75 4 Distribution and Effects of Chemicals in Communities and Ecosystems 4.1 INTRODUCTION In Chapter 3, the distribution of environmental chemicals through compartments of the gross environment was related to the chemical factors and processes involved, and models for describing or predicting environmental fate were considered. In the early sections of the present chapter, the discussion moves on to the more complex question of movement and distribution in the living environment—within individuals, com- munities, and ecosystems—where biological as well as physical and chemical factors come into play. The movement of chemicals along food chains and the fate of chemi- cals in the complex communities of sediments and soils are basic issues here. Ecotoxicology deals with the study of the harmful effects of chemicals in ecosys- tems. This includes harmful effects upon individuals, although the ultimate concern is about how these are translated into changes at the levels of population, community, and ecosystem. Thus, in the concluding sections of the chapter, emphasis will move from the distribution and environmental concentrations of pollutants to consequent effects at the levels of the individual, population, community, and ecosystem. The relationship between environmental exposure (dose) and harmful effect (response) is fundamentally important here, and full consideration will be given to the concept of biomarkers, which is based on this relationship and which can provide the means of relating environmental levels of chemicals to consequent effects upon individuals, populations, communities, and ecosystems. 4.2 MOVEMENT OF POLLUTANTS ALONG FOOD CHAINS The pollutants of particular interest here are persistent organic chemicals—com- pounds that have sufciently long half-lives in living organisms for them to pass along food chains and to undergo biomagnication at higher trophic levels (see Box 4.1). Some compounds of lesser persistence, such as polycyclic aromatic hydrocar- bons (PAHs) (Chapter 9), can be bioconcentrated/bioaccumulated at lower trophic levels but are rapidly metabolized by vertebrates at higher levels. These will not be discussed further here, where the issue is biomagnication with movement along the © 2009 by Taylor & Francis Group, LLC 76 Organic Pollutants: An Ecotoxicological Perspective, Second Edition entire food chain. The best studied examples of this are the organochlorines (OCs) dieldrin and p,pb-DDE (see Chapter 5), and the PCBs (see Chapter 6), where concen- trations in the tissues of predators of the highest trophic levels can be 10 4 –10 5 -fold higher than in organisms at the lowest trophic levels. Other examples include poly- chlorinated dibenzodioxin (PCDDs), polychlorinated dibenzofurans (PCDFs), and some organometallic compounds (e.g., methyl mercury). Biomagnication along terrestrial food chains is principally due to bioaccumu- lation from food, the principal source of most pollutants (Walker 1990b). In a few instances, the major route of uptake may be from air, from contact with contaminated surfaces, or from drinking water. The bioaccumulation factor (BAF) of a chemical is given by the following equation: Concentration in organism/concentration in food = BAF Biomagnication along aquatic food chains may be the consequence of biocon- centration as well as bioaccumulation. Aquatic vertebrates and invertebrates can absorb pollutants from ambient water; bottom feeders can take up pollutants from sediments. The bioconcentration factor (BCF) of a chemical absorbed directly from water is dened as Concentration in organism/concentration in ambient water = BCF One of the challenges when studying biomagnication along aquatic food chains is establishing the relative importance of bioaccumulation versus bioconcentration. The processes that lead to biomagnication have been investigated with a view to developing predictive toxicokinetic models (Walker 1990b). When organisms are continuously exposed to pollutants maintained at a fairly constant level in food and/ or in ambient water/air, tissue concentrations will increase with time until either (1) a lethal concentration is reached and the organism dies or (2) a steady state is reached when the rate of uptake of the pollutant is balanced by the rate of loss. The BCF or BAF at the steady state is of particular interest and importance because (A) it rep- resents the highest value that can be reached and therefore indicates the maximum risk, (B) it is not time dependent, and (C) the rates of uptake and loss are equal, thereby facilitating the calculation of the rate constants involved. BCFs and BAFs measured before the steady state is reached have little value because they are dependent on the period of exposure of the organism to the chemi- cal, and thus may greatly underestimate the degree of biomagnication that is possible. This statement should be qualied by the reservation that there may be situations in which the duration of exposure cannot be long enough for the steady state to be reached, for example, where the life span of an insect is very short. The principal processes of uptake and loss by different types of organisms are indicated in Table 4.1 (see also Box 4.2). A rough indication of the relative importance of different mechanisms of uptake and loss is given by a scoring system on the scale +−> ++++. Within each category of organ- ism there will be differences between compounds in the relative importance of differ- ent mechanisms, for example, due to differences in polarity and biodegradability. © 2009 by Taylor & Francis Group, LLC Distribution and Effects of Chemicals in Communities and Ecosystems 77 BOX 4.1 PERSISTENT ORGANIC POLLUTANTS (POPS) A list of hazardous environmental chemicals, sometimes referred to as the “dirty dozen,” has been drawn up by the United Nations Environment Programme (UNEP). These are: POP Year of Introduction Classification Aldrin 1949 Insecticide Chlordane 1945 Insecticide DDT 1942 Insecticide Dieldrin 1948 Insecticide Endrin 1951 Insecticide and rodenticide Heptachlor 1948 Insecticide Hexachlorobenzene 1945 Fungicide Mirex 1959 Insecticide Toxaphene 1948 Insecticide and acaricide PCBs 1929 Various industrial uses Dioxins 1920s By-products of combustion, for example, of plastics, PCBs Furans 1920s By-products of PCB manufacture The selection of these compounds was made on the grounds of their tox- icity, environmental stability, and tendency to undergo biomagnication; the intention was to move toward their removal from the natural environment. In the REACH proposals of the European Commission (EC; published in 2003), a similar list of 12 POPs was drawn up, the only differences being the inclusion of hexachlorobiphenyl and chlordecone, and the exclusion of the by-products, dioxins, and furans. The objective of the EC directive is to ban the manufac- ture or marketing of these substances. It is interesting that no fewer than eight of these compounds, which are featured on both lists, are insecticides. TABLE 4.1 Principal Mechanisms of Uptake and Loss for Lipophilic Compounds Mechanisms of Uptake Mechanisms of Loss Habitat/Type of Organism Diffusion From Food From Ingested Water Diffusion Metabolism Aquatic Mollusks ++++ + ++++ Fish ++++ +n +++ ++++ +n +++ Terrestrial Vertebrates ++++ < ++ ++++ © 2009 by Taylor & Francis Group, LLC 78 Organic Pollutants: An Ecotoxicological Perspective, Second Edition BOX 4.2 MODELS FOR BIOCONCENTRATION AND BIOACCUMULATION As indicated in Table 4.1, aquatic mollusks present a relatively simple picture because they have little capacity for biotransformation of organic pollutants, the principal mechanism of both uptake loss being diffusion. It is not sur- prising, therefore, that bioconcentration factors (BCFs) for diverse lipophilic compounds, measured at the steady state, are related linearly to log K ow val- ues (Figure 4.1). Thus, the more hydrophobic a compound is, the greater the tendency to partition from water into the lipids of the mollusk. The relation- ship shown in Figure 4.1 has been demonstrated in several species of aquatic mollusks, including the edible mussel (Mytilus edulis), the oyster (Crassostrea virginica), and soft clams (Mya arenaria) (Ernst 1977). A similar relationship has also been found with rainbow trout and other sh for some pollutants. On the other hand, some organic pollutants do not t the model well (Connor 1983). It seems probable that some compounds that are metabolized relatively rapidly by sh will be eliminated faster than would be expected if diffusion were the only process involved (Walker 1987). Such compounds would not be expected to follow closely a model for BCF based on K ow alone. This point aside, K ow values can give a useful prediction of BCF values at the steady state for lipo- philic pollutants in aquatic invertebrates. A great virtue of the approach is that K ow values are easy and inexpensive to measure or predict (Connell 1994). Other more complex and sophisticated models have been developed for sh (see, for example, Norstrom et al. 1976) but are too time-consuming/expensive to be used widely in environmental risk assessment where cost-effectiveness is critically important. Modeling for bioaccumulation by terrestrial animals presents greater problems, and BAFs cannot be reliably predicted from K ow values (Walker 1987). For example, benzo[a]pyrene and dieldrin have log K ow values of 6.50 and 5.48, respectively, but their biological half-lives range from a few hours in the case of the former to 10 –369 days for the latter. Endrin is a stereoisomer of dieldrin with a similar K ow , but has a half-life of only 1 day in humans, compared with 369 days in the case of dieldrin. These large differences in persistence have been attributed to differences in the rate of metabolism by P450-based monooxygenases (Walker 1981). Effective predic- tive models for bioaccumulation of strongly lipophilic compounds by terrestrial animals need to take account of rates of metabolic degradation. This is not a straightforward task and would require the sophisticated use of enzyme kinet- ics to be successful. In one model, it has been suggested that Lineweaver–Burke plots for microsomal metabolism might be used to predict BAF values in the steady state (Walker 1987) (Figure 4.2). In principle, when an animal ingests a lipophilic compound at a constant rate in its food, a steady state will eventually be reached where the rate of intake of the compound is balanced by the rate of its metabolism. It is assumed that the rate of loss of the unchanged compound by direct excretion is negligible. Primary metabolic attack upon many highly © 2009 by Taylor & Francis Group, LLC Distribution and Effects of Chemicals in Communities and Ecosystems 79 The main points to bring out are as follows: 1. The uptake and loss by exchange diffusion is important for aquatic organ- isms but not for terrestrial ones. 2. Metabolism is the main mechanism of loss in terrestrial vertebrates, but is less important in sh, which can achieve excretion by diffusion into ambi- ent water. 3. Most aquatic invertebrates have very little capacity for metabolism; this is particularly true of mollusks. Crustaceans (e.g., crabs and lobsters) appear to have greater metabolic capability than mollusks (see Livingstone and Stegeman 1998; Walker and Livingstone 1992). The balance between competing mechanisms of loss in the same organism depends on the compound and the species in question. In sh, for example, some compounds lipophilic compounds (e.g., polyhalogenated aromatic compounds and PAHs) takes place predominantly in the endoplasmic reticulum, particularly that of the liver in vertebrates. Thus, microsomes (especially hepatic microsomes of verte- brates) can serve as model systems for measuring rates of enzymic detoxication. Lineweaver–Burke and similar metabolic plots can relate concentrations of pol- lutants in microsomal membranes to rates of metabolism. In the steady state, rate of intake of chemical should equal rate of metabolism in the membranes of the endoplasmic reticulum. The concentration of the chemical required in the membranes to give this balancing metabolic rate can be estimated from the Lineweaver–Burke plot. The necessary balancing metabolic rate can be calcu- lated from the dened rate of intake in food, and then the microsomal concen- tration that will give this rate can be read from the plot. Thus, the concentration in endoplasmic reticulum can be compared to the dietary concentration to give an estimate of BAF. Estimates can also be made of BAF for the liver or the whole body if approximate ratios of concentrations of chemical in different compartments of the body when at the steady state are known. Log K ow Log BCF FIGURE 4.1 Relationship of BCF to log K ow values. © 2009 by Taylor & Francis Group, LLC 80 Organic Pollutants: An Ecotoxicological Perspective, Second Edition that are good substrates for monooxygenases, hydrolases, etc., can be metabolized relatively rapidly even though they, as a group, have relatively low metabolic capac- ity (Chapter 2). So, in this case metabolism as well as diffusion is an important fac- tor determining rate of loss. By contrast, many polyhalogenated compounds are only metabolized very slowly by sh, so metabolism does not make a signicant contribu- tion to detoxication, and loss by diffusion is the dominant mechanism of elimination. Some further aspects of detoxication by sh need to be briey mentioned. When sh inhabit polluted waters, exchange diffusion occurs until a steady state is reached, and no net loss will occur by this mechanism unless the concentration in water falls. When a recalcitrant pollutant is acquired from prey, digestion can lead to the tissue levels of that pollutant temporally exceeding those originally existing while in the steady state. Here, diffusion into the ambient water may provide an effective excre- tion mechanism in the absence of effective metabolic detoxication. Seen from an evolutionary point of view, the requirements of sh for metabolic detoxication would appear to have been limited on the grounds that loss by diffusion would often have prevented tissue levels becoming too high. The poor metabolic detoxication sys- tems of sh relative to those of terrestrial omnivores and herbivores are explicable on these grounds (Chapter 2). However, the advent of refractory organic pollutants, which combine high toxicity with high lipophilicity, has exposed the limitations of existing detoxication systems of sh. The very high toxicity of compounds such as dieldrin and other cyclodiene insecticides to sh was soon apparent, with sh kills occurring at very low concentrations in water (see Chapter 5) and metabolically resistant strains of sh being reported in polluted rivers such as the Mississippi. Gut Liver Peripheral tissues Redistribution Metabolism and excretion (a) (b) R U C L R M 1 K m 1 V max 1 v 1 s FIGURE 4.2 (a) A bioaccumulation model for terrestrial organisms. A kinetic model for liver. R U , rate of uptake from the gut; R M , rate of metabolism in liver; C L , concentration of pollutant in liver. The arrows indicate the routes of transfer of pollutant within the animal. The rates of uptake and metabolism are expressed in terms of kilograms of body weight. The nal elimination of water-soluble products (metabolites and conjugates) is in the urine. (b) Lineweaver–Burke plot to estimate the bioaccumulation factor; V max and v are expressed as milligrams of pollutant metabolized per kilogram of body weight per day; S is expressed as the concentration of pollutant, ppm by weight (either in terms of grams of liver or milligrams of hepatic microsomal protein) (from Walker 1987). © 2009 by Taylor & Francis Group, LLC Distribution and Effects of Chemicals in Communities and Ecosystems 81 More rapid elimination was needed than could be provided by passive diffusion in order to prevent tissue concentrations reaching toxic levels. Some models for predicting bioconcentration and biomagnication are presented in Box 4.1. 4.3 FATE OF POLLUTANTS IN SOILS AND SEDIMENTS Regarding soils, a central issue is the persistence and movement of pesticides that are widely used in agriculture. Many different insecticides, fungicides, herbicides, and molluscicides are applied to agricultural soils, and there is concern not only about effects that they may have on nontarget species residing in soil, but also on the pos- sibility of the chemicals nding their way into adjacent water courses. Soils are complex associations between living organisms and mineral particles. Decomposition of organic residues by soil microorganisms generates complex organic polymers (“humic substances” or simply “soil organic matter”) that bind together mineral particles to form aggregates that give the soil its structure. Soil organic matter and clay minerals constitute the colloidal fraction of soil; because of their small size, they present a large surface area in relation to their volume. Consequently, they have a large capacity to adsorb the organic pollutants that con- taminate soil. Within a freely draining soil there are air channels and soil water, the latter being closely associated with solid surfaces. Depending on their physical properties, organic compounds become differentially distributed between the three phases of the soil, soil water, and soil air. Hydrophobic compounds of high K ow become very strongly adsorbed to soil col- loids (Chapter 3, Section 3.1), and consequently tend to be immobile and persistent. OC insecticides such as DDT and dieldrin are good examples of hydrophobic com- pounds of rather low vapor pressure that have long half-lives, sometimes running into years, in temperate soils (Chapter 5). Because of their low water solubility and their refractory nature, the main mechanism of loss from most soils is by volatiliza- tion. Metabolism is limited by two factors: (1) being tightly bound, they are not freely available to enzymes of soil organisms, which can degrade them, and (2) they are, at best, only slowly metabolized by enzyme systems. Because of strong adsorption and low water solubility, there is little tendency for them to be leached down the soil pro- le by percolating water. The degree of adsorption, and consequently the persistence and mobility, is also dependent on soil type. Heavy soils, high in organic matter and/ or clay, adsorb hydrophobic compounds more strongly than light sandy soils, which are low in organic matter. Strongly lipophilic compounds are most persistent in heavy soils. When OC insecticides are rst incorporated into soil, they are lost relatively rapidly, mainly due to volatilization, before they become extensively adsorbed to soil colloids (Figure 4.3). With time, however, most residual OC insecticide becomes adsorbed, and subsequently there is a period of very slow exponential loss. In marked contrast to hydrophobic compounds, more polar ones tend to be less adsorbed and to reach relatively high concentrations in soil water. Phenoxyalkanoic acids such as 2,4-D and MCPA are good examples (Figure 4.3). Their half-lives in soil are measured in weeks rather than years, and they are more mobile than OC insec- ticides in soils. When rst applied they are lost only slowly. After a lag period of a © 2009 by Taylor & Francis Group, LLC 82 Organic Pollutants: An Ecotoxicological Perspective, Second Edition few days, however, they disappear very rapidly as a consequence of metabolism by soil microorganisms. This has been explained on the grounds that it takes time for a buildup in numbers of strains of microorganisms that can metabolize them; these microorganisms use the herbicides as an energy source. It has also been suggested that the lag period relates to the time it takes for enzyme induction to occur. Whatever the explanation, soils treated with these compounds stay enriched for a period, and further additions of the original compounds will be followed by rapid metabolism without a lag phase. If, however, the soils are untreated for a long period, they will revert to their original state and not show any enhanced capacity for degrading the herbicides. An important difference from the OC insecticides and related hydrophobic pollutants is that, because of their polarity and water solubility, they are freely available to the microorganisms that can degrade them. Interestingly, the phenoxyalkanoic acid 2,4,5-T is more persistent than either 2,4-D or MCPA. With three substituted chlorines in its phenyl ring, it is metabolized less rapidly than the other two compounds, and it would appear that metabolism is a rate-limiting factor determining rate of loss from soil. It was long assumed that there is little tendency for most pesticides or other organic pollutants to move through soil into drainage water. Indeed, this is to be expected with intact soil proles. Hydrophobic compounds will be held back by adsorption, whereas water soluble ones will be degraded by soil organisms. Some soils, how- ever, depart from this simple model. Soils high in clay can crack and develop deep ssures during dry weather. If rain then follows, pesticides, in solution or adsorbed to mobile colloids, can be washed down through the ssures, to appear in neighbor- ing drainage ditches and streams. This was found to happen with pesticides such as carbofuran, isoproturon, and chlorpyrifos in the Rosemaund experiment conducted in England during the period 1987–1993 (Williams et al. 1996). 0 0 10203040 506070 8090100 Days after Application Time Following Application (b) (a) 20 Herbicide Concentration (ppm) Log Concentration in Soil 40 60 80 100 120 Lag period Lag period 2,4-D MCPA Period of rapid loss during application and cultivation and for a time afterwards Concentration that would have been found if all applied material were retained by soil Period of slo w exponential loss FIGURE 4.3 Loss of pesticides from soil. (a) Breakdown of herbicides in soil. (b) Dis- appearance of persistent organochlorine insecticides from soils (from Walker et al. 2000). © 2009 by Taylor & Francis Group, LLC Distribution and Effects of Chemicals in Communities and Ecosystems 83 The inuence of polarity on movement of chemicals down through the soil prole has been exploited in the selective control of weeds using soil herbicides (Hassall 1990). In general, the more polar and water soluble the herbicide, the further it will be taken down into the soil by percolating water. Insoluble herbicides such as the triazine compound simazine (water solubility, 3.5 ppm), remain in the rst few cen- timeters of soil when applied to the surface. More water-soluble compounds such as the urea herbicides diuron and monuron (water solubilities 42 ppm and 230 ppm, respectively) are more mobile, and can move farther down the soil prole. Selective weed control can be achieved in some deep-rooting crops by judicious selection from this range of herbicides, so that the herbicide will only percolate far enough down the soil prole to control surface rooting weeds without reaching the main part of the root system of the crop (depth selection). Thus, when applied to the soil surface, simazine should only be toxic to shallow rooting weeds and should not affect crops that root farther down. Other more water-soluble herbicides can give weed control to greater depths in situations where the rooting systems of the crops are sufciently deep. When attempting depth selection in weed control, account needs to be taken of soil type. Herbicides will move farther down the prole in the case of light sandy soils than they will in heavy clays or organic soils. Although the major concern about the fate of organic pollutants in soil has been about pesticides in agricultural soils, other scenarios are also important. The dis- posal of wastes on land (e.g., at landll sites) has raised questions about movement of pollutants contained in them into the air or neighboring rivers or water courses. The presence of polychlorinated biphenyls (PCBs) or PAHs in such wastes can be a signicant source of pollution. Likewise, the disposal of some industrial wastes in landll sites (e.g., by the chemical industry) raises questions about movement into air or water and needs to be carefully controlled and monitored. In certain respects, sediments resemble soils. Sediments also represent an asso- ciation between mineral particles, organic matter, and resident organisms. The main difference is that they are situated underwater and are, in varying degrees, anaerobic. The oxygen level inuences the type of organisms and the nature of biotransformations that occur in sediments. A feature with sediments, as with soils, is the limited availability of chemicals that are strongly adsorbed. Again, compounds with high K ow tend to be strongly adsorbed, relatively unavailable, and highly persistent. There is much interest in the question of sediment toxicity and the availability to bottom-dwelling organisms of compounds adsorbed by sedi- ments (Hill et al. 1993). One case in point is pyrethroid insecticides (see Chapter 12), which are strongly retained in sediments on account of their high K ow values. Because of their ready biodegradability, they are not usually biomagnied in the higher trophic levels of aquatic food chains. However, they are available to bottom-dwelling organisms low in the food chain. Questions are asked about the possible long-term buildup of pyrethroids in sedi- ments and their effects on organisms in lower trophic levels. © 2009 by Taylor & Francis Group, LLC 84 Organic Pollutants: An Ecotoxicological Perspective, Second Edition 4.4 EFFECTS OF CHEMICALS UPON INDIVIDUALS— THE BIOMARKER APPROACH Until now this narrative has been concerned with questions about the movement and distribution of chemicals in the living environment, a topic that relates to the eld of toxicokinetics in classical toxicology, although on a much larger scale. It is now time to move to a consideration of the effects that chemicals may have upon living organ- isms, which relates to the area of toxicodynamics in classical toxicology. Effects upon individuals will be discussed before dealing with consequent changes at the higher levels of biological organization—population, community, and ecosystem. Measuring effects of chemicals upon free-living individuals in the natural envi- ronment is not an easy matter. Mobile animals need to be captured so that samples of tissues can be taken for analysis, but this is often difcult to do in a properly controlled way in the eld. It is easier to obtain samples from sedentary species (e.g., mollusks or plants) or of eggs in the case of birds, reptiles, and some invertebrates. All too often sampling is destructive, which raises problems of experimental design and statistical evaluation of results. In principle, measuring behavioral effects of chemicals is an attractive option, but this can be hard to achieve in practice because of the difculty of making reliable measurements in the eld. The problems of sam- pling can, to some extent, be circumvented by deploying indicator species that have been maintained in a “clean” environment into the eld. Thus, control sh from the laboratory can be held in cages in contaminated waters and samples taken from them after periods of exposure to pollutants. Uncontaminated birds’ eggs can be intro- duced into the nests of birds of the same species that are breeding in a polluted area. In this way, changes caused by pollutants can be measured and evaluated. Problems of sampling aside, the success of any strategy of this kind depends on the availability of reliable tests that can measure harmful effects of chemicals under eld conditions. Reference has already been made to biomarkers (Chapter 2, Box 2.2). In the following account they are dened as “biological responses to environ- mental chemicals at the individual level or below, which demonstrate a departure from normal status.” This denition includes biochemical, physiological, histologi- cal, morphological, and behavioral changes, but does not extend to changes at higher levels of organization. Changes at population, community, or ecosystem level are regarded instead as bioindicators. The concept of biomarkers is illustrated in Figure 4.4. As the dose of a chemical increases, the organism moves from a state of homeostasis to a state of stress. With fur- ther increases in dose, the organism enters rst the state of reversible disease, and even- tually the state of irreversible disease, which will lead to death. In concept, all of these stages can be monitored by biomarker assays (lower part of conceptual diagram). Some biomarker responses provide evidence only of exposure and do not give any reliable measure of toxic effect. Other biomarkers, however, provide a measure of toxic effects, and these will be referred to as mechanistic biomarkers. Ideally, bio- marker assays of this latter type monitor the primary interaction between a chemical and its site of action. However, other biomarkers operating “down stream” from the original toxic lesion also provide a measure of toxic action (see Figure 14.3 in Chapter 14), as, for instance, in the case of changes in the transmission of action potential © 2009 by Taylor & Francis Group, LLC [...]... cytochrome P450 and an insensitive form of the sodium channel (Table 4. 3 and McCaffery 1998) Apart from the resistance of insects to insecticides, resistance has been developed by plants to herbicides, fungi to fungicides, and rodents to rodenticides Rodenticide resistance is discussed in Chapter 11, Section 11.2.5 © 2009 by Taylor & Francis Group, LLC 96 4. 7 Organic Pollutants: An Ecotoxicological Perspective, ... toxic effects are manifest) Mechanistic biomarkers (see Table 4. 2 for examples) have potential for measuring adverse effects of chemicals in the field—effects that may be translated into changes at the population level and above © 2009 by Taylor & Francis Group, LLC 86 Organic Pollutants: An Ecotoxicological Perspective, Second Edition TABLE 4. 2 Some Mechanistic Biomarker Assays Chapter in This Text... in Chapter 17, after consideration of the individual examples given in Part 2 © 2009 by Taylor & Francis Group, LLC 98 4. 9 Organic Pollutants: An Ecotoxicological Perspective, Second Edition SUMMARY The movement of organic pollutants along food chains, and their fate in soils and sediments, is dependent upon biological as well as chemical factors The chemical and biochemical properties of pollutants... environment, and these are described in the chapters that © 2009 by Taylor & Francis Group, LLC 90 Organic Pollutants: An Ecotoxicological Perspective, Second Edition follow These have included cases where the recovery of populations followed the reduction of pollutant levels in the environment Examples include the declines of predatory birds caused by cyclodiene insecticides and p,p -DDE, and the decline... organisms since very early in evolutionary history There is abundant evidence of compounds produced by plants and animals that are toxic to species other than their own and which are used as chemical warfare agents (Chapter 1) Also, as we have seen, wild animals can develop resistance mechanisms to the toxic compounds produced by plants In Australia, for example, some marsupials have developed resistance... continuing use of pesticides, problems of resistance began to emerge The emergence of strains of pest species possessing genes that confer resistance was an inevitable consequence of the © 2009 by Taylor & Francis Group, LLC 94 Organic Pollutants: An Ecotoxicological Perspective, Second Edition continuing selection pressure of the pesticides Indeed, this can be seen to mirror the development of defense... in the liver Certain metabolites of coplanar PCBs such as 4- OH, 3,3 ,4, 4 -TCB can compete with thyroxine (T4) for binding sites on the protein transthyretin in blood This interaction leads to the breaking apart of © 2009 by Taylor & Francis Group, LLC Distribution and Effects of Chemicals in Communities and Ecosystems 87 the protein and the loss of thyroxine and retinol from the blood Here, measurement... by organophosphorous insecticides, and eggshell thinning of some predatory birds caused by p,p -DDE (Table 4. 2) Sensitivity is another desirable characteristic that can facilitate the early detection of sublethal effects The detection of later effects seen during the terminal stages of poisoning is of less value © 2009 by Taylor & Francis Group, LLC 88 Organic Pollutants: An Ecotoxicological Perspective, ... associated with changes in toxicokinetics are predominately cases of enhanced metabolic detoxication With readily biodegradable insecticides such as pyrethroids and carbamates, enhanced detoxication by P450-based monooxygenase is a common resistance mechanism (see Table 4. 3) TABLE 4. 3 Examples of Resistance of Insects to Insecticides Insecticide Species Strain (RF) Mechanism Comment Cypermethrin (cis-isomers)... studies on fish, rodents, and birds in the field and/or in the laboratory give evidence for a range of sublethal neurotoxic and behavioral effects of OP insecticides when brain acetylcholinesterase inhibition is in the range 40 –50%—before the appearance of severe toxic manifestations and death (see Chapters 10 and 16) A few OP compounds cause delayed neuropathy in mammals and birds, and this has been related . Francis Group, LLC 76 Organic Pollutants: An Ecotoxicological Perspective, Second Edition entire food chain. The best studied examples of this are the organochlorines (OCs) dieldrin and p,pb-DDE. Taylor & Francis Group, LLC 78 Organic Pollutants: An Ecotoxicological Perspective, Second Edition BOX 4. 2 MODELS FOR BIOCONCENTRATION AND BIOACCUMULATION As indicated in Table 4. 1, aquatic. in lower trophic levels. © 2009 by Taylor & Francis Group, LLC 84 Organic Pollutants: An Ecotoxicological Perspective, Second Edition 4. 4 EFFECTS OF CHEMICALS UPON INDIVIDUALS— THE BIOMARKER