Underlying issues in bioaccessibility and bioavailability:

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Ecotoxicology and Environmental Safety 56 (2003) 52–62

Underlying issues in bioaccessibility and bioavailability: experimental methods K. Hund-Rinke and W. Ko¨rdel Fraunhofer Institute for Molecular Biology and Applied Ecology, P.O. Box 1260, 57377 Schmallenberg, Germany Received 19 March 2003

Abstract This article presents experimental designs focusing on assessing of the bioavailability of metals in aquatic organisms, soil organisms, microorganisms, plants, birds, and mammals. Standardized test systems receive the greatest emphasis. With regard to microorganisms, animals, and plants, the study concentrates on toxicity as an indicator for bioavailability. In respective test procedures, results are usually calculated for total chemical concentrations; chemical analyses are not commonly in routine assessments. For soil organisms chiefly exposed by the water pathway, the bioavailable fraction of contaminants can be roughly determined by chemical analyses in aqueous soil extracts simulating soil pore water concentrations. Human toxicity can be determined using adequate in vitro test designs. In addition to experimental designs, results from the literature dealing with specific problems of bioavailability are presented. r 2003 Elsevier Inc. All rights reserved. Keywords: Metal; Bioavailability; Experimental methods; Aquatic organisms; Soil organisms; In vitro tests; Metal speciation

1. Introduction In the context of monitoring compliance with water or soil quality standards and for investigating potential adverse effects of metals on aquatic and terrestrial ecosystems, including mammals, a large set of standardized test methods is available. Assessments of potential impacts of specific compounds on terrestrial or aquatic habitats, however, require estimates considering the different characteristics of a specific site, unless there are site-relevant data available. Quality assessments of the terrestrial compartment for example, have to deal with the problems of comparability of test results elaborated with different soils (soils of different physicochemical properties and background concentrations, or artificial soils), as well as the generalization of effects elaborated with a few specifically selected soils to be transferred to the variety of soils of a certain region. 

Corresponding author. Fax: +49-2972-302-319. E-mail addresses: [email protected] (K. Hund-Rinke), [email protected] (W. Ko¨rdel). 0147-6513/03/$ - see front matter r 2003 Elsevier Inc. All rights reserved. doi:10.1016/S0147-6513(03)00050-2

Bioavailability of metals in soil, sediment, and food is affected by many conditions, for example: *

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Sorption and binding to constituents of soil and sediment matrix, such as clays, organic matter and oxides. Soil and sediment conditions, such as redox conditions, pH value, moisture content. Actual composition of the soil and sediment pore water, including pH value, dissolved organic matter content, complexing agents, composition of interfering anions and cations. Sequestration and binding in plants and other foodstuffs. Species-dependent regulation mechanisms for uptake, excretion, and storage (see, for example, Peijnenburg, in this issue). Metal speciation and bioavailability (see Peijnenburg, 2003, this issue). Uptake route and specific habitats of test species: Organisms living in the pores of soils and sediment are mainly exposed via the pore water, whereas organisms living in the bulk soil, in air-filled soil pores, or on the soil surface are mainly exposed by means of food and direct contact with the pollutant.

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The interactions of the parameters influencing the bioavailability of metals and their speciation cannot be modeled with sufficient precision (Peijnenburg et al., 1999a, b). However, as a first approximation, for organisms mainly exposed to the water phase, the concentrations of the pollutants in the pore water of soils and sediments are considered the bioavailable portions. The metal concentration is calculated using sorption coefficients (equilibrium partitioning method, EPM) (Crommentuijn et al., 1997; Pederson et al., 1994). This brief description shows that bioavailability is a complex term dealing with species-specific and dynamic processes. It is strongly associated with the species considered and with the type of exposure, including metal speciation. The complexity of parameters and interactions is not always sufficiently considered in toxicity testing. The following sections present several test systems as well as results from the literature for aquatic organisms, terrestrial plants, soil organisms, microorganisms, birds, and mammals. Additionally, in vitro test designs for obtaining information on accessibility and bioavailability for humans are described.

2. Aquatic organisms Aquatic single-species laboratory tests are routinely used to assess possible effects of chemicals on the aquatic environment. Standardized international guidelines exist for fish, algae, daphnids, and plants. Included are acute tests (e.g., OECD, 1984d, f, 1992a), chronic tests (e.g., OECD, 1984a, d, 1998a, 2000c), and tests with sensitive live stages of fish (e.g., OECD, 1992b, 1998b, 2000b). Furthermore, a specific bioaccumulation test for fish was developed (OECD, 1996). Further to testing pollutant effects on organisms in the water phase, tests with chironomids were elaborated to consider sediment organisms as well (OECD, 2001a, b). All tests are conducted under standardized conditions. However, the toxicity of heavy metals can vary significantly depending on the applied conditions, for example, the temperature used in the test (Spurgeon et al., 1997). For testing metals, either the total or the dissolved metal concentrations are usually determined. These approaches, however, do not reflect the realistic exposure of aquatic organisms and the pollutant effects (Morel, 1983; Pagenkopf, 1983; Playle et al., 1993). The importance of considering bioavailability for the development of water and sediment quality criteria for metals has been described by several authors (Di Toro et al., 1991; Ankley et al., 1996; Allen and Hansen, 1996). Research has led to an improved understanding of the influence of site-specific water chemistry in aquatic

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systems on the bioavailability of metals and the prediction of toxic effect levels, resulting, for example, in the biotic ligand models (BLM; Di Toro et al., 2001b; Santore et al., 2001). The extrapolation of simple laboratory ecotoxicity data to realistic field responses must be considered critically. Simple model calculations may under- or overestimate impacts of metals on freshwater ecosystems. The derivation of environmental quality standards for metals, for example, water-quality criteria (WQC), or predicted no-effect concentrations (PNECs) is based on data obtained from standard laboratory tests by adding assessment factors to the lowest laboratory toxicity values or by using statistical extrapolation models (EU, 1996). The deficiencies of these standard procedures are well known, and there is a serious need for an in-depth consideration of biovailability for heavy-metal testing and assessment, including the following considerations. 2.1. Bioavailability considerations 2.1.1. Application of metals to the test systems The frequent use of soluble metal salts simulates worst-case scenarios, because their bioavailability is higher than metal availability in the real environment. A realistic simulation of exposure should consider the complex chemical reactions of dissolution, binding, and complexation between metal species and the constituents of the environmental aquatic phase (Di Toro et al., 2001a). 2.1.2. Regional specific background concentrations in various aquatic ecosystems Higher background concentrations may act as triggers for high biodiversity during evolution and changes in sensitivity of the local communities in the specific aquatic ecosystem. Consequently, concentrations that are toxic to one community may be essential for another. 2.1.3. The need for more realistic simulation of the exposure route of metals There have been attempts made to improve the realism of single-species test procedures in aquatic testing by introducing a sediment layer in the test vessels, so as to observe distribution and sorption processes between the water and sediment phases. By using site-specific samples, the complex water chemistry that determines bioavailability is considered at least partly, because changes of the natural state through sampling and laboratory testing are reduced. However, the complexity of the test parameters lowers comparability. Compared with single-species tests there is greater ecological relevance in micro- and mesocosm testing,

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including water, sediment, and a complex biocenosis. However, there is a concomitant increase in complexity, costs, and the degree of expertise needed. These integrative test systems can significantly alter the bioavailability of the test substances. For example, metals applied in the water phase can bind to the sediment resulting in a decreased bioavailability for organisms living in the water phase. Complexation of the metals by natural complexing agents (e.g., humic acids) will also alter bioavailability, as will changes in pH and redox values due to biological activities. Furthermore, the test systems can consider fluctuations within and interactions between populations. Therefore, micro- and mesocosm studies are commonly used as higher tier tests in pesticide registration. 2.2. General considerations The briefly discussed approaches are, for the most part, relevant for the terrestrial compartment as well. They show that a test system must be designed for each specific question considering the site-specific conditions. Also there should be further investigation of the effects of essential metal compounds compared to non-essential metals, and the transferability of results from laboratory experiments to outdoor conditions and ecosystems with low metal background concentrations to which organisms may be adapted. Moreover, for each test species, the specific uptake and excretion (regulation mechanisms) for essential metal species have to be considered in testing, risk assessment, and derivation of WQC (Slijkerman et al., 2000; Janssen et al., 2000).

3. Terrestrial plants National as well as international guidelines exist or are in preparation for testing side effects of chemicals on plants. These refer to acute tests considering germination, plant growth, or growth of roots (ISO, 1993a, 1995; OECD, 1984g). In addition, chronic tests are under development (ISO, in preparation). Quantitative as well as phenotypic parameters (e.g., chlorosis and necrosis) are considered. These, however, do not reflect effects on metabolism that are already occurring at much lower metal concentrations. Another parameter that seems to be very sensitive to heavy-metal contamination is the extent of nodulation in the legume root nodule symbiosis (Neumann et al., 1998). The guidelines propose cultivated plants as test species. Presently, two principal approaches are followed for risk assessment. Either only two species (one monocotyledonous and one dicotyledonous) are tested to achieve comparability, or testing of a wide variety of plants is requested, including regional specific wild plants, to cover the variability of the flora. In Canada,

for example as many as 30 species (representing 10 families) are proposed to be tested for the registration of pesticides (Boutin et al., 1995). The tests described in the guidelines are usually conducted in small pots with a ratio between soil and roots that differs significantly from field conditions. As a result, the plants may take up larger amount of test substance than they would under the field conditions. All the guidelines referring to effects on terrestrial plants do not require chemical analyses. Results are compared to nominal concentrations. No specific international guidelines exist for accumulation studies. Nevertheless, heavy metals have been well investigated because of their high ecotoxicological and human toxicological relevance, and chemical determinations are easily done. For example, the accumulation of Zn and Cd depends on the pH of the soil, the clay content and the accumulating plant species. Table 1 compares the capacity of various plants to take up Cd, Pb, and Tl. Terrestrial plants have evolved a complex suite of strategies aimed at increasing the bioavailability of essential trace elements (micronutrients) in the rhizosphere and transporting these metals into the root so as to ensure an adequate supply (Marschner, 1995). However, at higher concentrations, many of the essential micronutrients (such as Zn, Cu, Ni) have the potential for phytotoxicity; alternately, phytotoxic heavy-metal analogues are taken up in addition to micronutrients (e.g., Cd, Pb). Therefore, their uptake is subject to physiological regulating and tolerance mechanisms. Two main types of response are avoidance and sequestration (Kochian, 1999). Such plant, metal, and soil-specific processes must be considered in effects testing a complicating aspect in elaborating a general test strategy and the following assessment. Considerable research has concerned on the uptake of heavy metals by cereals. The results show that in the concentration range between ‘‘sufficiently available as micronutrients’’ and ‘‘not yet phytotoxic’’, the main uptake route of some metals (e.g., As, Cd, Cu, Ni, Pb, Zn) is by pore water, and the bioavailable fraction can be simulated by extraction with NH4NO3 or CaCl2 (Delschen and Ru¨ck, 1997; Welp et al., 1999; Davidson et al., 1998). In addition to the impact on metals bioavailability of soil characteristics and plant species, anthropogenic forces can have a significant influence. Simultaneous application of several heavy metals, for example by the addition of contaminated sewage sludge, can yield results that differ substantially from the application of single elements (Parkpain et al., 2000). Irrigation with saline waters, a common practice in developing countries, can also significantly alter the uptake rate (Helal et al., 1999). Examples of phytoremediation measures show that chemical mobilization of heavy metals by chelators increases their concentration in soil pore

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Table 1 Ability of several plant species to enrich heavy metalsa Ability for enrichment

Cd

Pb

Tl

High

Endive, Lollo rosso, beet, celeriac, spinach, wheat, leaves of sugar beet Cauliflower, broccoli, chinese cabbage, kale, oat, carrot, leek, beet root, cabbage lettuce, black salsify

Endive, Lollo rosso

Kale, rape

Cauliflower, broccoli, chinese cabbage, beet, celeriac, spinach, wheat, kale, oat, carrot, leek, beet root, cabbage lettuce, black salsify, radish, rye, brussels sprouts, red cabbage, white cabbage, savoy, onion, lamb’s lattuce, barley, turnip, cabbage Bush bean, pea, cucumber, cabbage lettuce, potato, runner bean, tomato, zucchini

Celeriac, spinach, beet, broccoli, beet root, black salsify, carrot, leek, lettuce, radish, savoy

Medium

Low

a

Bush bean, pea, field salad, Cucumber, potato, turnip cabbage, pumpkin, paprika, radish, rye, brussels sprouts, red cabbage, runner bean, tomato, white cabbage, savoy, zucchini, onion

Cauliflower, endive, Chinese cabbage, bush bean, pea, cucumber, turnip, cabbage, pumpkin, paprika, red cabbage, runner bean, tomato, white cabbage, zucchini, onion

Asterisks indicate uncertainty. Source of data, LABO (1997).

Phytophagous Nematodes

Collembolans

Predaceous Mites

Cryptostigmatid Mites

Roots

Predaceous Collembolans

Mycorrhiza

Saprophytic Fungi

Noncryptostigmatid Mites Fungivorous Nematodes

Detritus

Nematode Feeding Mites

Predaceous Nematodes

Earthworms

Enchytraeids

Omnivorous Nematodes

Bacteriophagous Nematodes Amoebae Bacteria

Flagellates

Bacteriophagous Mites

Fig. 1. Diagram of a soil food web (based on De Ruiter et al., 1993, 1997).

water. The uptake of these metals by plants is, however, far less increased, because complexation significantly decreases bioavailability (Felix et al., 1999).

4. Soil organisms To obtain information on potential effects of heavy metals on soil organisms and consequently on the

habitat function of soils, it is necessary to conduct investigations with representative organisms. In this context the complexity of the biocoenosis has to be taken into account. Fig. 1 gives an example of such complex interactions. Depending mainly on their size, organisms colonize different ‘‘regions’’ in soils. A classification of soil organisms and their habitats is presented in Fig. 2. Several standardized guidelines exist for testing chemicals with respect to their toxicity on soil

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Central pores

Macropores, cracks, worm-/root channels 0.2 mm

Micro fauna

20 mm

2 mm Meso fauna

Macro fauna

Mega fauna

Bacteria (au,he) Fungi (he) Algae (ph) Protozoa (mi) Nematodes (mi,ma) Enchytraeids (ma,mi)

Earthworms (ma) Mites (mi,zo) Spiders (zo) Millipedes (zo,ma) Collembolans (mi) Beetles and their larvae (zo,ma) Diptera larvae (ma) Gastropods (ma)

Mammals (zo,ma)

Fig. 2. Classification of common soil organisms based on body length, assignment to trophic groups (au=autotrophic, he=heterotrophic, ph=phototrophic, ma=macrosapro- and macrophytophagous; mit=microphytophagous, zo=zoophagous incl. necrophagous) and indication of colonized pore types (based on Gisi (1990) and Ro¨mbke and Schick, 2000).

Table 2 International standardized test guidelines to determine the toxicity of chemicals on soil organisms (laboratory tests) Organism

Measuring parameter

Guideline

Microorganisms (microflora)

Nitrification, N-mineralization (incubation: 28 days)

OECD (2000e) ISO (1997a) ISO (2001a)a ISO (2001b)a OECD (2000d) ISO (1997c) ISO (2001b)a ISO (1997b) OECD (2000a) ISO (2000)a ISO (1999)a OECD (1984e) ISO (1993b)a ISO (1998)a

Nitrification Respiration Substrate-induced respiration

Enchytraeids: Enchytraeus albidus (mesofauna) Collembolans: Folsomia candida (mesofauna) Earthworm: Eisenia fetida (macrofauna)

Biomass (fumigation-extraction) Reproduction Reproduction Mortality Reproduction

a

These tests were validated for the assessment of soils in a ring test.

organisms (Table 2). The selected test organisms represent different trophic levels, habitats, and exposure routes.

The suitability of some of these test guidelines for soil assessments was validated in a round-robin test (HundRinke et al., 2002). All test systems proved to be suitable

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for assessing the toxicity of heavy metals on soil organisms. Influencing the results obtained are soil characteristics, the form in which the test chemical is added, environmental factors, and especially the organism itself. The concentration of a chemical in a test organism is not usually analyzed after the exposure to determine the bioavailability of the toxicant. A prerequisite for determining the bioavailable portion of a contaminant is the complete elimination of adhering soil particles (on the surface of the organisms or in the intestine). The concentrations in the organisms that cause an observed effect can significantly differ from the concentrations determined in soil. Mechanisms that regulate the uptake of contaminants have been reported for some organisms. Such mechanisms depend on the heavy metal as well as on the organism itself (Peijnenburg et al., 1999a, b; Vijver, 2000). In this context the application of (radio) labeled isotopes may improve the prediction of uptake rate constants (Peijnenburg et al., 1999b). Bioavailability and toxicity are substantially influenced by the testing conditions. Spurgeon et al. (1997) showed that temperature significantly influenced the test results for Zn. Spurgeon and Hopkin (1995) also studied the influence of soil characteristics on the toxicity of heavy metals (Cd, Cu, Pb, Zn) by using an artificial soil as well as contaminated soils. In these studies the most toxic element was Zn, which was at least an order of magnitude more toxic to Eisenia fetida in artificial soil than in contaminated soil. Differences in the availability were explained by the soil characteristics (Corg, clay) and by the form in which the test chemical was added (soluble nitrate salt and anthropogenic contamination). To increase the ecological relevance, additional singlespecies investigations in terrestrial model ecosystems (TME) have been proposed (Kula and Ro¨mbke, 1998).

5. Microorganisms Studies on the bioavailability of metals for microorganisms often consider only toxicity. There exist several standardized guidelines for the determination of microbial activities, and numerous methods are described in the literature (e.g., Schinner et al., 1996). For example, microbial respiration (OECD, 2000d; ISO, 1997c, 2001b), enzymatic activities (OECD, 2000e; ISO, 1997a, 2001a), or colony-forming units are determined. Furthermore, the assessment of phospholipids as an indicator for the structure of the microbial biocenosis is applied (Frostegard et al., 1993, Zelles et al., 1992). Molecular methods give information on the activity and structure of the biocenosis (e.g., Manz et al., 1992; Liu et al., 1997; Fleming et al., 1998; Noda et al., 1999; Fritze et al., 2001). The effects found are related to the total concentration of heavy metals or to the watersoluble concentration. However, because bacteria are

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present within colonies in soil (Harris, 1994) or are protected by clays (Ladd et al., 1995), they often may not be exposed to the equilibrium solution activity of heavy metals. A review about toxicity of heavy metals to microorganisms and microbial processes in agricultural soils is given by Giller et al. (1998). Besides the investigation of toxic effects and changes in the structure of the biocenoses, uptake mechanisms are studied using molecular methods to obtain information on the mechanism of metal bioavailability. To understand metal homeostatic mechanisms, model organisms such as Saccharomyces cerevisiae and Escherichia coli have been studied comprehensively. The advantage of using these organisms is the relative ease with which molecular genetic techniques can be done and the availability of the complete genomic sequence. For details see the article ‘‘Underlying issues including approaches and information needs in risk assessment: Chapter III.3—Processes at the cellular level’’ (Baker et al., 2003, elsewhere in this issue). Furthermore, microorganisms are used as biosensors to detect bioavailable metal contaminants in the environment. They can complement other analytical methods by distinguishing the bioavailable fraction from the total amount of contaminant present. Numerous bioreporter systems have been constructed. For details see article by Rensing and Maier (2003, this issue).

6. Birds To evaluate the toxicity of chemicals to avians there exist two standardized guidelines, a short-term toxicity study (OECD, 1984b; Observation: Mortality, Body Weight, Food Consumption, Signs of Intoxication) and a reproduction study (OECD, 1984c). The test substance is applied in the diet. In research projects including monitoring some further endpoints and species of birds are studied. Blood concentrations are used as a function of current exposure and mobilization from internal tissues. Concentrations of metals in feathers can be used as a marker of exposure during the 2–4 weeks of formation. Burger and Gochfeld (2000) review laboratory and field studies with respect to the effects of lead on birds. Birds are useful models in behavioral toxicology because, like humans, they rely on visual and vocal (rather than olfactory and tactile) cues for communication, and because their young ones have an extended developmental period during which they are dependent upon their parents for protection and food.

7. Mammals Studies with mammals are conducted mostly to obtain information on human toxicity of a contaminant.

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Several standardized test designs exist. For assessing the availability and accessibility of metals, oral toxicity studies ranging from acute toxicity to long-term investigations (e.g., OECD, 1995, 1998c, d) are of value. Also studies of two (or more) generations can be done (e.g., OECD, 1983, 2001c). Among rodents, the preferred species is the rat, although other rodent species, such as mice and guinea pigs, may serve as test organisms. Although dogs are the most commonly used non-rodent species, other species such as pigs or minipigs, may also be used. Either gavage or diet are proposed methods for feeding. Toxic effects are determined by several parameters, including mortality, clinical observations, ophthalmological and functional examinations, body weight change, and food and water consumption. Researchers determine concentrations of chemicals, in hair, bones, or internal organs (Hac and Krechniak, 1996). If necessary for the purpose of the study, the test substance is dissolved or suspended in a suitable vehicle usually increasing bioavailability. Concerning particlebound contaminants it seems that contaminants ingested with soil are associated with lower absorption and toxicity than are contaminants ingested with food or liquids (Freeman et al., 1992, 1994, 1996; Casteel et al., 1997). Furthermore, it is well known that the diet itself significantly influence the absorbability of substances. For example, the absorption of Zn (Likuski and Forbes, 1965), Fe, Cu, Mg (Davies and Nightingale, 1975), and Cd (Turecki et al., 1994) is reduced by phytic acids, the major phosphorous storage constituents of most cereals, legumes and oilseeds (Reddy et al., 1982). Several in vivo investigations conducted with humans showed a decreasing uptake of iron in the presence of whole-wheat bread (Kobza and Steenblock, 1977; Elwood et al., 1968). Relative intestinal absorption of iron decreases in the presence of the following components of a meal (Rossander et al., 1979) as in coffee/ chocolate4milk4tea. In studies with mammals, particle-bound metals can be mixed and administered with the diet. However, simulating special diets for humans will be difficult. The guidelines require only biological data, and no chemical data of absorbed chemicals. Therefore, the effects can only be related to nominal concentrations in food or drinking water. Concentrations of the bioavailable portions of contaminants usually remain unknown.

8. In vitro test designs for the evaluation of particlebound metals with respect to humans There are several in vivo approaches available to obtain information on the toxicity of metals for humans (e.g., from soil or food). However, each has its limitations. Human balance studies are susceptible to

considerable error (Isaksson and Sjogren, 1967), and in vivo animal studies are problematic with respect to extrapolation to humans because of potential discrepancies due to differences in the biochemical and physiology of the gastrointestinal (GI) tract in humans and animals (Finch et al., 1978). Such problems led to the development of in vitro systems that aim to simulate various stages of digestion. Synthetic digestive fluids are used, because natural ones cannot be standardized to a sufficient degree. The Commission of the European Community proposes the application of synthetic saliva and synthetic gastric juice (EU, 1988). These fluids are commonly applied to investigate the resorption of essential food components, the mobilization of active ingredients of pharmaceutical products, and the mobilization of organic and inorganic contaminants (Ritschel, 1973; Rotard et al., 1993; Ruby et al., 1993, 1996; Wolters et al., 1993; Sheppard et al., 1995; Hack and Selenka, 1996). The elution models differ in kind and concentration of organic and inorganic components of the synthetic gastric juices, the duration of the elution procedure, the concentration of the contaminants or contaminated materials as well as the methodology applied for the separation of the mobilized contaminants. In addition to the composition of saliva, stomach, and small intestinal fluid, the pH must be considered. To study the influence of food using these simulation systems appropriate components can be added to the synthetic fluids, using such complex amendments as fullcream milk powder (E-DIN, 2000), or such defined amendments as acetate, citrate, lactate, and malate (Ruby et al., 1993, 1996). Full-cream milk powder is mainly used to simulate the mobilization of contaminants in the gut of infants, under the assumption that it qualifies as a standardized material. Thus far no systematic investigations on the effects of single components of the gastric juices have been done. Depending on the problem to be investigated, single compartments of the GI tracts are considered and modeled (e.g., Sheppard et al., 1995), or a process consisting of several stages is simulated (e.g., Ruby et al., 1993, 1996; Sheppard et al., 1995; Miller et al., 1981; Rotard et al., 1993; Hamel et al., 1999). A complex system is presented by Molly et al. (1993). Measurements of bioavailability may be restricted to the determination of the solubilized metal content after simple separation of fluids and particles, for example by centrifugation (E-DIN, 2000). The determined amount is considered to be the maximum amount available for intestinal adsorption (bioaccessible amount; see Peijnenburg, 2003, this issue). Furthermore, the process of resorption can be integrated in the experimental design. Simple physical methods, as for example online ultrafiltration (Wolters et al., 1993) or dialysis (Miller et al., 1981), are described. These procedures, however, do not consider active-transport mechanisms. To study the

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transport of the mobilized contaminants, across the intestinal wall and to obtain information on the bioavailable portion of the contaminants, biological experiments can be conducted, for example using in vitro-differentiated intestinal cells (caco2-cells) (Oomen, 2000). In comparison with the total content of contaminants, the models can be used to achieve a more realistic assessment of the toxic potential of particle-bound contaminants. However, it would be difficult to include in one experiment antagonistic effects or demobilization influences.

9. Summary Standardized experimental methods for investigating the bioavailability of chemicals and metals exist only for the determination of toxic effects to organisms. The guidelines are restricted to single-species tests and cover a spectrum of test organisms considering different trophic levels, habitats, and exposure routes. However, because chemical analyses usually are not integrated and effects are compared to nominal concentration, information on the bioavailable portion of the contaminants is not obtained. Particularly for soil organisms affected mainly via the water pathways, the water-soluble concentration is more appropriate for a comparison of chemical data and observed effects. Micro- and mesocosm studies that include a biocenosis increase the environmental relevance. Several studies described in the literature include chemical analyses. Some of these describe the total content of a contaminant in the organism (body burden) without differentiating between the available (effective) concentration and the detoxified fraction. For larger animals parts of the organisms are investigated (e.g., liver, hair, feather). In most of these investigations there is no differentiation between the available and detoxified fractions. Especially for the assessment of heavy-metal toxicity to humans, in vitro systems have been developed. However, specific resorption processes are not included in the guidelines. The concentration of the solubilized heavy metal is regarded as the toxic fraction in these systems.

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