology of the fish and the most probable source from which it accumulates the chemical of concern. It is then posssible after exposure to compare the concentration in the tissues of interest with that in wild caught fish in the polluted habitat. With in vitro experiments, the situation is much more complex since it is not at all clear whether a tissue concentration of 10 μg/g, for example, in a fish from a polluted habitat equates to adding 10μg to an incubation of 1 g of tissue or to 1 g of incubation medium, and in neither case does this relate to pollutant which may be localised within certain cellular components.

The liver is the major organ of detoxification in all vertebrates and would therefore be expected to play a major role in determining the sensitivity of an organism to a particular pollutant. In fish it is particularly rich in lipids and it is therefore not surprising to find that it is also a major site of bioaccumulation for many organic pollutants (see Chapter 3). Such bioaccumulation increases the risk both of morphological damage to the liver, and of disruption to the activities of detoxifying enzymes. These changes may in turn alter the rate at which further accumulations of the toxin can be deactivated and removed from the organism. This could result in synergisms between different pollutants if the disruption of the deactivating systems caused by one pollutant results in failure to detoxify the second. The hepatic enzymes used for detoxifying xenobiotics are in fact closely related to the enzymes that are used for both synthesis and deactivation of other chemicals naturally present in the body. P450 and conjugating enzymes, for example, play a major role in the deactivation of steroid hormones and if these enzymes are preferentially used to metabolise a pollutant, the hormones may be less rapidly deactivated resulting in a higher plasma concentration. They are also essential for the synthesis of steroid hormones by the interrenal and gonads. The situation is further complicated since the liver is both the target tissue and site of deactivation of the ovarian steroid estradiol. As a target tissue it has estrogen receptors which specifically induce synthesis of the yolk protein vitellogenin. The balance between binding to such receptors and deactivation could be affected by changes both in the deactivating enzymes and in the structural integrity of the liver and its constituent cells.

Two types of parameters have to be considered in assessments of chemical contaminants in the environment: exposure-related factors that determine the spatial and temporal abundance of the contaminants, and effects-related endpoints that indicate the potential toxicity. The observed interactions and effects are assumed to be related directly to the concentration of the compounds at the site of action. By appropriate combination of the respective information, a probability of negative impacts can be obtained - risk assessment (Chapter 9). The most toxic agents are not necessarily the most dangerous ones. The statement by Paracelsus that ‘sola dosis facit venenum’ still holds true (Ariens, Mutschler and Simonis, 1978); any substance, xenobiotic or biogenetic, will be toxic if applied in a sufficiently high dose.