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Contaminant bioavailability and bioaccessibilityPart 1: A scientific and technical review

technicalreport

14no.

CRC for Contamination Assessment and Remediation of the Environment

J.C. Ng, A.L. Juhasz, E. Smith and R. Naidu

CRC for Contamination Assessment and Remediation of the Environment

Technical Report no. 14

Contaminant bioavailability and bioaccessibility

Part 1: A scientific and technical review

J.C. Ng1,2, A.L. Juhasz2,3, E. Smith2,3 and R. Naidu2,3

1 National Research Centre for Environmental Toxicology, University of Queensland 2 CRC CARE Pty Ltd

3 Centre for Environmental Risk Assessment and Remediation (CERAR), University of South Australia

April 2010

Cooperative Research Centre for Contamination Assessment and Remediation of the Environment, Technical Report series, no. 14 April 2010 Copyright © CRC CARE Pty Ltd, 2010 This book is copyright. Except as permitted under the Australian Copyright Act 1968 (Commonwealth) and subsequent amendments, no part of this publication may be reproduced, stored or transmitted in any form or by any means, electronic or otherwise, without the specific written permission of the copyright owner. CRC CARE gives permission to NEPC to use extracts of this document for the purpose of the NEPM, with appropriate citation as indicated below. ISBN: 978-1-921431-20-3 Enquiries and additional copies: CRC CARE, PO Box 486, Salisbury South, South Australia, Australia 5106 Tel: (61) (08) 8302 5038 Fax: (61) (08) 8302 3124 www.crccare.com This report should be cited as: Ng, JC, Juhasz, AL, Smith, E & Naidu, R 2010, Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review, CRC CARE Technical Report no. 14, CRC for Contamination Assessment and Remediation of the Environment, Adelaide, Australia. Disclaimer: This publication is provided for the purpose of disseminating information relating to scientific and technical matters. Participating organisations of CRC CARE do not accept liability for any loss and/or damage, including financial loss, resulting from the reliance upon any information, advice or recommendations contained in this publication. The contents of this publication should not necessarily be taken to represent the views of the participating organisations. Acknowledgement: CRC CARE acknowledges the contribution made by Jack Ng (University of Queensland, CRC CARE), and Albert Juhasz, Euan Smith and Ravi Naidu (University of South Australia, CRC CARE) towards the writing and compilation of this technical report.

CRC CARE Technical Report no. 14 i Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review

Table of contents

1. NEPM bioavailability review 1

1.1 Scope 1

1.2 Specification 2

1.2.1 General 2

1.2.2 Specific outcomes 2

1.2.3 Presentation 2

2. Background 3

2.1 National and international approaches 3

2.1.1 Key findings 4

3. Bioavailability and bioaccessibility concepts 5

3.1 Bioavailability processes as defined by NRC (2003) 6

3.2 Human contaminant exposure 10

3.2.1 Ingestion and release of contaminants from the soil matrix 10

3.3 Methods for the determination of contaminant bioavailability 12

3.3.1 Assessment of contaminant bioavailability – organics 14

3.3.2 Assessment of contaminant bioavailability – inorganics 19

3.4 Factors influencing contaminant bioavailability 31

4. In vitro methods for assessing contaminant bioaccessibility 33

4.1 Chyme composition 39

4.2 Gastric and intestinal phase pH 40

4.3 Gastric and intestinal phase residence time 40

4.4 Soil:solution ratio 40

4.5 Amendments to gastric and intestinal phases 41

4.6 Other parameters 41

5. Bioaccessibility assessment for contaminated soils 42

5.1 Correlation between in vivo bioavailability and in vitro bioaccessibility 47

6. Knowledge gaps 52

7. Conclusions 53

8. References 55

CRC CARE Technical Report no. 14 ii Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review

Tables

Table 1. Biological endpoints for the determination of contaminant bioavailability in vivo

13

Table 2. Absorption/bioavailability of orally administered benzo[a]pyrene

16

Table 3. Assessment of organic contaminant bioavailability using in vivo animal models

17

Table 4. Cumulative 48 hour elimination (% of dose) of arsenic in urine and faeces of laboratory animals following oral and parenteral administration of inorganic arsenic

21

Table 5. Absolute (comparison to intravenous route) oral bioavailability of arsenic from soil in laboratory animals

22

Table 6. Relative bioavailability (comparison to oral gavage) of arsenic from soil in laboratory animals by comparison of blood concentrations otherwise specified

23

Table 7. Metabolism and urinary excretion of inorganic and organic arsenicals in human following experimental administration

25

Table 10. Relative bioavailability (RBA) ranking order found in various soil and soil-like materials obtained from juvenile swine dosing experiments

27

Table 8. Absolute (comparison to intravenous route) oral bioavailability of lead in laboratory animals

29

Table 9. Relative bioavailability (comparison to oral gavage) of lead from soil in laboratory animals by comparison of blood concentrations otherwise specified

30

Table 11. Composition and in vitro parameters for commonly utilised in vitro bioaccessibility assays (SBRC, IVG, PBET and DIN) for inorganic contaminants

34

Table 12. Methods for the assessment of contaminant bioaccessibility which include esophageal, gastric and intestinal phases

35

Table 13. Methods for the assessment of contaminant bioaccessibility which include gastric and intestinal phases

37

Table 14. Assessment of organic contaminant bioaccessibility using in vitro gastrointestinal extraction methods

44

Table 15. Assessment of arsenic bioaccessibility using in vitro gastrointestinal extraction methods

45

Table 16. Assessment of lead bioaccessibility using in vitro gastrointestinal extraction methods

46

CRC CARE Technical Report no. 14 iii Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review

Figures

Figure 1. Bioavailability processes in soil or sediment, including release of a solid-bound contaminant and subsequent transport, direct contact of a bound contaminant, uptake by passage through a membrane, and incorporation into a living system

7

Figure 2. Correlation between in vivo bioavailability and in vitro bioaccessibility for organic contaminants

48

Figure 3. Correlation between arsenic bioaccessibility (in vitro SBET assay) and arsenic bioavailability (in vivo swine assay) for railway corridor, dip site, mine site, gossan and arsenic-spiked soils

49

Figure 4. Schematic diagram for the determination of contaminant bioavailability and bioaccessibility for human health risk assessment

54

CRC CARE Technical Report no. 14 1 Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review

1. NEPM bioavailability review

The NEPM has undergone a statutory five-year review resulting in a Review Report consented to by the National Environment Protection Council (NEPC) in November 2006. The report makes 27 recommendations including one relating to the provision of guidance on bioavailability and associated leachability testing methods and their application. A Variation Team has been established by the NEPC to implement the report recommendations. Site assessment on land potentially affected by contamination is often undertaken at the termination or transfer of property leases, on property sale, as part of redevelopment to new and often more sensitive land uses, or for sites that pose unacceptable human health and environmental risk and are the subject of statutory assessment and remediation notices. Inadequate assessment that may over or understate the risk of site contamination can result in significant cost burdens for development with resulting negative impacts on economic growth.

An integral part of site assessment involves reference to interim urban ecological investigation levels (EILs) and health investigation levels (HILs) as screening criteria for soil contaminants. Provided competent site investigation processes are followed, chemical analyses of soil that fall below the EILs and HILs usually mean that no further site investigation or management are required depending on the land uses involved. When these levels are exceeded, site-specific qualitative or quantitative health and environmental risk assessment is required. The results of these risk-based approaches are related to proposed land uses and conditions that may be needed for the ongoing management of site contamination.

The NEPM HIL derivation methodology includes the conservative assumption that the contaminant of concern is 100% bioavailable. In most cases the actual bioavailability of the soil contaminant to humans and other organisms is significantly less. An accurate determination of bioavailability enables more realistic site-specific remediation criteria to be developed when a detailed health risk assessment is undertaken. Apart from human health concerns, bioavailability testing methods may also affect determination of acceptable soil criteria for ecological protection in relation to effects of contaminants on domestic and native animals, soil organisms and exotic and native flora. This report specifically addresses Recommendation 24 as scoped by the variation team.

1.1 Scope

The purpose of this specification is to provide the basis for the preparation of a report that incorporates guidance to address the bioavailability and related leachability aspects of Recommendation 24 (below) of the NEPM Review Report.

Undertake a review of current bioavailability and leachability approaches, methods and limitations to provide general guidance in the NEPM for determining their use and application in site assessment.

This specification does not relate to the more limited US EPA Toxicity Characteristic Leaching Procedure SW-846 and other chemical leaching test procedures that provide a basis to estimate the potential of the contaminant to mobilise in various environments and impact on receptors such as surface and groundwater.


1.2 Specification

1.2.1 General

The work undertaken is to be a desktop study only. A draft report must be prepared for consideration by the Review Team.

1.2.2 Specific outcomes

Provide acceptable and relevant definitions related to bioavailability and associated parameters such as relative bioavailability and oral bioaccessibility.

Review and contrast national and international approaches to in vitro and in vivo testing to assess human health impacts.

Identify acceptable bioavailability testing methods for soil contaminants in Australian jurisdictions including related leachability testing and quality assurance and quality control (QA/QC) of procedures.

Discuss the application and limitations of methods including the effects of soil pH and clay content in bioavailability assessment.

Provide practical information on when and how bioavailability testing should be considered using a risk-based tiered framework that includes economic and land use considerations. The framework must ensure that bioavailability considerations are only undertaken when required in the NEPM risk-based approach to ensure that testing is not unnecessarily mandated in the regulation of site assessment.

1.2.3 Presentation

The guidance must be in a form that may be directly incorporated into Schedule B2 of the NEPM or into other Schedules related to risk assessment for direct use by contaminated land professionals. The technical details of methods need not be reproduced but all technical material that includes laboratory methods and QA/QC requirements should be fully referenced. Tabulated information or flow diagram/decision tree approaches that are consistent with the risk assessment processes in the NEPM should be used to assist the user to determine the need for testing in site-specific circumstances. While the draft report may contain information relating to practices in various developed jurisdictions, the details of comparison of methodologies and other relevant scientific material, the recommended guidance component for incorporation in the NEPM should preferably not exceed five A4 pages including any tables and flow diagrams. The NEPM document has been widely adopted by environmental professionals, industry and government agencies alike. The recent review of the NEPMs has resulted in further recommendations, and by addressing those, will improve this document. This report aims to address Recommendation 24.


2. Background

When quantifying exposure to environmental contaminants for human health risk assessment calculations, contaminant bioavailability is assumed to be 100% which presumes that all the contaminant ingested is solubilised in the gastrointestinal tract and absorbed into systemic circulation. In reality, a fraction of the contaminant may only be bioavailable and as such this assumption may grossly overestimate the chemical daily intake thereby influencing risk assessment. Contaminant bioavailability is now regarded as a priority research area for both remediation and risk assessment as it is an important yet poorly quantified regulatory factor.

2.1 National and international approaches

NEPM health investigation levels (HILs) are generally set using a default of 100% bioavailability. This is similar to the approach adopted by many environmental regulatory agencies worldwide. Bioavailability data is helpful for Tier 2 risk assessment process. It is seldom used for setting health criteria; rather, more often than not, for proposing remediation goals. As a result of past and present commercial and industrial activity in many countries, including Australia, there is a need for extensive remediation of contaminated areas.

Although there is general acceptance that contaminant bioavailability is an important consideration in the site assessment process, there is considerable reticence to include bioaccessibility assessment into regulatory guidelines. Typically, this lack of acceptance is based on the following reasons:

bioaccessibility results depend on the assessment method used, the soil type and the contaminant

a method developed and validated for one contaminant is not always appropriate for other contaminants

a method developed and validated for one soil type is not always appropriate for other soil types, and

there are no standard reference materials that can be used to validate the results from bioaccessibility tests or to check their reproducibility (Environment Agency 2007).

Furthermore, most countries do not advocate the incorporation of in vitro bioaccessibility assessment data into risk assessment without supporting evidence from in vivo testing.

Specific analysis of international regulatory guidelines show that:

Canadian regulatory guidance allows the incorporation of bioavailability data from in vivo methods as a more accurate risk assessment on a site-specific basis. Canadian regulators have yet to issue guidance on the use of in vitro bioaccessibility methods for assessing contaminant bioaccessibility.

Dutch researchers at the National Institute for Public Health and the Environment (RIVM) have recommended the RIVM in vitro method for use in site-specific risk


assessments for lead. However, no decision on the inclusion of in vitro methodologies has been made in the Netherlands by regulatory authorities.

The Environment Agency responsible for contamination risk assessment guidelines in the UK has extensively reviewed the current literature and decided not to incorporate bioaccessibility guidelines into risk assessment policy until scientific evidence shows that in vitro data correlates with in vivo data for contaminants, and bioaccessibility methodologies are shown to be robust and reproducible.

In Denmark, Danish regulators support the use of in vitro methods for lead bioaccessibility but only to complement the current risk assessment tools. No formal policy, position or guidance is publicly available.

The US EPA is the only regulatory agency that has validated an in vitro methodology that correlates with in vivo studies for lead contaminated soils. The methodology is based on the correlation between in vivo and in vitro studies for 19 lead contaminated soils and a Standard Operating Procedure has been defined for the in vitro assay (US EPA 2008, 2009). No other in vitro methodologies have been validated for other soil contaminants, although considerable research is focused on in vitro methods for arsenic and polyaromatic hydrocarbons.

Lack of regulatory acceptance of in vitro methods has not limited research into assessing soil contaminant bioavailability issues as this is an active area of scientific research. There are several large international collaborations aimed at improving the understanding of the scientific validity of in vitro research. These include Bioavailability Research Canada (BARC), the Bioavailability Research Group Europe (BARGE) and the Solubility/Bioavailability Research Consortium (SBRC, USA). In Australia, in vitro and in vivo assessment has yet to gain widespread acceptance, although a considerable amount of research has been undertaken to validate in vitro methods for predicting in vivo relative bioavailability.

2.1.1 Key findings

In vivo bioavailability and in vitro bioaccessibility data need to be validated before acceptance by regulatory agencies.

The US EPA has validated an in vitro method for the determination of in vivo lead relative bioavailability on a site-specific basis.


3. Bioavailability and bioaccessibility concepts

The term ‘bioavailability’ has been used extensively in scientific literature; however, its definition may vary depending on discipline-specific designation (Juhasz, Smith & Naidu 2003). In pharmacology, bioavailability is used to describe the fraction of an administered dose of unchanged drug that reaches the systemic circulation, one of the principal pharmacokinetic properties of drugs. By definition, when a medication is administered intravenously, its bioavailability is 100%. However, when a medication is administered via other routes (such as orally), its bioavailability decreases (due to incomplete absorption and first-pass metabolism). Bioavailability is one of the essential tools in pharmacokinetics, as bioavailability must be considered when calculating dosages for non-intravenous routes of administration. This definition under the clinical setting has since been widely adopted by various environmental jurisdictions. Many variations of bioavailability definition have been described and are listed in ‘bioavailability of contaminants in soils and sediments’ (NRC 2003). Taking into consideration multiple exposure pathways, bioavailability is broadly defined as follows:

Bioavailability is the amount of a contaminant that is absorbed into the body following skin contact, ingestion, or inhalation.

More specific definitions for bioavailability may include:

Absolute bioavailability (ABA) – the fraction of a compound which is ingested, inhaled or applied to the skin that is actually absorbed and reaches systemic circulation (NEPI 2000). Absolute bioavailability can be expressed using Equation [1]:

ABA = Absorbed Dose x 100% / Administered Dose Equation [1]

For solid matrices (e.g. soil), it is more practical to determine ABA by comparing absorbed dose in blood or urine of the treatment group via the oral route to that of a soluble reference compound (e.g. sodium arsenate, or lead acetate) via the intravenous injection (representing 100% ABA) and is expressed using Equation [2]:

ABA = (AUCoral-test x Doseiv-reference) x 100% / (AUCiv-reference x Doseoral-test) Equation [2]

Where:

oral = oral gavage

iv = intravenous injection

AUC = the area of under the curve of blood or urinary concentrations of the respective test and reference items over prescribed time intervals.

Alternatively, specific tissue concentration for the respective treatment group has been used in lieu of AUC for the estimation of bioavailability.

The ABA of a compound therefore determines the extent of transfer of the administered dose across absorptive barriers of the organism, and cannot be greater than 100%. When considering an adjustment for human health risk assessment calculation via the inclusion of bioavailability estimates, it is important to note that ABA values are not


always appropriate. For example, it may not be applicable to merely extrapolate ABA values from animal models to humans, from fasted to fed subjects, from one route of exposure to another or from simple (highly soluble) to complex matrices. Some of the confounding factors must be strictly controlled and understood. These will be discussed in the latter sections. More commonly, the relative bioavailability estimate is used (RBA) for adjustment in risk assessment calculations. The RBA of a compound is defined as the comparative bioavailability of the same compound from different exposure media via the oral route (e.g. soil versus water) (Ruby et al. 1999), and is expressed using Equation [3]:

RBA = (ABAtest x 100% / ABAreference) Equation [3]

Again it can be represented by Equation [4]:

RBA=(AUCoral-test x Doseoral-reference) x 100%/(AUCoral-reference x Doseoral-test) Equation [4]

The in vivo method used to estimate the RBA of a compound in a test material compared to that in a reference material is based on the principle that equal absorbed doses will produce the same increase in concentration of the compound in the tissues of exposed animals. Typically, RBA is used to describe the bioavailability of compounds such as arsenic and lead from soil material relative to soluble forms of the compounds in water, both via the oral route (Ruby et al. 1996). When using an ABA value from an animal model for human health risk assessment, the calculation assumes that the compound would have both the same solubility and absorption characteristics in both animal species. However, the assumptions are reduced when RBA values are used instead of ABA. That is, the BA of the soluble form of compounds such as lead and arsenic have been determined in humans, and hence, RBA values from animal models can be corrected and extrapolated to humans (Ruby et al. 1996). This extrapolation accounts for the variation in solubility of the reference material between animal species, and assumes only that the test material behaves proportionally to the reference material in both species.

3.1 Bioavailability processes as defined by NRC (2003)

In light of the differing viewpoints on the definition of bioavailability, the Committee on Bioavailability of Contaminants in Soils and Sediments (NRC 2003) prefers to use the term ‘bioavailability processes’ to encapsulate the mechanisms involved in the dissolution, transport and absorption of environmental contaminants by a receptor organism. Figure 1 (from NRC 2003) provides a visual representation of the term bioavailability processes.


Figure 1. Bioavailability processes in soil or sediment, including release of a solid-bound contaminant and subsequent transport, direct contact of a bound contaminant, uptake by passage through a membrane, and incorporation into a living system. Note that A, B, and C can occur internal to an organism such as in the lumen of the gut (from NRC 2003).

Bioavailability processes are defined (NRC 2003) as the individual physical, chemical and biological interactions that determine the exposure of plants and animals to chemicals associated with soils and sediments. In the broadest sense, bioavailability processes describe a chemical’s ability to interact with the biological world, and they are quantifiable through the use of multiple tools. Bioavailability processes incorporate a number of steps, not all of which are significant for all contaminants or all settings, and there are barriers that change exposure at each step. Thus, bioavailability processes modify the amount of chemical in soil or sediment that is actually absorbed and available to cause a biological response. Presently, our mechanistic understanding of the bioavailability processes described below is highly variable, and quantitative descriptive models of bioavailability processes in most cases are lacking (perhaps with the exception for Pb bioavailability).

‘A’ in Figure 1 – contaminant binding and release – refers to the physical and [bio]chemical phenomena that bind, unbind, expose or solubilise a contaminant associated with soil or sediment. Binding may occur by adsorption on solid surfaces or within a phase like natural organic matter, or by change in form as by covalent bonding or precipitation. Contaminants become bound to solids as a result of chemical, electrostatic and hydrophobic interactions, the strength of which vary considerably. An important aspect governing contaminant-solid interactions is time; with aging, a contaminant is generally subject to transformation or incorporation into a more stable solid phase that can lead to a decrease in contaminant bioavailability. Contaminants can be released to fluid in contact with soil or sediment in response to changes in water saturation, in water and gas chemistry, and in solid surface properties. Biologically induced release is common in natural systems, including release mediated by digestive processes, microorganisms, plants, and bioturbating invertebrates.

Bioavailability processes (A, B, C, and D)

Contaminant interactions between phases

Transport of contaminants to organisms

Passage across physiological membrane

Circulation within organism, accumulation in target organ, toxicokinetics, and toxic effect

C

B

A

Bound contaminant

Released contaminant

Association Dissociation E

Absorbed contaminant in

organism

Site of biological response

Biological membrane

D


‘B’ in Figure 1 involves the movement of a released contaminant to the membrane of an organism, while ‘C’ involves the movement of contaminants still bound to the solid phase. Contaminants dissolved in the aqueous or gas phases are subject to transport processes such as diffusion, dispersion and advection that may carry the contaminant to the surface of a living organism. These same processes can also transport contaminants still bound to small solid particles (colloids) to within close proximity of potential receptors. As contaminants are being transported, they can undergo transformation reactions (including oxidation-reduction reactions, hydrolysis, acid-base reactions, and photolysis) that can affect greatly affect the bioavailability and toxicity of the contaminant. It should be noted that if association-dissociation processes have occurred internally (as in the gut lumen), fate and transport processes prior to uptake across a biological membrane may be limited.

The bioavailability process depicted as ‘D’ entails movement from the external environment through a physiological barrier and into a living system. Because of the enormous diversity of organisms and their physiologies, the actual process of contaminant uptake into a cell – or factors that may impede or facilitate uptake – varies depending on receptor type. One common factor among all organisms is the presence of a cellular membrane that separates the cytoplasm (cell interior) from the external environment. Most contaminants must pass through this membrane (by passive diffusion, facilitated diffusion, or active transport) before deleterious effects on the cell or organism occur. For bacteria and plants, contaminants must be dissolved in the aqueous phase before they can be taken up. However, elsewhere in the natural world there are exceptions to the notion that bioavailability is directly dependent on solubility. For example, contaminant-laden particles that undergo phagocytosis can be delivered directly into some cells (although within the cell the contaminant may eventually need to be solubilised to reach its site of biological action). Uptake mechanisms relevant to humans include absorption across the gut wall, the skin, and the lining of the lungs.

‘E’ in Figure 1 refers to paths taken by the chemical following uptake across a membrane, for example, metabolic processing or exerting a toxic effect within a particular tissue. In general, the magnitude and the nature of the effect will be determined by the form and concentration of the chemical at its active site(s). If concentrations of the chemical achieved at the biological targets are too low, or if the chemical has been converted to a form that no longer interacts with the target, no effect will be observed. On the other hand, exposure may lead to concentrations that are sufficiently high so as to be lethal. Between these extremes is the potential for non-lethal, yet deleterious effects such as reduced metabolic activity, impaired reproduction, and increased sensitivity to physical or chemical stresses. Of particular importance is the bioaccumulation of contaminants (e.g. polychlorinated biphenyls or PCBs) to storage sites within tissues that are often inaccessible to normal elimination mechanisms such as metabolism and excretion. Slow release of the chemical from these storage sites can result in protracted ‘exposure’ within the body even when exposure outside the body has been reduced.

Bioaccumulated contaminants may become available at some point to higher order organisms that eat the plant or animal in which the contaminants are stored. In fact, food chain transfer is probably a more important exposure pathway to contaminants in soils and sediment for higher-order animals than is direct ingestion of soil or sediment. Depending on the extent of bioaccumulation in each organism, animals can be


exposed to contaminants at concentrations higher than those found in the solids from which the compound originated (biomagnification).

The potential for the consideration of bioavailability processes to influence risk-based decision-making is greatest where certain chemical, environmental and regulatory factors align, that is:

where the contaminant is (and is likely to remain) the risk driver at a site

where the default assumptions made for a particular site are inappropriate

where significant change to remedial goals is likely (e.g. because large amounts of contaminated soil or sediment are involved)

where conditions present at the site are unlikely to change substantially over time, and

where regulatory and public acceptance is high.

These factors should be evaluated before committing the resources needed for a detailed consideration of bioavailability processes (NRC 2003).

Another related term is bioaccessibility – the fraction of a compound that is soluble in the gastrointestinal tract and is therefore available for absorption – which is specifically referred to when in vitro assessment models are used (Rodriguez et al. 1999; Ruby et al. 1996, 1999).

There is evidence about the inter-mixed use of the bioavailability and bioaccessibility terms. It has caused confusion if non-conforming definitions are applied in the literature.

Key findings

Bioavailability is the amount of a contaminant that is absorbed into the body following skin contact, ingestion or inhalation:

absolute bioavailability is the fraction or percentage of a compound which is ingested, inhaled or applied to the skin that is actually absorbed and reaches systemic circulation

relative bioavailability is the ratio of the absorbed fraction from the exposure medium in the risk assessment (e.g. soil) to the absorbed fraction from the dosing medium used in the critical toxicity study.

Bioaccessibility refers to the fraction of a compound that is soluble in the gastrointestinal tract and is therefore available for absorption – which is specifically referred to when in vitro assessment models are used.


3.2 Human contaminant exposure

Human exposure to contaminants may result from a variety of pathways including inhalation, oral ingestion or dermal absorption. Inhalation of contaminants in a volatile state is dependent on the compound’s vapour pressure, although exposure via this pathway may occur through inhalation of contaminated dust and particulate matter. The major non-dietary pathway for contaminant exposure is via incidental ingestion of contaminated soil. When contaminants are added to a soil environment they bind to soil in a process known as sequestration, which is believed to result in a decrease in contaminant bioavailability over time. The extent of bioavailability reduction may vary depending on the nature of the soil and the physico-chemical properties of the contaminant. When assessing the potential risk associated with contaminants, often the total concentration of the soil-borne contaminant is included into risk models, however this does not account for bioavailability issues. As a result, the risk posed by soil-bound contaminants to human and ecosystem health may be overestimated due to the assumption that all the soil-bound contaminant is bioavailable. In order to determine the extent to which humans are at risk from exposure to chemicals in the environment, an understanding of contaminant bioavailability is essential.

The following sections focus on the soil ingestion pathway for human health exposure to environmental contaminants. It aims to highlight methodologies available for the in vivo assessment of contaminant bioavailability and the development and implementation of surrogate assays (in vitro) for its determination.

3.2.1 Ingestion and release of contaminants from the soil matrix

Following incidental ingestion of contaminated soil, a number of digestive processes may lead to release of contaminants from the soil matrix and potential absorption into systemic circulation. Initially ingested material is exposed to a variety of saliva enzymes (e.g. amylase, lipase) in the oral cavity. Following mastication, ingested material passes into the stomach via the oesophagus canal as a result of peristaltic action. In the stomach, material is exposed to a low pH environment (pH 1–4 depending on the degree of fasting) as a result of the secretion of hydrochloric acid from the stomach cell wall mucous membrane under gastric condition (Johnson 2001). Following transition of the chyme (partially digested food, hydrochloric acid and enzymes) from the stomach to the duodenum, bile and pancreatin is secreted from the gall bladder and pancreas respectively while bicarbonate in pancreatin juices neutralises the pH of the small intestine (Johnson 2001). In addition, other enzymes such as pepsin, amylases, lipases, proteases and nucleases are secreted by the pancreas for the breakdown of carbohydrates, fats and proteins (Gorelick & Jamieson 1994).

From the duodenum, the chyme passes through the ileum and jejunum respectively. During this transition, the chyme is in contact with enterocytes, the most common epithelial cells that are principally responsible for fat, carbohydrate, protein, calcium, iron, vitamins, water and electrolyte absorption (Hillgren, Kato & Borchardt 1995). Each enterocyte contains several microvilli (Madara & Trier 1994) which results in a large surface area (approximately 200 m2 for adults) for absorption within the small intestines (Tso 1994).

In cells there are mechanisms for metal ion homeostasis that involves a balance between uptake and efflux in order to maintain normal bodily physiological functions.


Metal transporters have been periodically discovered that transport metals across cell membranes and organelles inside the cells. The importance of metal transporters is best illustrated by copper-transport proteins involved in Menkes disease (copper deficiency) and Wilson disease (copper overload).

Due to the hydrophobic and lipophilic nature of organic contaminants, these compounds traverse the intestinal membrane to reach systemic circulation via the lipid transport pathway (Seydel & Schaper 1982). The amphipathic nature of bile salts micelles in chyme act as a vector for organic contaminant transportation in the duodenum. Water and electrolytes are also absorbed from the large intestine and colon after which residuals (indigestible material and bacterial) are excreted.

A uniform mechanism for all toxic metals is implausible because of their wide variety of chemical and toxic properties. In general terms, metals in their ionic form can be very reactive and interact with biological systems in various ways. For example, cadmium and mercury readily bind to sulfur in proteins (cycteine sulphur of amino acid residues) as a preferred bio-ligand resulting in dysfunction of biomolecules (Kasprzak 2002). The preferential binding of thiol groups can also inhibit the normal function of many enzymes in the body. One of the best known examples is the way Pb interacts with enzymes involved in the haem synthesis pathways, and results in alteration of porphyrin profile (Moore et al. 1987). Similarly, arsenic can also interfere with haem pathways (Krishnamohan et al. 2007a, 2007b; Ng et al. 2005).

Metals can also exert their toxicity via mimicry of essential elements by binding to physiological sites that are normally reserved for essential elements. For example:

cadmium, copper and nickel can mimic zinc

thallium can mimic potassium

manganese can mimic iron

arsenate and vanadates can mimic phosphate

selenate, molydate, and chromate can mimic sulphate and compete for sulphate transporters and in chemical sulfation reaction (Bridges & Zalpus 2005).

Ionic metals can form adducts with DNA and protein molecules. For example, once hexavalent chromium is absorbed (entered the cells) it can be reduced to reactive trivalent chromium species that form adducts with DNA or protein. If the adducts are not repaired it could result in mutagenicity (Zhitkovich 2005).

Metals can directly act on catalytic centres for redox reactions and induce oxidative modification of biomolecules such as proteins or DNA. This may be a significant pathway leading to carcinogenicity by certain metals (Kasprzak 2002). On the other hand, some metals/metalloids such as arsenic can produce free radicals during its metabolism and result in oxidative damage (Wang et al. 2009), mutagenesis (Ng et al. 2001) and carcinogenesis (Krishnamohan 2008) of the receptor. The metabolism of arsenic can also explain its inhibitory effect on DNA repairs (Tran et al. 2002).

Following absorption into systemic circulation, organic contaminants may be metabolised to daughter products, accumulated in selected organs or fatty tissue or excreted as parent and/or daughter products (Fries, Marrow & Somich 1989). In addition, exposure to organic contaminants may result in the expression of various


enzymes (cytochrome P450 monooxygenases, mixed function oxidases, epoxide hydrolase etc.) or the induction of DNA adducts (Bauer et al. 1995; Ramesh et al. 2004). Prolonged exposure may result in a variety of adverse health effects as outlined in IARC (1983), ATSDR (2000) and Ramesh et al. (2004).

Key findings

Contaminant exposure may occur via a number of exposure pathways, but soil ingestion is the major exposure pathway for many contaminants.

Contaminant exposure may affect various metabolic processes including those at the cellular level.

3.3 Methods for the determination of contaminant bioavailability

In order to determine the bioavailability of contaminants in soil for human health exposure assessment, in vivo assays utilising animal models have been applied. For inorganic contaminants, mice, rats, rabbits, dogs, swine, cattle and primates have been used. Being closely related to man, primates are the first choice for bioavailability studies, however the cost associated with their use are prohibitive (Rees et al. 2009). Young swine are considered to be a good physiological model for the gastrointestinal absorption of contaminants in children (Weis et al. 1991). An outline of advantages of utilising swine for contaminant bioavailability assessment is provided in Weis et al. (1991). Rodents are the most commonly used vertebrate species for bioavailability studies because of their availability, size, low cost and ease of handling. However, the endpoint used for the assessment for contaminant bioavailability will be influenced by the animal model. For example, only single blood samples may be appropriate/feasible for experiments conducted with small laboratory rodents (e.g. mice), while repeated blood sampling may be suitable for swine and primate studies.

These in vivo assays involve dosing an animal with a reference or test material (intravenously and orally via gavage or incorporation into daily feed) and monitoring bioavailability endpoints following an exposure period. Bioavailability endpoints may include the determination of the contaminant in blood, organs, fatty tissue, urine and faeces, urinary metabolites, DNA adducts and enzyme induction (e.g. cytochrome P450 mono-oxygenases). Table 1 provides details of bioavailability endpoints including advantages and disadvantages of each method.

In brief, common endpoints used to monitor bioavailability include:

blood (e.g. arsenic, lead)

urinary excretion (e.g. arsenic)

faecal excretion

target organs (adipose tissue for organic contaminants, liver and kidney for cadmium).

No single method of in vivo bioavailability has been identified as being suitable for all inorganic contaminants and the methodologies selected are dependent on the contaminant of interest and the resource available to undertake the bioavailability studies.

Table 1. Biological endpoints for the determination of contaminant bioavailability in vivo.

Endpoint Contaminant Advantages Disadvantages References

Blood PAHs

Dioxin

PCDD/F

As, Pb

Measures bioavailable fraction (concentration in systemic circulation). Clearance times can be calculated.

Limitation in sampling frequent for small animal models. Large administration dose required for contaminant detection.

van Schooten et al. 1997 Cavret et al. 2003 Wittsiepe et al. 2007 Juhasz et al. 2007b, 2008 Juhasz, Weber et al. 2009

Urine PAHs

As

Abundant sampling Relies on the quantification of a limited number of transformation products. Detection issues at low concentrations.

van Schooten et al. 1997 Buckley and Lioy 1992 Jongeneelen et al. 1988 Rodriguez & Basta 1999

Faeces PAHs Abundant sampling Represents the non-bioavailable fraction. Issues with extraction recoveries from matrix.

van Schooten et al. 1997

Adipose tissue PAHs

PCDD/F

Abundant tissue Useful only for some hydrophobic contaminants

ATSDR 1995 Wittsiepe et al. 2007

Target organ PAHs

PCDD/F

Cd

Some contaminants may accumulate in specific organs.

Not applicable for some organs and contaminants.

Wall et al. 1991 Robinson et al. 1975 Wittsiepe et al. 2007 Schroder et al. 2004

DNA adduct formation PAHs

Biomarker for carcinogenicity

Studies.

Expensive and time consuming. Correlation with bioavailable fraction unknown. Adduct formation not specific for individual compounds.

Schoket et al. 1993

P450 Monooxygenase induction

PAHs

Relevant to POP metabolism and pharmacokinetics studies.

Hard to detect. Correlation with bioavailable fraction unknown. Enzyme induction not specific for individual compounds.

Zhang et al. 1997a, 1997b Roos 2002 Thomas et al. 1983 Chaloupka et al. 1995


3.3.1 Assessment of contaminant bioavailability – organics

Many pharmacokinetic studies have been undertaken to determine the uptake of organic compounds but almost no studies have investigated organic contaminant bioavailability in soils. In these studies, the contaminant of interest is administered to the test animal in a suitable vehicle (oil, synthetic diet etc.) intravenously or via gavage, and bioavailability endpoints determined over a time course period. The mode of administration may vary depending on whether the relative or absolute bioavailability is being determined. In these studies, organic contaminant bioavailability varied depending on the animal model used, bioavailability endpoint assessed, dose and the dosing vehicle. For example, Ramesh et al. (2004) provides a summary of absorption/bioavailability data for orally administered PAHs. The bioavailability of orally administered benzo[a]pyrene varied from 5.5–102% (Ramesh et al. 2004) illustrating the variability in bioavailability data when various animal models and dosing vehicles are used (see Table 2). Bioavailability data derived from the aforementioned studies are important in order to establish dose responses for reference compounds. In addition, data of this type may be utilised in comparison calculations, i.e. bioavailability of a contaminant in a matrix (e.g. soil or food) relative to the bioavailability of a reference dose.

While a considerable amount of bioavailability data is available for pure organic compounds, only a limited number of studies have been undertaken to assess the bioavailability of organic contaminants in soil (see Table 3). In a number of studies presented in Table 3, absolute or relative bioavailability were not calculated, however the impact of soil matrix and contaminant ageing on organic contaminant bioavailability was noted. For example, Bordelon et al. (2000) assessed the bioavailability of PAHs from coal tar in Fischer 344 male rats using DNA adduct formation as the endpoint for bioavailability assessment. Coal tar was spiked into soil which was then incorporated into rodent diets, resulting in a PAH concentration of 100–250 mg kg-1 feed. Following a 17 day feeding trial, DNA adducts were assessed in both liver and lung tissue. While DNA adducts were detected in both liver and lungs, adduct levels were three-fold higher in lung DNA compared to hepatic DNA. However, when spiked soils were aged for nine months, the soil-POP residence time had no effect on the bioavailability of coal tar constituents.

A similar effect of matrix on organic contaminant bioavailability has been observed by other researchers (Table 3). Koganti et al. (1998) determined that the presence of soil reduced PAH bioavailability when assessed using a mouse model. Urinary excretion of 1-hydroxypyrene decreased by 9–75% when the mouse diet contained PAH-contaminated soil compared to organic extracts of the soil.

Determination of the absolute or relative organic contaminant bioavailability in contaminated soil is shown in Table 3. Pu et al. (2004; 2006) determined the bioavailability of phenanthrene and PCBs in spiked soil with varying soil properties. In both cases, bioavailability was determined by monitoring blood-contaminant concentrations in male Sprague-Dawley rats following a single gavage dose of contaminated soil. Blood-contaminant concentration versus time data were plotted and absolute or relative bioavailability determined by comparing soil dose data to intravenous and oral (phenanthrene dissolved in corn oil) administered doses. For phenanthrene, the absolute bioavailability of an oral dose administered in corn oil was approximately 25% (at 400 and 800 µg kg-1 body weight doses). However, the absolute bioavailability of phenanthrene-spiked soil varied depending on the soil type and dose


administered. Absolute phenanthrene bioavailability ranged from 15–49% and 22–23% at doses of 400 and 800 µg kg-1 respectively. Generally, the relative phenanthrene bioavailability (compared to the corn oil administered dose) was <100%, however, when phenanthrene was spiked into a low organic carbon and clay content soil, the relative bioavailability of phenanthrene was 203% and 138% for the 400 and 800 µg kg-1 doses respectively. The authors did not provide any suggestions as to the reasons for the enhancement in phenanthrene bioavailability in this soil type.

When Pu et al. (2006) assessed PCB bioavailability in the Sprague-Dawley rat model, the absolute bioavailability of corn oil administered PCB #52 and PCB #118 was 53% and 70% while the absolute bioavailability of PCB #52 and PCB #118 spiked soils ranged from 53–67% and 61–70% respectively. While the absolute bioavailability of PCBs was reduced when administered in soil by gavage, soil properties (i.e. organic carbon content) did not appear to have a significant effect on PCB bioavailability. In a similar PCB experiment using 14C-labelled compounds (PCB #18, #52, #101), Fries, Marrow and Somich (1989) also observed a small reduction in PCB bioavailability when compounds were administered in soil versus corn oil although the absolute bioavailability of these compounds ranged from 78–88% (Table 3).

In vivo bioavailability assessment has also been undertaken with PCDD/F-contaminated soil. In the study of Wittsiepe et al. (2007), PCDD/F bioavailability was assessed by determining the concentration of PCDD/Fs in blood, liver, brain, muscle and adipose tissue of swine following daily oral doses of contaminated soil or soil extracts for 28 days. The absolute bioavailability of PCDD/Fs was significantly less when PCDD/F soil doses were administered in feed compared to when soil extracts were administered following their addition to feed. 2378-chlorosubstituted congener absolute bioavailability ranged from 0.6–22% for soil compared to 3–60% when administered via soil extracts in the diet. When relative bioavailability was calculated, values ranged from 2–42% for the 2378-chlorosubstituted congeners. Using an alternative bioavailability endpoint (EROD induction), Budinsky et al. (2007) determined the relative bioavailability of PCDD/Fs to be 20 and 25% for two soils of varying origin with TCDD/F TEQs of 651 and 264 respectively. When swine (EROD induction) bioavailability values were compared to values derived using an alternative animal model (Sprague-Dawley rats and EROD induction), relative PCDD/F bioavailability was 1.8–2.4 times higher when measured using the rat model. Budinsky et al. (2007) proposed that the variability in bioavailability protocols (i.e. administration of soil wrapped in dough for swine versus incorporation of soil into lipid rich feed for rats) and/or species differences in gastrointestinal absorption of PCDD/Fs may account for the difference in bioavailability measurements.

Table 2. Absorption/bioavailability of orally administered benzo[a]pyrene (adapted from Ramesh et al. 2004).

Animal model Benzo[a]pyrene

dose (mg kg-1)

Dosing vehicle Absorption/ bioavailability (%)

Reference

Rats

F-344 (male) 0.37-3.7 Peanut oil 91 Hecht, Grabowski & Groth 1979

F-344 (male) 0.002 Char-broiled hamburger 89 Hecht, Grabowski & Groth 1979

F-344 (male) 100 Peanut oil 40 Ramesh et al. 2001

SD (male) 2-60 20% emulphor and 80% isotonic glucose solution

>90 Bouchard & Viau 1997

SD (male) 0.002 Gavage >80 Yamazaki & Kakiuchi 1989

Wistar (male) 1000 Synthetic diet 89 Rabache, Billaud & Adrian 1985

Wistar 1 mmol Olive oil 28 Van de Wiel et al. 1993

Hamster (male Syrian golden)

0.16-5.5 Corn oil + synthetic diet 97 Mirvish et al. 1981

Sheep (female merino) 10-20 µCi +

1 mg (unlabelled)

Toluene and lucerene chaff 54-67 West & Horton 1976

Goat 2.5 x 106 Bq Vegetable oil 5.5 Grova et al. 2002

Swine 50 µCi Spiked milk 33 Laurent et al. 2002

Table 3. Assessment of organic contaminant bioavailability using in vivo animal models.

Animal model Contaminant Contaminated matrix Bioavailability Reference

Mouse (urine, DNA adducts)

PAHs Manufacturing gas plant soil: total PAHs up to 3120 mg kg-1. Soil organic extracts were also administered to mice.

The absolute or relative bioavailability of PAHs was not determined, however, the presence of soil resulted in a considerable decrease (9-75%) in PAH bioavailability as measured by 1-hydroxypyrene in the urine.

Koganti et al. 1998

Sprague-Dawley male rats (blood)

PAHs, phenanthrene Soils with varying properties (TOC: 0.5-1.7%; Clay: 8-38%; pH: 4.6-7.4) were spiked with phenanthrene. Soils were administered to rats orally at two dose levels (400 and 800 µg kg-1 bw). Phenanthrene was also supplied orally in corn oil and via intravenous injection.

The absolute bioavailability of phenanthrene in corn oil was 24% and 25% for the 400 and 800 µg kg-1 body weight (bw) doses respectively. At the 400 µg kg-1 bw dose, the absolute bioavailability of phenanthrene in soil ranged from 15-49%. At 800 µg kg-1 bw dose, the absolute bioavailability of phenanthrene ranged from 22-34%.

Pu et al. 2004

Fischer 344 male rats (urine, organs)

PAHs Coal tar spiked soils. Ageing of coal tar spiked soils had minimal impact on coal tar bioavailability, however, incorporation of coal tar into soil resulted in decreased bioavailability as demonstrated by a reduction in 1-hydroxypyrene elimination and a lower concentration of PAHs in the liver.

Reeves et al. 2001

Lewis rats (urine, faeces, blood endpoints)

PAHs

Anthracene, pyrene and benzo[a]pyrene as pure compounds

Industrial contaminated soil with benzo[a]pyrene at 9.2 mg kg-1.

Low faecal excretion of parent compounds was observed for both pure compounds and contaminated soils. Urinary excretion of 1-hydroxypyrene was 17.7 times higher when pyrene was dosed as a pure compound compared to soil.

Van Schooten et al. 1997

Fischer 344 rats (faeces)

PAHs Manufacturing gas plant soil: total PAHs from 7-1040 mg kg-1.

90-95% of PAHs were recovered in the faeces. The uptake of PAHs from treated soil (3 months of air sparging) was reduced to 60%.

Stroo et al. 2000

Sprague-Dawley rats (EROD activity, DNA adducts)

PAHs PAH contaminated soil from a coke plant: total PAHs at 509 mg kg-1.

The absolute or relative bioavailability of PAHs was not determined, however, EROD activity was enhanced in the presence of PAHs.

Fouchécourt et al. 1999

Fischer 344 male rats (DNA adducts)

PAHs Coal tar contaminated soils: total PAHs at 3200-3500 mg kg-1. Following incorporation into rodent diets, total PAH concentrations were 100-250 mg kg-1.

Liver weights as a percentage of body weight were 30% higher in animals exposed to coal tar. DNA adduct levels were 3-fold higher in lung DNA compared to hepatic DNA. The presence of soil decreased the bioavailability of genotoxic components in the coal tar, however, ageing (9 months) had no effect on bioavailability.

Bordelon et al. 2000

Table 3 (cont.) Assessment of organic contaminant bioavailability using in vivo animal models.

Animal model Contaminant Contaminated matrix Bioavailability Reference

Swine (organs, tissue)

PCDD/F Animals were administered daily doses for 28 days of 0.5 g soil kg-1 bw (2.63 ng I-TEq/(kg bw . d) or 1.58 ng I-TEq/(kg bw . d) provided in the diet via solvent addition.

The absolute bioavailability of 2,3,7,8-chlorosubstituted congeners ranged from 0.6–22% from soil and 3–60%when administered via solvent in the diet. The relative bioavailability of 2,3,7,8-chlorosubstituted congeners in soil ranged from 2–42%.

Wittsiepe et al. 2007

Sprague-Dawley female rats and swine (EROD induction)

PCDD/F Flood plain soil (TEQ = 651) and urban soil (TEQ = 264).

In the rat, the relative bioavailability of PCDD/Fs in flood plain and urban soil was 37% and 60% respectively compared to PCDD/Fs administered in corn oil. In swine, PCDD/F bioavailability was 20% and 25%.

Budinsky et al. 2007

Sprague-Dawley male rats (blood)

PCBs ,#52 and #118

Soils with varying properties (TOC: 0.5-2.9%; Clay: 8-38%; pH: 4.6-7.4) were spiked with PCBs. Soils were administered to rats orally at a dose of 600 µg kg-1 bw. PCBs were also supplied orally in corn oil and via intravenous injection.

The absolute bioavailability of PCB #52 ranged 53-67% while the relative bioavailability ranged from 109-126%. The absolute bioavailability of PCB #118 ranged 61-70% while the relative bioavailability ranged from 87-99%.

Pu et al. 2006

Sprague-Dawley male rats (urine, faeces, body fat)

PCBs Soil was spiked with 0.25,-1.0 µCi kg-1 14C-PCBs (#18, #52, #101).

PCB bioavailability in spiked soil was 78, 88 and 77% for #18, #52, #101 respectively.

Fries, Marrow & Somich 1989


3.3.2 Assessment of contaminant bioavailability – inorganics

In contrast to the limited studies undertaken to determine organic bioavailability in contaminated soils, a considerable amount of research has been undertaken on inorganic contaminant bioavailability. This may be attributed to the relatively uncomplicated approach in determining inorganic contaminants in animal tissues. The two most commonly studied inorganic contaminants are arsenic and lead.

Arsenic

The absorption coefficient for arsenic has been reported as 0.98 with a range of 0.70–0.98 (Owen 1990), suggesting soluble arsenical compounds are readily accessible for uptake by a receptor. Both pentavalent and trivalent arsenic are rapidly and extensively absorbed from the gastrointestinal tract of common laboratory animals following a single oral dose (see Table 4). Based on the mouse data of Vahter and Norin (1980), arsenite is more extensively absorbed from the gastrointestinal (GI) tract compared to arsenate at lower doses (e.g. 0.4 mg As/kg), whereas the reverse is true at higher doses (e.g. 4.0 mg As kg-1). The latter is consistent with the greater whole body retention of arsenite compared to arsenate at higher doses. As demonstrated in a small number of Australian studies (Ng & Moore 1996; Ng, Bruce & Noller 2003, Ng et al. 1998), arsenic speciation of soil is important so that comparison to appropriate positive control can be made when calculating bioavailability. These studies also illustrate that bioavailability is dependent on solubility and the parent compounds, including arsenic species sodium arsenate, sodium arsenite and calcium arsenite. Ng and Moore (1996) reported in a rat study that arsenic-contaminated soils obtained from formal cattle tick dip sites contained significant amounts of arsenite, and most likely in the calcium arsenite form. Liming was a general practice to precipitate arsenical compound in the dip bath, forming calcium arsenite for disposal purpose. This relatively insoluble trivalent arsenical remains unchanged in the soil for several decades, in contrast to the generalisation that arsenite converts to arsenate readily in the soil. The study demonstrates that the bioavailability of sodium arsenite, sodium arsenate and calcium arsenite differs. Soil containing natural mineral phase could also contain a significant proportion of arsenite (Ng et al. 1998). It is important, in this case, to include positive control groups of rats dosed with these different arsenical compounds separately.

Studies using mice conducted by Odanaka, Matano and Goto (1980) suggest that much less pentavalent arsenic is absorbed from the GI tract following oral administration compared to the results of Vahter & Norin (1980) – 48.5% of dose (5 mg kg-1) in urine compared to 89% dose (4 mg kg-1) in urine. This difference may be attributable to the fact that the mice in the study of Vahter and Norin were not fed for at least two hours before and 48 hours after dosing, whereas the mice in the Odanaka, Matano and Goto studies were not food restricted. Kenyon, Hughes and Levander (1997) found that feeding a diet lower in fibre or ‘bulk’ to female B6C3F1 mice increased absorption of pentavalent arsenic by ~10%, compared to standard rodent chow diet. The bioavailability of arsenic from soils has been assessed using various animal models because this can be a significant issue in risk assessment for contaminated industrial sites where there is potential for arsenic exposure via soil ingestion. Absolute bioavailability and relative (comparative) bioavailability data are summarised in Tables 5 and 6. These studies indicate that oral bioavailability of arsenic in a soil or dust vehicle is considerably lower compared to the pure soluble salts typically used in toxicity studies. Davis et al. (1992) have pointed out that this is due mainly to mineralogic factors which control solubility in the gastrointestinal tract,


such as the solubility of arsenic bearing mineral itself and encapusulation within insoluble matrices (e.g. silica). It is worthy to note that a fasting human stomach has a pH of about 1.3–1.5, higher pH of 2.5 for a partially fed and 4.5 for a fully fed stomach respectively. The small intestinal tract has a near-neutral pH of about 7. Obviously these pH values alone will influence the solubility of contaminants throughout the digestive system once ingested.

In a recent study conducted with Australian soils, Juhasz et al. (2007b) utilised an in vivo swine assay for the determination of arsenic bioavailability in contaminated soils. Arsenic bioavailability was assessed using pharmacokinetic analysis encompassing area under the blood plasma-arsenic concentration time curve following zero correction and dose normalisation. In contaminated soil studies, arsenic uptake into systemic circulation was compared to an arsenate oral dose and expressed as relative arsenic

bioavailability. Arsenic bioavailability ranged from 6.9 5.0 to 74.7% 11.2% in 12 contaminated soils collected from former railway corridors, dip sites, mine sites and naturally elevated gossan soils. Arsenic bioavailability was generally low in the gossan

soils and highest in the railway soils, ranging from 12.1 8.5 to 16.4 9.1% and 11.2

4.7 to 74.7 11.2% respectively.

In a follow-up study, Juhasz et al. (2008) assessed arsenic bioavailability in spiked soils aged up to 12 months using in vitro and in vivo methodologies. Ageing (natural attenuation) of spiked soils resulted in a decline in in vivo arsenic bioavailability (swine assay) of over 75% in Soil A (Red Ferrosol), but had no significant effect on in vivo arsenic bioavailability even after 12 months of ageing in Soil B (Brown Chromosol). Sequential fractionation, however, indicated that there was repartitioning of arsenic within the soil fractions extracted during the time course investigated. In Soil A, the arsenic fraction associated with the more weakly bound soil fractions decreased while the residual fraction increased from 12% to 35%. In contrast, little repartitioning of arsenic was observed in Soil B, indicating that natural attenuation may be only applicable for arsenic in soils containing specific mineralogical properties.

In results similar to those found using experimental animals, arsenic ingestion studies in humans indicate that both tri- and pentavalent arsenic are well absorbed from the GI tract (see Table 7). For example, Pomroy et al. (1980) reported that healthy male human volunteers excreted 62.3 ± 4.0% of a 0.06 ng dose of 74As-arsenic acid (AsV) in urine over a period of seven days, whereas only 6.1 ± 2.8% of the dose was excreted in the faeces. Few other controlled human ingestion studies have actually reported data on both urine and faecal elimination of arsenic. However, between 45% and 75% of the dose of various trivalent forms of arsenic are excreted in the urine within a few days (see Table 4), which suggests that gastrointestinal absorption is both relatively rapid and extensive.

Table 4. Cumulative 48 hour elimination (% of dose) of arsenic in urine and faeces of laboratory animals following oral and parenteral administration of inorganic arsenic (i.v. = intravenous; s.c. = subcutaneous).

Species Arsenic form Dose Route Urine Faeces Total Reference

Rat arsenic acid 5 mg/kg Oral 17.2 33.0 50.2 Odanaka, Matano & Goto1980

1 mg/kg i.v. 51.0 0.8 51.8

Hamster arsenic acid 5 mg/kg Oral 43.8 44.1 87.9 Odanaka, Matano & Goto1980

1 mg/kg i.v. 83.9 4.0 87.9

Hamster arsenic trioxide 4.5 mg/kg Oral 43.5 9.4 52.9 Yamauchi & Yamamura 1985

Mouse arsenic acid 5 mg/kg Oral 48.5 48.8 97.3 Odanaka, Matano & Goto1980

1 mg/kg i.v. 86.9 2.6 89.5

Mouse sodium arsenate 0.4 mg As/kg s.c. 86 6.4 92.4 Vahter & Norin 1980

0.4 mg As/kg Oral 77 8.0 85

4.0 mg As/kg Oral 89 6.1 95.1

Mouse sodium arsenite 0.4 mg As/kg s.c. 73 3.8 76.8 Vahter & Norin 1980

0.4 mg As/kg Oral 90 7.1 97.1

4.0 mg As/kg Oral 65 9.1 74.1

Mouse sodium arsenate 0.00012 mg As/kg Oral 65.0 16.5 81.5 Hughes, Menache & Thompson 1994

0.0012 mg As/kg Oral 68.3 13.5 81.8

0.012 mg As/kg Oral 72.1 10.5 82.6

0.12 mg As/kg Oral 71.0 14.6 85.6

1.2 mg As/kg Oral 68.7 18.2 86.9

Rabbit sodium arsenite 0.050 mg As/kg i.p. 75.7 9.9 85.6 Marafante et al. 1982

Table 5. Absolute (comparison to intravenous route) oral bioavailability of arsenic from soil in laboratory animals.

Species Duration i.v. dose Soil or sample Soil dose Bioavailability Method Reference (hours) (%, mean±SD)

International data

Beagle dog 120 2 mg As5+ Netherlands bog Ore

6.6-7.0 mg As 8.3±2.0 AUC for urinary excretion

Groen et al. 1994

New Zealand white rabbit

120 1.95 mg As5+/kg

Smelter impacted soil (Anaconda, Montana USA)

0.78 mg As/kg 1.95 mg As/kg 3.9 mg As/kg

24±3.2 AUC for urinary excretion, no dose-dependency observed

Freeman, Johnson, Llao et al. 1993

sodium arsenate 1.95 mg/kg 50±5.7

Cynomolgus monkey

168 0.62 mg As5+/kg

Soil (Anaconda, Montana USA)

0.62 mg As/kg 14 (11) AUC for urinary excretion (AUC

Freeman et al. 1995

House Dust (same location)

0.26 mg As/kg 19 (10) for blood in parentheses)

sodium arsenate 0.62 mg As/kg 68 (91)

Immature swine 144 0.01 - 0.31 mg As/kg

Soil

0.04 - 0.24 mg As/kg 52 AUC for blood US EPA 1996

slag 0.61 - 1.52 mg As/kg 28

sodium arsenate 0.01 - 0.11 mg As/kg 68

Australian data

Wistar rat 96 0.5 mg As3+/kg Soil (Canberra, 0.5 mg As/kg 4.31-9.87* AUC for urinary Ng et al. 1998 Australia) (n=6) 1.27-2.98# excretion

0.5 mg As5+/kg 5 mg As/kg 1.02-1.96* (n=4) 0.26-0.67#

Sprague Dawley rat

168

0.5 mg As3+/kg 0.5 mg As5+/kg

Gold mine tailings, waste rock, heap leach material; Former arsenic mine mixed waste

0.5 mg As/kg (n=4)

0.5±0.14 - 8.0±1.2

AUC for urinary excretion

Bruce 2004

Cattle 240 0.5 mg As3+/kg 0.5 mg As5+/kg


1.0 mg As/kg (n=3)

7±0.34 - 58±6.49 AUC for blood Bruce 2004

Where n = no. of composite soil samples; * compared to arsenite AUC; # compared to arsenate AUC

Table 6. Relative bioavailability (comparison to oral gavage) of arsenic from soil in laboratory animals by comparison of blood concentrations otherwise specified .

Species Duration Oral gavage Soil or sample Soil dose Bioavailability Reference (hours) (%, mean±SD)

International data


120 1.95 mg As5+/kg Smelter impacted soil (Anaconda, Montana USA)

0.78 mg As/kg 1.95 mg As/kg 3.9 mg As/kg

505.7

Freeman, Johnson, Llao et al. 1993



1.25 mg As/kg 70 Freeman, Johnson, Killinger et al. 1993

Cynomolgus monkey


1.25 mg As/kg 43.6 Freeman, Johnson, Killinger et al. 1993

Cynomolgus monkey


0.62 mg As/kg 10-30 Freeman et al. 1995

Dust (same location) 0.26 10-30

Immature swine 144 0.01 - 0.31 mg As/kg Soil

0.04 - 0.24 mg As/kg

78 US EPA 1996

slag 0.61 - 1.52 mg As/kg

42

Australian data

Wistar rat 96 Sodium arsenite: 5 mg As3+/kg

Soils (from 16 cattle dip 5 mg As/kg 8.14.0 Ng & Moore 1996

Calcium arsenite: 5 mg As3+/kg

sites, NSW, Australia) 14.47.1

Sodium arsenate: 5 mg As5+/kg

6032.4

Table 6 (cont.) Relative bioavailability (comparison to oral gavage) of arsenic from soil in laboratory animals by comparison of blood concentrations otherwise specified.

Species Duration Oral gavage Soil or sample Soil dose Bioavailability Reference (hours) (%, mean±SD)

Wistar rat 96 Sodium arsenite: 0.5 mg As3+/kg

Soils (9 samples from one CCA contaminated site)

0.5 mg As/kg 13.04.5 Ng & Moore 1996

Calcium arsenite: 0.5 mg As3+/kg

32.211.2

Sodium arsenate: 0.5 mg As5+/kg

3813.2

Sprague Dawley rat

168

Sodium arsenite: 5 mg As3+/kg Sodium arsenate: 5 mg As5+/kg


0.5 mg As/kg

10.28 - 203.09 (urine)

Bruce 2004

Sprague Dawley rat

240 Sodium arsenite: 5 mg As3+/kg Sodium arsenate: 5 mg As5+/kg

4 categories of mine wastes 0.5 mg As/kg 1.6-8.9% Diacomanolis, Ng & Noller 2007

Swine (20-25 kg b.w.)

26 Sodium arsenate: 0.1 mg As5+/kg

Herbicide/pesticide impacted, mine site and gossan soils (n = 12)

0.03-0.4 mg As/kg 6.9 5.0% - 74.7 11.2%

Juhasz et al. 2007b

Swine (20-25 kg b.w.)

26

Sodium arsenate: 0.1 mg As5+/kg

Red Ferrosol soilA Brown Chromosol soilB

0.2-0.25 mg As/kg

0.2-0.5 mg As/kg

5333.9 - 24.19.5g

97.21.9 - 92.95.3

110.6 - 697.24

Juhasz et al. 2008

Cattle 266 days (repeated

doses)

Sodium arsenite: 0.5 mg As3+/kg Sodium arsenate: 0.5 mg As5+/kg


0.5 mg As/kg (20.3 - 2919)Liver

Bruce et al. 2003; Bruce 2004

Where in Juhasz et al. 2008: A= Red Ferrosol soil contained 47.5% sand, 24% clay, 20% silt, spiked with sodium arsenate to achieve final arsenic concentration of 1000 mg/kg. B= Brown Chromosol soil contained 87% sand, 7.8% clay, 2.7% silt, spiked with sodium arsenate to achieve final arsenic concentration of 1000 mg/kg. Both A and B were aged for 12 months. The bioavailability data showed the decrease of bioavailability from 0 months to 12 months after aging. Data from 3 and 6 months are not shown here.

Table 7. Metabolism and urinary excretion of inorganic and organic arsenicals in human following experimental administration.

Form Adult Dose and frequency Time % Dose in % of Total urinary metabolites Reference

subjects duration urine As5+ As3+ MMA DMA

Arsenic acid 6 0.01 µg 5 days 57.9 27.2 20.6 51.0 Tam et al. 1979

Arsenic acid 6 0.06 ng 7 days 62.3 Not Determined Pomroy et al. 1980

Arsenic trioxide 1 700 µg 3 days 68.2 7.9 31.7 28.2 32.2 Yamauchi & Yamamura 1979

Sodium

Arsenite

3 500 µg 4 days 45.1 25 21.3 53.7 Buchet, Lauwerys & Roels 1981a

Sodium metaarsenite

1 125 µg x 5 days 14 days 54 16 34 50 Yamauchi & Yamamura 1979

1 250 µg x 5 days 14 days 73 7 20 73

1 500 µg x 5 days 14 days 74 19 21 60

1 1000 µg x 5 days 14 days 64 26 32 42

Sodium mono-methyl arsonate

4 500 µg 4 days 78.3 ND ND 87.4 12.6 Buchet, Lauwerys & Roels 1981a

Sodium dimethyl arsinate

4 500 µg 4 days 75.1 ND ND ND 100 Buchet, Lauwerys & Roels 1981b


Lead

The absorption coefficient of lead is from 0.01–0.14 (mean 0.10) (Owen 1990). A key study on absorption and retention of lead by infants was reported in 1978 (Ziegler et al. 1978). The authors conducted 89 metabolic balance studies with 12 normal infants aged 14–746 days over a 72 hour duration. Subjects were fed milk or formula and commercially prepared strained foods. The estimated Pb intakes were 0.83–22.61 (9.44 ± 5.27) µg/kg/day. Faecal and urinary excretions of lead were measured, and net absorption and retention were calculated. Out of all the balance studies – 61 studies with intakes of greater than 5 µg/kg/day – the averaged net absorption was 41.5% of lead intake and net retention was 31.7% of intake. Both absorption and retention were inversely correlated with intake of calcium, suggesting dietary calcium reduces lead absorption. Seven of 28 studies with lead less than 5 µg/kg/day and in only 3 of 61 studies at higher intakes returned a higher faecal excretion than the intake. Intake less than 5 µg/kg/day were more likely to be unreliable. The authors speculated infants in this treatment group not under special study conditions and might have received higher intakes than this value. Further, this study didn’t account for non-dietary intakes and excretion other than those with urine and faeces. Excretions via other routes (e.g. sweat) are considered to be low. However, intakes from other sources could be significant, including inhalation of fine dust and ingestion of dust with hand-to-mouth activity. If the actual intakes were higher than those measured by the authors then the real absorption could be lower than the reported value here.

In an earlier report from 11 balance studies carried out with eight subjects aged three months to eight years with intakes ranging about 5–17 µg/kg/day (mean of 10.6 µg/kg/day), the average absorption was 53% of the intake and average retention was 18% of the intake (Alexander, Clayton & Delves 1974). Lead radioisotope studies reported about 10% absorption of lead by human adults (Hursh & Suomela 1968; Rabinowitz, Wetherill & Kopple 1976).

Measurements of lead relative bioavailability are often low, with ranges from as low as 6% for New Zealand White rabbits fed contaminated soils (Davis, Ruby & Bergstrom 1992), and 10.7% observed in monkeys, also fed contaminated soils (Roberts et al. 2002). The US Environmental Protection Agency (US EPA) views only in vivo studies conducted using the juvenile swine as the appropriate animal to undertake lead studies to estimate lead bioavailability in young children. The US EPA assumed 30% of ingested lead in soil will be bioavailable (absolute bioavailability) when assessing lead bioavailability of contaminated sites (US EPA 1994). However, the recent bioavailability studies using animal models mentioned above have demonstrated that the BA of lead from some soils and mine waste materials may indeed be considerably lower or higher as confirmed in later studies.

Multiple bioavailability studies done by US EPA Region 8 on 18 soil and soil-like samples and one NIST paint material using the juvenile swine model was reported recently (US EPA 2007b). Although the gastroinstestinal tract of immature swine is believed to be similar to that of young humans, the absolute bioavailability of lead acetate was found to be 10, 14,16, and 19% as measured using blood (AUC), femur, liver and kidney respectively (average 15 ± 4%). Lead absorption in juvenile swine is apparently lower than for young human (42% to 53% – see below). RBA varied widely between different test materials with the lowest RBA of about 1% in California Gulch Oregon Gulch tailings and the highest of 105% in California Gulch Fe/Mn PbO. Galena-enrich soil (PbS) had the lowest RBA range of -1% to 4%. The report


concluded that the wide variability highlights the importance of obtaining and applying reliable RBA data in order to help improve risk assessment for lead exposure. Although available data are not yet sufficient to establish reliable quantitative estimates of RBA for each of the different mineral phases of lead in the test materials, the report provides a tentative ranking order of the phases into three semi-quantitative categories (low, medium, or high RBA) as shown in Table 10.

Table 10. Relative bioavailability (RBA) ranking order found in various soil and soil-like materials obtained from juvenile swine dosing experiments (from US EPA 2007b).

Low bioavailability

RBA< 25%

Medium bioavailability

RBA = 25–75%

High bioavailability

RBA > 75%

Fe(M) sulfate Lead phosphate Cerussite (lead carbonate)

Anglesite Lead oxide Mn(M) oxide

Galena

Pb(M) oxide

Fe(M) oxide

In a recent study by Juhasz, Weber et al. (2009), in vivo swine experiments were performed to determine the bioavailability of Pb in (Australian) contaminated soils. Lead doses were administered under fasting conditions to represent a worst-case scenario for Pb exposure. Feed was administered to swine two hours after Pb dosage to ensure the lowest gastric conditions at the time of Pb exposure. The relative bioavailability of Pb in contaminated soils was determined by comparing the area under the blood-Pb concentration time curve for orally administered contaminated soil and a Pb acetate reference dose. Mean relative Pb bioavailability for contaminated soils ranged from 10.1 ± 8.7% to 19.1 ± 14.9%. Variability in relative Pb bioavailability was observed between triplicate soil analysis due to physiological intra-species differences including genetic factors, disparity in stomach clearance times, stomach pH and rate of Pb absorption. Previous Pb bioavailability assays utilising swine have reported relative Pb bioavailability in contaminated soils to range from 1% to 87% (Marschner et al. 2006; Schroder et al. 2004), although in the study of Marschner et al. (2006), relative Pb bioavailability (17–63%) was calculated following the determination of Pb in selected organs, bone and urinary excretions but not blood analysis. Examples of bioavailability data obtained from various animal models are summarised in Tables 8 and 9.

Based on available literature data on lead in humans, the Integrated Exposure Uptake Biokinetic Model for Lead (IEUBM) used by US EPA estimates that the absolute bioavailability of lead from water and diet is usually about 50% in children (US EPA 1994). Thus, when a reliable site-specific RBA value for soil is available, it may be used to estimate a site-specific absolute bioavailability in that soil using Equation [5]:

ABAsoil = 50% x RBAsoil Equation [5]


Key findings

A range of in vivo animal models have been utilised to assess contaminant bioavailability in soils.

In vivo organic bioavailability is poorly quantified for all organic compounds of concern.

In vivo inorganic bioavailability studies have mostly focused on the bioavailability of arsenic and lead in contaminated soils.

Bioavailability of arsenic and lead in contaminated soils may range from <10 to >90%.

The US EPA recommends that the juvenile swine model be used to assess bioavailability of arsenic or lead in contaminated soils.

Table 8. Absolute (comparison to intravenous route) oral bioavailability of lead in laboratory animals.

Species Duration i.v. dose Soil or sample Soil dose Bioavailability Method Reference

(hours) (%, mean±SD)

International data

Humans

72

0.02, 0.20, 2.0 mg/kg b.w. for 29 days

Food

0.83-22.61 µg Pb/kg/day

(Food)

41.5

Urinary & faecal excretion

Ziegler et al. 1978*

Sprague Dawley rat

29-30 days

Lead acetate

0.08-26.2 mg Pb/kg/day (lead

acetate spiked feed)

0.12-23.8 mg Pb/kg day (mine waste)

15±8.1(blood) 7.4±4.1(bone) 11±7.1(liver)

2.7±1.5(blood) 0.4±0.16(bone)

0.55±0.68 (liver)

Blood, bone or liver

Freeman et al. 1992

Juvenile swine

360

100 µg Pb/kg

Lead acetate solution

0

10-19

AUC for blood, necropsy Pb in liver, kidney or bone

US EPA 2007b

Juvenile swine

15 days

100 µg Pb/kg

Lead acetate solution

75, 225 or 675 µg Pb/kg/day (berm or

residential soil)

28 (blood) & 43 (liver) for berm

soil; 29 (blood) & 37 (liver) for

residential soil

AUC for blood or necropsy Pb in liver

Casteel et al. 1997

Australian data

Sprague Dawley rat

240

0.5 mg Pb/kg Lead acetate

solution

0.01-2.7 mg Pb/kg (Various mine wastes)

0.6-1.4 AUC for urinary excretion

Diacomanolis, Ng & Noller 2007

Sprague Dawley rat

168 1 mg Pb/kg Lead acetate solution

10 mg Pb/kg (Zinc concentrate)

1±2.15 AUC for urinary excretion

Bruce 2004

Cattle

240 0.5 mg Pb/kg Lead acetate solution

5 mg Pb/kg (in Zinc concentrate)

3±0.51 AUC for blood Bruce 2004

*This food study is included as cross-reference to soil exposure

Table 9. Relative bioavailability (comparison to oral gavage) of lead from soil in laboratory animals by comparison of blood concentrations otherwise specified.

Species Duration Oral gavage Soil or sample Soil dose Bioavailability Reference

(hours) (%, mean±SD)

International data

Juvenile swine

360

0.025-0.225 mg Pb/kg-day (lead acetate)

18 soil of various mineral phases and 1 NIST paint

0.075-0.625 mg Pb/kg-day

0-105

US EPA 2007b

Juvenile swine 15 days

0, 75, 225 µg Pb/kg-day

Berm or residential soil

75, 225 or 675 µg Pb/kg-day (berm or residential soil)

56-58 (blood), 74-86 (liver),

68-74 (kidney), 68-72 (bone)

Casteel et al. 1997

Sprague Dawley rat

30 days

0.076-25.7 mg Pb/kg-day

Mine waste soil

0.119-23.2 mg Pb/kg-day

12.1-26.8 (blood), 4.8-13.3 (bone), 0.6-13.6 (liver)

Freeman et al. 1992

Australian data

Sprague Dawley rat

168 10 mg Pb/kg (lead acetate)

Zinc concentrate 10 mg Pb/kg 38±6.8 Bruce 2004

Juvenile swine 5 days 0.025-0.225 mg Pb/kg (lead acetate)

5 soils from a domestic incinerator site and residential land developed on quarry fill material

0.225-0.53 mg Pb/kg 7.4 ± 4.2 – 19.1 ± 1.9

Juhasz, Weber et al. 2009

Cattle 240

5 mg/kg (lead acetate)

Zinc concentrate

5 mg Pb/kg

80±13.2

Bruce 2004

Cattle 266 days 0.5 mg Pb/kg-d (lead acetate)

Zinc concentrate 0.5 mg Pb/kg-d 48±3.2Liver Bruce 2004


3.4 Factors influencing contaminant bioavailability

The bioavailability of ingested contaminants will vary depending on the matrix in which it is ingested (e.g. food, water, beverages, soil). Indeed, solubility of the arsenical compound itself, the presence of other food constituents, and nutrients in the gastrointestinal tract may all influence the bioavailability of arsenic. There are a range of other factors which may influence bioavailability. These include pH of the gastrointestinal conditions of the receptor, pH of the contaminated material, and the chemical and physical properties of the material. Furthermore, it is generally true that smaller particle size material has a greater bioavailability compared to larger size particles of the same material. Since particle size of <250 µm is the most likely fraction ingested via hand-to-mouth activities by children (US EPA 2009), it is important that bioavailability be determined using this particle size fraction. In brief, both environmental and physiological conditions can influence the bioavailability. Owen (1990) reported absorption coefficients for 39 chemicals via oral and inhalation routes of exposure which could be indicative for bioavailability. Although there are differences in physiological features and metabolisms in animals and humans, bioavailability considering only the absorption fraction via the GI tract can be determined using various animal models provided a proper positive control group of animals is included in the same study. Ng and Moore (1996) proposed the use of a single rat blood for relative bioavailability (they also referred to comparative bioavailability in the literature). The advantage of rats is their greater capacity to bind arsenic in the red blood cells compared to other animal species, including humans. For a same dose, the rat blood arsenic is 60-fold greater than that of a guinea pig. They are of the view that rats are sufficiently large enough compared to mice, and more cost-effective than larger mammals such as dogs, pigs and monkeys or perhaps the ultimate human model. The metabolism (methylation of pathway) of arsenic is different amongst various animal species. For example, chimpanzees and marmoset monkeys are unable to methylate arsenic (Vahter et al. 1995; Zakharyan et al. 1996). It has been often cited that there is no perfect animal model to replicate arsenic metabolism in humans. However, for bioavailability studies, given the way in which the relative bioavailability of a given metal/metalloid is operationally defined and measured, differences in one’s metabolism, even if they exist, would not be a confounding factor (Roberts et al. 2002). Hence, many animal species have been employed for bioavailability studies.

Although in vivo studies utilising animal models are an appropriate method for determining contaminant bioavailability in soil for inclusion in human health exposure assessment, the time required for in vivo studies, expense of animal trials and ethical issues preclude their use as routine bioavailability assessment tools. As a result, rapid, inexpensive in vitro methods simulating gastrointestinal conditions in the human stomach have been developed as surrogate bioavailability assays. These assays determine contaminant concentrations that are solubilised following gastrointestinal extraction, and are therefore available for absorption into systemic circulation. This fraction is referred to as the ‘bioaccessible fraction’ and is specifically used when in vitro assessment models are used. The following sections detail the development of in vitro methods, key assay parameters and examples where the bioaccessibility of organic and inorganic contaminants in soils have been determined.


Key findings

Soil particle size, soil pH and gastrointestinal pH all influence contaminant bioavailability.

The US EPA recommends that the <250 µm soil particle size is used to determine contaminant bioavailability in contaminated soils.


4. In vitro methods for assessing contaminant bioaccessibility

In order to determine the bioaccessibility of contaminants in soil, a variety of in vitro gastrointestinal extraction methods have been developed (see Tables 11–13). These methodologies attempt to simulate processes that occur in the human body that lead to the release of contaminants from the soil matrix. As the human digestive system is an extremely complex system, these methodologies do not attempt to replicate conditions found in various compartments, but mimic key processes.

The forerunners of contaminant bioaccessibility were developed for nutritional studies in order to estimate the availability of iron in food (Schricker et al. 1981). Subsequently, in vitro assays were developed to determine the bioaccessibility of inorganic contaminants (e.g. arsenic and lead) in soil in order to estimate potential exposure from the incidental ingestion of contaminated soil. While a considerable amount of research has been performed on the assessment of arsenic and lead bioaccessibility in contaminated soil, limited research has been performed on organic contaminants although research efforts have been increased in the past few years. When developing in vitro assays for assessing contaminant bioaccessibility, parameters should be representative of the physiology of children, the group most at risk from exposure to environmental contaminants.

As seen in Tables 11–13, in vitro methodologies may consist of up to three phases including esophageal, gastric and intestinal. However, due to the short residence time that material spends in the mouth (approximately two minutes), this phase may not contribute significantly to contaminant release and therefore may be optional. The gastric and intestinal phases are present in most in vitro methodologies, although compartment parameters may vary between methods. Described below are key variables that need to be considered when designing and developing in vitro methodologies for the assessment of contaminant bioaccessibility.

There are several other methodologies that are sometimes utilised to assess contaminant bioaccessibility. These include methodologies to assess the potential leaching characteristics of waste material (Toxicity Characteristic Leaching Procedure (TCLP) and the Australian Standard Leaching Procedure). These methods have been developed for estimating the potential leaching characteristic of waste material and should not be utilised to estimate contaminant bioaccessibility. Similarly, strong acids (i.e. 0.1M HCl) have often been utilised to estimate potential metal bioaccessibility in contaminated soils. This methodology has not been correlated with in vivo bioavailability data and has limited applicability for in vitro estimations.


Table 11. Composition and in vitro parameters for commonly utilised in vitro bioaccessibility assays (SBRC, IVG, PBET and DIN) for inorganic contaminants.

In vitro parameters

Method/phase Composition (g l-1) pH Soil:solution ratio

Extraction time (hr)

SBRC

Gastric

Intestinal

30.03 g glycine

1.75 g bile, 0.5 g pancreatin

1.5

7.0

1:100

1:100

1

4

IVG

Gastric

Intestinal

10 g pepsin, 8.77 g NaCl


1.8

5.5

1:150

1:150

1

1

PBET

Gastric

Intestinal

1.25 g pepsin, 0.5 g sodium malate, 0.5 g sodium citrate, 420 µl lactic acid, 500 µl acetic acid


2.5

7.0

1:100

1:100

1

4

DIN

Gastric

Intestinal

1 g pepsin, 3 g mucin, 2.9 g NaCl, 0.7 g KCl, 0.27 g KH2PO4

9.0 g bile, 9.0 g pancreatin, 0.3 g trypsin, 0.3 g urea, 0.3 g KCl, 0.5 g CaCl2, 0.2 g MgCl2

2.0

7.5

1:50

1:100

2

6

Table 12. Methods for the assessment of contaminant bioaccessibility which include esophageal, gastric and intestinal phases.

Phase Parameter Bioaccessibility method

A B C

Reference TIM (Minekus et al. 1995)a BARGE; Oomen et al. 2000, 2001b Wittsiepe et al. 2001c

Target Contaminant - PCBs, Lindane PCDD/F

Esophageal Composition (per litre) Saliva solution – details not available 1.79 g KCl, 0.4 g KSCN, 1.78 g NaH2PO4, 1.14 g Na2PO4, 0.6 g NaCl, 0.14 g NaOH, 0.4 g urea, 290 mg α-amylase, 30 mg uric acid, 100 mg mucin

0.9 g KCl, 0.2 g KSCN, 0.57 g Na2SO4, 0.29 g NaCl, 0.89 g NaH2PO4, 0.145 g α-amylase, 5 mg mucin, 0.2 g urea, 15 mg uric acid

Soil:solution ratio 1:5 1:15 1:10

pH 5.0 6.5 -

Mixing Peristaltic 60 rpm 200 rpm

Residence time 5 min 5 min 0.5 hr

Gastric Composition (per litre) Full details not available. Gastric juice containing lipase and pepsin.

Initial gastric pH decreasing from 5.0 to 3.5, 2.5 and 2.0 after 30, 60 and 90 minutes respectively.

5.5 g NaCl, 0.54 g NaH2PO4, 1.64 g KCl, 0.8 g CaCl2

. 2H2O, 0.62 g NH4Cl, 1.3 g glucose, 0.04 g glucuronic acid, 0.17 g urea, 0.66 g glucoseamine hydrochloride,2 g bovine serum albumin (BSA), 2 g porcine pepsin, 6 g mucin

0.2 g CaCl2, 0.82 g KCl, 2.75 g NaCl, 0.27 g NaH2PO4, 0.31 g NH4Cl, 0.33 g glucosamine hydrochloride, 0.65 g glucose, 0.02 g glucuronic acid, 1.5 g mucin, 0.05 g n-acetylneuraminic acid, 1.0 g porcine pepsin, 1.0 g BSA, 0.08 g urea, 58.3 g whole milk powder (optional)

Soil:solution ratio 1:25 1:37.5 1:30

pHd 2.0 (HCl) 1.2 (HCl) 2.0 (HCl)

Mixing peristaltic 60 rpm 200 rpm

Residence time 1.5 hr 2 hr 3 hr

Table 12 (cont.) Methods for the assessment of contaminant bioaccessibility which include esophageal, gastric and intestinal phases.


A B C

Intestinal Composition (per litre) Full details not available. Intestinal juice containing porcine bile and pancreatin.

3 sections, duodenum (pH 6.5), jejunum (pH 6.8) and ileum (pH 7.2)

Duodenal juicec: 14.02 g NaCl, 6.78 g NaHCO3, 0.16 g KH2PO4 1.13 g KCl, 0.1 g MgCl2, 0.2 g urea, 0.4 g CaCl2.2H2O, 2.0 g BSA,6 g porcine pancreatin, 1.0 g lipase

Bile solution: 10.52 g NaCl, 11.57 g NaHCO3, 0.75 g KCl, 0.5 g urea, 0.44 g CaCl2.2H2O, 3.6 g BSA, 12 g chicken bile

Duodenal juiced: 0.2 g CaCl2, 0.56 g KCl, 0.08 g KH2PO4, 0.05 g MgCl2, 7.01 g NaCl, 1.8 g NaHCO3, 0.5 g lipase, 3.0 g porcine pancreatin, 1.0 g BSA, 5 mg stearic acid, 0.1 g urea

Bile solution: 0.22 g CaCl2, 0.37 g KCl, 5.25 g NaCl, 4.2 g NaHCO3, 3.0 g chicken bile, 1.8 g BSA, 0.25 g urea

Soil:solution ratio - 1:97.5 1:60

pHb 7.2 (NaHCO3) >5.5 7.5

Mixing peristaltic 60 rpm 200 rpm

Residence time 6 hr 2 hr 3 hr

a The TIM bioaccessibility method developed by Minekus et al. 1995 is a dynamic system. b The in vitro method consisted of 9 ml of esophageal phase, 13.5 ml of gastric juice, 27 ml of duodenal juice and 9 ml of bile solution. c The in vitro method consisted of 20 ml of esophageal phase, 40 ml of gastric juice, 40 ml of duodenal juice and 20 ml of bile solution. d Information in parenthesis indicates the mode of pH adjustment.

Table 13. Methods for the assessment of contaminant bioaccessibility which include gastric and intestinal phases.


D E F

Reference Ruby et al. 2002 Hack & Selenka 1996 Wittsiepe et al. 2001

Target POP Dioxins, furans PAHs, PCBs PCDD/F

Gastric Composition (per litre) 15 g glycine, 8.8 g NaCl, 1.0 g porcine pepsin, 5.0 g BSA, 2.5 g mucin, 6 ml oleic acid

83.3 g porcine pepsin,58.3 g whole milk powder, 1.8% w/v NaCl solution, 2.9 g mucin

95 mg porcine pepsin, 3.3 g mucin, 58.3 g whole milk powder, or 15.2 g grape seed oil


pHa 1.5 (HCl) 2.0 (HCl) 2.0 (HCl)

Mixing 30 rpm 200 rpm 200 rpm


Intestinal Composition (per litre) 600 mg porcine pancreatin, 4 g bovine bile

2.9 g porcine pancreatin, 2.9 g bovine bile, 83.3 mg trypsin

2.9 g chicken bile, 2.9 g porcine pancreatin, 88.3 mg trypsin


pHa 7.2 ± 0.2 (NaOH) 7.0 (NaHCO3) 7.0 (NaHCO3)



aInformation in parenthesis indicates method of pH adjustment

Table 13 (cont.) Methods for the assessment of contaminant bioaccessibility which include gastric and intestinal phases.


G H I

Reference Holman 2000a DIN 2000 Van de Wiele, Verstraete & Siciliano 2004 (SHIME)

Target POP PAHs Organics PAHs

Gastric Composition (per litre) - 1 g porcine pepsin, 3 g mucin, 2.9 g NaCl, 0.7 g KCl, 0.27 g KH2PO4

10.0 g KHCO3, 5.7 g NaCl, 10 mg porcine pepsin

Soil:solution ratio - 1:50 1.2-1.10-1.40

pHb - 2.0 (HCl) 1.5 (HCl)

Mixing - 200 rpm 150 rpm

Residence time - 2 hr 2 hr

Intestinal Composition (per litre) 8.8 g NaCl Mixed bile saltsc: 11.5 mg-2.3 g sodium glycocholate, 11.1 mg-2.2 g sodium glycochenodeoxycholate, 7.5 mg-1.5 g sodium glycodeoxycholate, 0.3-60 mg sodium glycolithocholate, 1.7-340 mg disodium glycolithocholate sulphate, 6.5 mg-1.3 g sodium taurocholate, 6 mg-1.2 g taurochenodeoxycholate, 4.2-840 mg sodium taurodeoxycholate, 0.2-40 mg sodium taurolithocholate, 0.6-120 mg disodium taurolithocholate sulphate Mixed intestinal lipidsd: 0.7 g cholesterol monohydrate, 5.5 g oleic acid, 1.4 g monoolein, 0.5 g diolein, 3.1 g lecithin

9.0 g bovine bile, 9.0 g porcine pancreatin, 0.3 g trypsin, 0.3 g urea, 0.3 g KCl, 0.5 g CaCl2, 0.2 g MgCl2

0.9 g porcine pancreatin (Duodenum model)e


pHb 6.5-7.3 7.5 (NaHCO3) 6.3 (NaHCO3)


Residence time 4 hr 6 hr 5 hr a Information in parenthesis indicates method of pH adjustment. b Information in parenthesis indicates method of pH adjustment. c Mixed bile salts were added at 0.1, 5.0 and 20 mM to represent fasting, fasting with gall bladder emptying and fat digestion states. d Mixed intestinal lipids were only added during the fat digestion state. . eThe duodenum digest suspension (100 ml) was supplemented with 100 ml of SHIME suspension (Simon and Gorbach, 1995). SHIME contained microbiota representative of the composition and concentration in the human colon. Lactobacilli, Bifidobacteria, Enterococci, Fungi, Staphylococci and Clostridia were included with total anaerobic and aerobic microorganisms at approximately 8.4 and 7.8 log colony forming units (CFU) ml-1. The colon digest was incubated at 37°C and stirred at 150 rpm for 18 h.


4.1 Chyme composition

The composition of gastric and intestinal phase range from simple systems containing few constituents (e.g. Rotard et al. 1995), to highly complex assays that contain numerous constituents including simulated intestinal microbial communities (e.g. SHIME – Molly, Van de Woestyne & Verstraete 1993). Pepsin is a base constituent of the in vitro gastric phase. It is a digestive protease which functions to break down proteins into peptides. It has also been suggested that pepsin may decrease the surface tension of chyme (Tang et al. 2006), thereby increasing the mobilisation potential and solubility of POPs (Charman et al. 1997). Alternatively, glycine is utilised as the main constituent of the gastric phase instead of pepsin. Mucins are also a commonly added constituent to the gastric phase of in vitro bioaccessibility assays. Mucins are a group of large glycosylated proteins that are secreted on the mucosal surface. Mucins are added to in vitro gastric phases to increase contaminant mobilisation in the digestive tract (Hack & Selenka 1996). Oomen et al. (2000), Wittsiepe et al. (2001) and Ruby et al. (2002) included protein in gastric phase solutions in order to enhance organic contaminant solubilisation from the soil matrix. Similarly, lipids (e.g. oleic acid) may be included in gastric phase solutions to increase organic contaminant solubilisation. The detergency effect of bile ingested lipids (e.g. triglycerides) leads to formation of mixed bile-lipid micelles. Soil-borne organics may be mobilised into these micelles which are potentially available for absorption (Hack & Selenka 1996; Holman 2000; Oomen 2000). Lipids may be obtained from milk derivatives (Wittsiepe et al. 2001) or oleic acid, a fatty acid abundant in animals and vegetables (Oomen 2000; Ruby et al. 2002).

Bile and pancreatin are central components of the intestine phase of in vitro bioaccessibility assays. The influence of bile on the mobilization of organic contaminants has been well established. Bile is secreted from the gall bladder into the duodenum during digestion in order to facilitate the breakdown of lipids. Consisting of cholesterol, phospholipids, bile pigments, bile salts and bicarbonate (Dean & Ma 2007), bile forms aggregates with monoglycerides and fatty acids (bile-fat micelles) as a result of lipase. As a result of micelle formation, contaminant solubility may be enhanced due to the incorporation of these compounds into high surface area amphipathic molecules (Gorelick & Jamieson 1994; Walsh 1994). While human bile may not be used in in vitro assays due to ethical issues (Alvaro et al. 1986; Wildgrube et al. 1986), a variety of animal bile has been used. Due to the similarity of bile salt percentage to human bile, bovine and porcine bile are considered most suitable for in vitro applications (Oomen et al. 2004). Different forms of bile have been used in various models such as purified uniform bile salts and animal origin bile (Friedman & Nylund 1980). Previous studies identified that mixed micelles, an important parameter for absorption, may not be formed when using purified uniform bile salts. As a result, some in vitro assays suggest the use of original freeze-dried bile obtained from animals (Hack & Selenka 1996; Minekus et al. 1995; Oomen et al. 2002; Rotard et al. 1995; Ruby et al. 1996). While chicken bile was used in some early in vitro assays (Rotard et al. 1995), its use has been discouraged due to its varying solubilising effect compared to other bile products. For example, lead bioaccessibility was 3–5.5 times greater when chicken bile was used in the intestinal phase compared to bovine and porcine bile (Oomen et al. 2003, 2004). Bovine or porcine bile is preferred to chicken bile, because chicken bile may lead to an irregular and unaccountable bioaccessibility pattern and the composition of chicken bile is significantly different from the composition of human bile. In addition, the quantity of


bile added to the intestinal phase may influence organic contaminant bioaccessibility. For example, Pu et al. (2004) observed a 45-fold increase in phenanthrene bioaccessibility when bile was included in the intestinal phase medium.

Pancreatin is also included in the simulated intestinal fluid. Pancreatin is a mixture of digestive enzymes (amylase, lipase, trypsin and protease) which hydrolyse proteins to oligopeptides (trypsin), hydrolyse starch to oligosaccharides (amylase) and hydrolyse triglycerides to fatty acids and glycerol (lipase). In some cases, trypsin is added to intestinal fluid to increase the capacity for protein hydrolysis (Hack & Selenka 1996; Wittsiepe et al. 2001).

4.2 Gastric and intestinal phase pH

The pH of the human gastric system can vary significantly depending on whether the subject is under fasting or fed states. Ruby et al. (1996) reported that the mean fasting pH value for young children ranged from 1.7 to 1.8, with a range of 1 to 4. Following ingestion of food, the stomach pH will rise to > 4 and will return to basal levels two hours following stomach emptying. Developers of in vitro assays have generally chosen a gastric phase pH value which represents a worst case scenario (fasted state) for young children. Low pH stomach values are particularly prudent for the assessment of metal bioaccessibility as the pH will drive the dissolution of metals and mineral phases, thereby controlling the fraction that is potentially available for uptake. As seen in Tables 11–13, the majority of bioaccessibility assays employ a gastric pH of 1–2. However, Molly, Van de Woestyne and Verstraete (1993) utilised a gastric phase pH of 5.2 in the SHIME method which was meant to represent an infant’s fed state stomach pH. The pH in the small intestine varies from 4–4.5 in the initial part of the small intestine to 7.5 in the ileum (Daugherty & Mrsny 1999; Guyton 1991; Johnson 2001; Sips et al. 2001). Most in vitro assays adjust the pH of the intestinal phase to near neutral, (6.5–7.5), however Oomen et al. (2000) employed a pH value of 5.5.

4.3 Gastric and intestinal phase residence time

Nutrition studies have demonstrated that complete emptying of the stomach may occur after 1–2 hours, while 3–5 hours is required for chyme to pass from the start of the small intestine to the start of the large intestine. Most bioaccessibility assays reflect the timeframes with gastric extraction times ranging from 1–3 hours and intestinal extraction times of 2–6 hours. It has been proposed that small variations in residence time do not have a significant impact on bioaccessibility (Daugherty & Mrsny 1999; Degan & Philips 1996; Guyton 1991; Johnson 2001).

4.4 Soil:solution ratio

Soil:solution ratio is one parameter that exhibits the greatest variability between in vitro assays. Ratios of 1:2 (g ml-1) up to 1:5000 (g ml-1) have been used by a variety of researchers. Ruby et al. (1992, 1996) proposed that assays which utilise soil:solution ratios of 1:5 to 1:25 may underestimate the bioaccessible fraction due to solubility issues at these ratios as a result of diffusion limited dissolution kinetics. However, Hamel, Buckley and Lioy (1998) demonstrated that little differences in metal


bioaccessibility values resulted from soil:solution ratios of 1:100 to 1:5000 (g ml-1). While this may be the case for metal contaminants, the influence of soil:solution ratio on organic contaminant bioaccessibility has received little attention. As a result, a soil:solution ratio of 1:100 has been selected for most in vitro assays for the assessment of metal bioaccessibility, as this ratio would not influence the test results. In addition, due to insufficient data for fasting children to support any soil:solution ratio, 1:100 was selected arbitrarily to represent a fasting child (Ruby et al. 1996).

4.5 Amendments to gastric and intestinal phases

Another variable between in vitro assays is the inclusion of food additives. Food may be added to in vitro assays for comparison of contaminant bioaccessibility between fed and fasted states (Ruby et al. 1996). In addition, the choice and amount of food added to in vitro assays may have a significant effect on bioaccessibility measurements depending on the fat and protein content of the additive. The inclusion of food additives have been shown to increase the bioaccessibility of organic contaminants. Hack and Selenka (1996) included lyophilised milk as part of the in vitro assay which increased PAH and PCB mobilisation from 5–40% to 40–85%.

4.6 Other parameters

Irrespective of methodology, bioaccessibility assays are conducted at 37°C, as this is the physiological temperature of the human body. In addition, agitation is required during bioaccessibility assessment, either in the form of shaking, stirring, end-over-end rotation, inert gas movement or peristaltic movement, in order to mimic gastrointestinal turbulence. Most in vitro methods are performed under batch condition, i.e. the simulated digestive contents is reacted and analysed in one vessel and compartments are added to the system step by step during process. Alternatively, in order to better represent the human digestive system and its motion, a ‘flow through’ model was developed by Molly, Van de Woestyne and Verstraete (1993) and Minekus et al. (1995).

Key findings

A considerable diversity of factors may be considered when determining contaminant bioaccessibility in contaminated soils.


5. Bioaccessibility assessment for contaminated soils

As outlined in Table 14, the bioaccessibility of organic contaminants in soil varies significantly (0.1–89%) depending on the physico-chemical properties of the compound and soil, soil-contaminant residence time and the in vitro methodology used. While limited information is available in the literature on organic contaminant bioaccessibility, some conclusions can be drawn from the aforementioned studies. Interactions between the soil matrix and organic contaminant significantly influence the bioaccessibility of these chemicals in the environment (Scott & Dean 2005). Sequestration via association, retention or sorption mechanisms with solid matrices in soils and sediments are important processes that affect contaminant retention and bioaccessibility. Organic contaminants may bind to soil or sediment components through hydrophobic partitioning or via chemical or physical bond formation. Non-polar, hydrophobic compounds are usually retained on the organic components of the soil, including condensed humic material or soot particles (Alexander, 2000; Chiou 1998; Ellickson et al. 2001; Juhasz, Mallavarapu & Naidu 2000; Luthy et al. 1997; Schlebaum et al. 1998; Semple, Morriss & Paton 2003; Xing & Pignatello 1997). As a result of these physico-chemical processes, the bioaccessibility of organic contaminants may be reduced with increasing soil-contaminant residence time (Alexander 1995; ATSDR 1995; Linz & Nakles 1997). This is highlighted by the differences in bioaccessibility values derived from contaminant-spiked versus contaminated soils (see Table 14).

With increasing contaminant-soil residence time, diffusional and diagenetic processes affect the retention of organic compounds in a process termed ‘ageing’. Organic contaminants may be incorporated into natural organic matter, diffuse into nanopores or absorb into organic matter. The concentration and composition of soil organic matter will significantly influence the sequestration of hydrophobic organic contaminants. In phenanthrene spiked soil studies, Pu et al. (2004) determined that soils with higher organic carbon contents tended to result in low phenanthrene bioaccessibility values. In addition, Pu et al. (2004) suggested that the clay composition of the soil could also play an important role in contaminant bioaccessibility due to enhanced sorption of hydrocarbon compounds with increasing soil-clay contents. In addition, physical, chemical and biological processes may cause changes in the soil sorptive matrix which influences the sequestration of organic contaminants.

Tables 15 and 16 outline the variability in arsenic and lead bioaccessibility when contaminated soils were assessed using a variety of in vitro methodologies. A number of parameters can influence the bioavailability of inorganic contaminants in soil. Studies by Yang et al. (2002) identified Fe oxide content and pH to be the principal soil factors controlling arsenic bioaccessibility. Similarly, Juhasz et al. (2007a, 2007b) identified that arsenic bioaccessibility was related to the free or total Fe content of the soil. Recent evidence has suggested that contaminant ageing decreases bioavailability due to changes in surface phase complexes with increasing arsenic-soil residence time (Fendorf, La Force & Li 2004). Once sorbed by the soil, increasing residence time (ageing) may lead to the development of inner sphere complexes, surface diffusion within micropores or surface precipitates (Aharoni & Sparks 1991), resulting in a decrease in As bioavailability.

Contaminant bioaccessibility may also be influenced by the in vitro methodology employed. As outlined in previous sections, the composition of the gastrointestinal fluid


(i.e. bile concentration, inclusion of food additives etc.) can significantly influence the release of contaminants from the soil matrix. While round robin studies have been undertaken to compare bioaccessibility methodologies for arsenic, cadmium and lead (see Tables 15 and 16) (Oomen et al. 2002; Van de Wiele, Verstraete & Siciliano 2007), such comparisons have not been undertaken for organic contaminants.

Table 14. Assessment of organic contaminant bioaccessibility using in vitro gastrointestinal extraction methods.

Contaminant Methoda Contaminated matrix Bioaccessibility Reference

PAHs I PAH contaminated top soil was assessed (atmospheric deposition).

Total concentration of PAHs was 49 ± 1.5 mg kg-1, organic matter content was 3.3%, soil pH was 6.5.

The soil contained 94% sand, 4% silt and 2% clay.

PAH bioaccessibility was low ranging from 0.1 to 1.4%.

Van de Wiele, Verstraete &

Siciliano 2004

PAHs B Soils with varying properties (TOC: 0.5-1.7%; Clay: 8-38%; pH: 4.6-7.4) were spiked with phenanthrene to achieve concentrations of approximately 200 and 400 mg kg-1.

Phenanthrene bioaccessibility ranged from 18-70% (at 200 mg kg-1) and 53-89% at (400 mg kg-1).

Pu et al. 2004

PAHs B Manufacturing gas plant, mine waste and household / construction waste soil: benzo[a]pyrene and dibenz[a,h]anthracene concentrations ranged from 6-270 and 1-43 mg kg-1 respectively.

Benzo[a]pyrene bioaccessibility ranged from 2-16%. Dibenz[a,h]anthracene bioaccessibility ranged from 11-21%.

Grøn et al. 2007

PAHs, PCBs E 22 polluted soils were assessed. PAH concentrations ranged from 20 to 5000 mg kg-1 while PCB concentrations ranged from 59 to 692 µg kg-1.

PAH and PCB bioaccessibility ranged from 5-40%. When food was added to the assay, bioaccessibility increased to 40-85%.

Hack & Selenka 1996

PCBs B Soils with varying properties (TOC: 0.5-2.9%; Clay: 8-38%; pH: 4.6-7.4) were spiked with PCBs to achieve a concentration of approximately 300 mg kg-1.

PCB #52 bioaccessibility ranged from 56-79% while PCB #118 bioaccessibility ranged from 40-63%.

Pu et al. 2006

Lindane, PCBs

B OECD-medium (10% peat, 20% kaolin clay, and 70% sand) was employed as an artificial standard soil. The OECD-medium was spiked with lindane (1-10 mg kg-1), PCB #52 (3.5-35 mg kg-1), PCB #118 (3.5-35 mg kg-1), PCB #153 (7-70 mg kg-1) and PCB #180 (3.5-35 mg kg-1).

PCB bioaccessibility ranged from 30-40%. Lindane bioaccessibility was 57%.

Oomen et al. 2000

Lindane, PCBs

B OECD-medium (10% peat, 20% kaolin clay, and 70% sand) was employed as an artificial standard soil. The OECD-medium was spiked with a mixture of lindane (2 mg kg-1), PCB #52, PCB #118, PCB #180 (7 mg kg-1) and PCB #153 (14 mg kg-1).

Under fasting conditions,

25%, 15% and 60% of PCBs were sorbed to bile salts, digestive proteins and soil respectively (23%, 32% and 40% for lindane). <1% of PCBs and 5% of lindane were freely dissolved.

Oomen et.al 2001

PCDD/F F PCDD/F concentration in red slag ranged from 10–100 µg kg-1 I-TEQ.

PCDD/F bioaccessibility ranged from 3-39%. When milk powder was added to the assay, bioaccessibility increased 2- to 4-fold.

Wittsiepe et al. 2001

TCDD D TCDD concentrations in 8 soils ranged from 1.7-139 ng kg-1 (6-340 ng kg-1 TEQ). Total organic carbon content of the soils ranged from 1-4%.

Dioxins/ furan bioaccessibility ranged from 19 to 34%.

Ruby et al. 2002

aSee Tables 12 and 13 for a description of the in vitro bioaccessibility method employed.


Table 15. Assessment of arsenic bioaccessibility using in vitro gastrointestinal extraction methods.

As source As (mg kg-1) In vitro method (see Tables 11 and

12)

As bioaccessibility

Reference

Herbicide (18) Pesticide (13) Mine waste (4 composite waste types from 60 samples) Mine waste (9) Mine waste (8) Geogenic (11)

22-1345 39-3034 180±50 - 1340±720 210-2570 606-12781 13-422

SBRC Gastric

PBET

PBET

6-89% 9-89% 3.9±0.5 - 8.5±4.35% 7.1±1.7 – 10.2±2.1% 5-35% 1-22%

Juhasz et al. 2007a Diacomanolis, Ng & Noller 2007 Bruce 2004 Smith, Weber & Juhasz 2009

Industrial (2)

55-236

SBET DIN (fasted) DIN (fed) BARGE SHIME TIM

11, 50% 18, 44% 11, 30% 19, 95% 1, 6% 15, 52%

Oomen et al. 2002

Mine waste (15) 233-17500 IVG Gastric IVG Intestinal PBET Gastric PBET Intestinal

3.6-24.8% 3.5-22.7% 1.4-18.3% 1.5-12.5%

Rodriguez et al. 1999

CCA (20) 37.4-310 IVG Gastric IVG Intestinal

15.8-63.6% 17.0-66.3%

Girouard & Zagury 2009

CCA (12) 23-220 IVG Gastric IVG Intestinal

20.7-63.6% 25.0-66.3%

Pouschat & Zagury 2006

Ironstone formations (15)

19-102 PBET 1-10% Wragg, Cave & Nathanail 2007

Pesticide (12) 31.3-2143 IVG Gastric

IVG Intestinal

2-76%

3-90%

Sarkar et al. 2007

Mine waste (87) Mineralised soil (20) Tailings (22)

249-68900 123-205 1280-204500

PBET 0.5-42% 6.8-16.7% 0.6-61.1%

Palumbo-Roe & Klinck 2008

Mine waste (3) 1406-20000 PBET Gastric PBET Intestinal

10-12.5% 16-35.6%

Williams et al. 1998

Residential (2) House dust (1)

170-3900 PBET Gastric PBET Intestinal

34-55% 31-50%

Ruby et al. 1996

River sediment (9)

33-264 PBET Gastric 1-11% Devesa-Rey et al. 2008

Mine waste / soil (27)

204-9025 BARGE Saliva-gastric-intestinal

10-34%

Button et al. 2009

Mine waste (18) 40-824 SBRC Gastric SBRC Intestinal

0.1-25% 0-3%

Navarro et al. 2006

Residential soil near smelter (10)

214-5214 PBET Intestinal 39-66% Carrizales et al. 2006


Table 16. Assessment of lead bioaccessibility using in vitro gastrointestinal extraction methods.

Pb Source Pb (mg kg-1) In vitro method Pb bioaccessibility

Reference

Urban soils (15) 32-6330 DIN 2-21% (without milk powder

11-56% (with milk powder)

Marschner et al. 2006


0.5-6% 0.1-0.7%

Ruby et al. 1993

Mine waste (6) 140-1800 SBET 65-110% Schaider et al. 2008

Mine waste (18) 1270-14200 IVG Gastric IVG Intestinal

0.7-36.3% 0.02-1.16%

Schroder et al. 2004


1.3-83% 2.7-54%

Ruby et al. 1996

Carpet dust (15) 209-1770 PBET Gastric PBET Intestinal

52-77% 5-32%

Yu, Yin & Lioy 2006

Mine waste (1) Urban soil (1) Residential soil (1)

68-2924 Saliva, gastric and intestinal phases

39-69% Hamel, Buckley & Lioy 1998

Residential soil (2) Mine waste (2)

2141-77007 BARGE Saliva-gastric Intestinal

15-56% 5-25%

Denys et al. 2007

Mine waste (1) 3060 PBET DIN BARGE SHIME

TIM

13% (fasted), 22% (fed) 14% (fasted), 29% (fed) 32% (fasted), 24% (fed) 2% (fasted), 24% (fed) 33% (fasted), 7% (fed)

Van de Wiele, Verstraete & Siciliano 2007

Mine waste (18) 153-4817 SBRC Gastric SBRC Intestinal

0.5-86% 0-80%

Navarro et al. 2006

Residential soil near smelter (10)

391-4062 PBET Intestinal 13-64% Carrizales et al. 2006

Soil contaminated with pottery flakes (10)

50-2400 BARGE 28-73% Oomen et al. 2003

Urban soil (2) Incinerator site (3)

646-3905 SBRC Gastric SBRC Intestinal Rel-SRBC Intestinal

36-64% 1.2-2.7% 14-26%

Juhasz, Weber et al. 2009

Industrial (2) Mine waste (4 composite waste types from 60 samples) Mine waste (9)

612-6380 550±80 - 5450±2700 170-12100

SBET DIN (unfed) DIN (fed) BARGE SHIME TIM PBET PBET

56, 91% 23, 40% 16, 31% 29,66% 1, 4% 4, 13% 10.2±1.5 – 13.4±2.6% 3.7±1.7 – 24.2±4.8%

Oomen et al. 2002 Diacomanolis, Ng & Noller 2007 Bruce 2004


5.1 Correlation between in vivo bioavailability and in vitro bioaccessibility

In vitro assays have the potential to overcome the time and expense limitations of in vivo studies, thereby providing a surrogate measurement of bioavailability that is quick and inexpensive compared to animal models (Basta, Rodriguez & Castell 2001; Ruby et al. 1996). However, in order to validate the use of in vitro assays as a surrogate measure of contaminant bioavailability, the relationship between in vivo bioavailability and in vitro bioaccessibility needs to be established. A dearth of information is available on the validation of in vitro assays for organic contaminants: Figure 2 shows the in vivo – in vitro relationship for five individual organic compounds utilising three in vivo animal models and two in vitro methodologies (data from Grøn et al. 2007; Pu et al. 2004, 2006). For most contaminants detailed in Figure 1 (benzo[a]pyrene, dibenz[a,h]anthracene, PCB #52, PCB #118), limited data points are available (≤ 4) which precludes confidence in determining in vivo – in vitro relationships. However, for PAHs (benzo[a]pyrene, dibenz[a,h]anthracene), in vitro bioaccessibility determination generally underestimated in vivo bioavailability. For PCBs, bioaccessibility and bioavailability values were similar, although variability did exist which may result in the underestimation of PCB #118 bioavailability when using the in vitro assay. For phenanthrene, a relationship was observed between in vivo bioavailability and in vitro bioaccessibility (r = 0.73, p < 0.05) for eight spiked soils representing four soil types and two phenanthrene loadings (Pu et al. 2004). While this relationship was significant, it would be interesting to note the performance of the in vitro assay in predicting phenanthrene bioavailability in contaminated soil (i.e. spiked versus aged PAHs; individual compounds versus multiple PAHs).


Data is presented for:

Phenanthrene – (in vivo method – rat [blood]; in vitro method – Oomen et al. 2001) from Pu et al. (2004)

Benzo[a]pyrene – ( in vivo method – mouse [urine]; in vitro method – Oomen et al. 2003)

Benzo[a]pyrene – (in vivo method – swine [urine, faeces]; in vitro method – Oomen et al. 2003) from Grøn et al. (2007)

Dibenz[a,h]anthracene – (in vivo method – swine [urine, faeces]; in vitro method – Oomen et al. 2003) from Grøn et al. (2007)

PCBs –

– – PCB #118 (in vivo method – rat [blood]; in vitro method – Oomen et al. 2001) from Pu et al. (2006)

– – PBC #52 (in vivo method – rat [blood]; in vitro method – Oomen et al. 2001) from Pu et al. (2006).

Figure 2. Correlation between in vivo bioavailability and in vitro bioaccessibility for organic contaminants.

Although data on arsenic bioavailability is accumulating in the literature, only a few studies have correlated arsenic bioavailability with arsenic bioaccessibility as determined by in vitro assays. In the study of Rodriguez et al. (1999), there was no statistical difference in the relative availability of arsenic in mining and smelting material when measured by in vitro (IVG method) or in vivo (swine feeding trials) methods. However, the calcine samples analysed by the in vitro methods were not statistically equivalent to the in vivo method, as arsenic bioavailability was underestimated by the in vitro method. In the study of Juhasz et al. (2007b), Pearson correlation method was used to calculate the relationship between arsenic bioaccessibility as measured by the SBET method and arsenic bioavailability as measured by the in vivo method (Figure 2). As the spiked soil study of Juhasz et al. (2008) utilised the SBET method, and in vivo experiments were conducted under the same condition as per the 2007 study, this data was also included in Pearson calculations. Comparison of in vitro and in vivo results indicate that the correlation between arsenic bioaccessibility and arsenic bioavailability was significant (p = 0.01; n = 49), with SBET assessment of contaminated soils providing a good prediction of in vivo arsenic bioavailability (in vivo arsenic


bioavailability (mg kg-1) = 14.19 + 0.93 x [SBET arsenic bioaccessibility (mg kg-1)]; r2 = 0.92) (see Figure 3). While some variability was observed in the data collected during in vivo studies, this may be attributable to physiological intra-species variability including genetic factors, disparity in stomach clearance times, stomach pH and rates of arsenic absorption. The same variability was not evident in the in vitro assay with the data being extremely reproducible under laboratory conditions.

Figure 3. Correlation between arsenic bioaccessibility (in vitro SBET assay) and arsenic bioavailability (in vivo swine assay) for railway corridor, dip site, mine site, gossan and arsenic-spiked soils (n = 49). 95% confidence limits are also shown (from Juhasz et al. 2007a).

In a recent study by Juhasz, Smith et al (2009), arsenic bioaccessibility in contaminated soils (n = 12) was assessed using four in vitro assays (SBRC, IVG, PBET, DIN). In vitro results were compared to in vivo relative arsenic bioavailability data (swine assay) to ascertain which methodologies best correlate with in vivo data. Arsenic bioaccessibility in contaminated soils varied depending on the in vitro method employed. For SBRC and IVG methods, arsenic bioaccessibility generally decreased when gastric phase values were compared to the intestinal phase. In contrast, extending PBET and DIN assays from the gastric to the intestinal phase resulted in an increase in arsenic bioaccessibility for some soils tested. Comparison of in vitro and in vivo results demonstrated that the in vitro assay encompassing the SBRC gastric phase provided the best prediction of in vivo relative arsenic bioavailability (r2 = 0.75, Pearson correlation = 0.87). However, relative arsenic bioavailability could also be predicted using gastric or intestinal phases of IVG, PBET and DIN assays but with varying degrees of confidence (r2 = 0.53–0.67, Pearson correlation = 0.73–0.82).


Research undertaken as part of a US EPA study (Drexler & Brattin 2007; US EPA 2007b) determined that dissolution of Pb phases following gastric phase extraction provided a good prediction of relative Pb bioavailability determined using juvenile swine. Similar results were obtained by Ruby et al. (1996) (PBET method and a Sprague-Dawley rat model) and Schroder et al. (2004) (IVG method and a swine model). Poor in vivo – in vitro correlations were obtained for intestinal phase data and relative Pb bioavailability, which was attributed to the complex non-equilibrium chemical system for Pb in the small intestines (Ruby et al. 1996). It was suggested, however, that the use of the small intestinal phase data would be preferable as a measure of Pb bioaccessibility (Ruby et al. 1996).

While a good relationship was observed between gastric phase Pb dissolution and in vivo relative Pb bioavailability, the US EPA (2007b) cautioned that the majority of samples used were derived from similar sources (mining and milling activities) and that some forms of Pb that were absent in these soils may not follow the observed correlation (US EPA 2007b). This was highlighted in the study of Marschner et al. (2006), who reported that the absolute and relative bioavailability of Pb in five urban and industrial soils, determined using liver, kidney, bone and urine data from soil dosed minipigs, was not related to Pb bioaccessibility determined using the standardised German in vitro assay (Hack & Selenka 1996).

In the study of Juhasz, Weber et al. (2009), Pb bioavailability in contaminated soils was determined using an in vivo swine assay while Pb bioaccessibility was assessed using an in vitro method (SBRC) encompassing gastric (SBRC-G) and intestinal (SBRC-I) phases. Initially, bioaccessibility studies were performed with a Pb reference material (Pb acetate, 1-10 mg l-1) in order to determine the influence of pH on Pb solubility. In the gastric phase (pH 1.5), Pb solubility was 100% (100 ± 2.9%, n = 16) irrespective of the Pb concentration added, however, when the pH of the intestinal phase was increased to near neutral, Pb solubility decreased to 14.3 ± 7.2%. In contaminated soils, Pb bioaccessibility varied from 35.7 to 64.1% and 1.2 to 2.7% for SBRC-G and SBRC-I phases respectively. When relative bioaccessibility (Rel-SBRC-I) was calculated by adjusting the dissolution of Pb from contaminated soils by the solubility of Pb acetate at pH 6.5 (intestinal phase pH), Rel-SBRC-I values ranged from 11.7 to 26.1%. The relationship between Pb bioaccessibility determined using either the SBRC-G, SBRC-I and Rel-SBRC-I methods and relative Pb bioavailability, determined using the in vivo swine assay, was ascertained. When relative Pb bioavailability was plotted against Pb bioaccessibility, three distinct data groupings were evident. Comparison of in vitro and in vivo results indicated that the correlation between Pb bioaccessibility and relative Pb bioavailability varied depending on the in vitro methodology used.

Regression models (SPSS, Release 15.0.1, 2006) determined that Rel-SBRC-I provided the best estimate of in vivo relative Pb bioavailability for soils used in this study. Although research undertaken as part of a US EPA study (Drexler & Brattin 2007; US EPA 2007b) determined that gastric phase extraction provided a good estimate of relative Pb bioavailability, the relationship between SBRC-G and in vivo relative Pb bioavailability was poor. However, if in vitro – in vivo relationships were calculated using bioaccessibility/bioavailability data expressed as mg available Pb kg-1, SBRC-G provided a good estimate of relative Pb bioavailability. This was not surprising as the amount of Pb in the intestinal phase is dependent on gastric phase dissolution. A poor relationship was observed between in vivo relative Pb bioavailability


and SBRC-I values as previously reported by Ruby et al. (1996), Schroder et al. (2004) and Marschner et al. (2006).

Key findings

There is a lack of consistent approaches to correlate in vivo and in vitro data in contaminated soils:

researchers have generally selected a range of in vitro bioaccessibility methodologies

no standard soil samples have been utilised to compare efficacy of different in vitro methodologies.


6. Knowledge gaps

There is a general lack of acceptance of the use of bioaccessibility data to predict contaminant bioavailability. Several major reasons have been identified for this approach in this review. These include:

lack of validation between in vivo and in vitro studies for many contaminants

a method developed and validated for one contaminant is not always appropriate for other contaminants, and

there are no standard reference materials that can be used to validate the results from bioaccessibility tests or to check their reproducibility.

One of the major limitations for addressing these concerns is the actual cost of any in vivo animal study. Lack of validation between in vitro and in vitro considerably inhibits the regulatory acceptance of any in vitro methodology. Only arsenic and lead bioavailability in contaminated soil has been reported extensively. However, no single methodology has been used consistently across a number of different inorganic and organic contaminants. Comparison of commonly employed in vitro bioaccessibility methodologies for inorganic and organic soil contaminants is severely lacking and needs urgent addressing. Furthermore, inclusion of standard reference soils needs to be included in experimental design by researchers to enable the comparison of laboratory efficacy in conducting research. The authors suggest standard reference soils NIST SRM 2711 or NIST SRM 2710.

Although it is commonly assumed a single bioaccessibility methodology may be used to determine bioaccessibility in a range of contaminants, there is no scientific evidence to support this view. Considerable research needs to be focused on identifying a single methodology or the methods that may be used effectively for estimating contaminant bioaccessibility. For organic contaminants there are little or no studies that validate any in vitro bioaccessibility methodology with in vivo bioavailability studies. To address this, researchers should focus on a single contaminant such as benzo(α)pyrene.


7. Conclusions

When assessing the impact of an ingested chemical on human health risk assessment, the chemical’s toxicity is influenced by the degree to which it is absorbed from the gastrointestinal tract into the body (i.e. its bioavailability). As oral references doses (RfDs) and cancer slope factors (CSFs) are generally expressed in terms of ingested dose, rather than absorbed dose, the variability in absorption between different exposure media, chemical forms etc. may significantly influence risk calculations. In Australia, NEPM HILs are derived using a bioavailability default value of 100%. However, the assumption that 100% of the soil-borne contaminant is bioavailable may overestimate exposure thereby influencing risk calculations. As a result, assessment of contaminant bioavailability may help to refine exposure modelling for Tier 2 human health risk assessment.

In the absence of human studies or the availability of suitable epidemiological data, the relative bioavailability of soil-borne contaminants may be assessed using in vivo methods. Bioavailability assessment using an in vivo model is considered to be the most reliable method for refining exposure models for Tier 2 human health risk assessment. While a variety of animal models have been utilised for the assessment of RBA, standard operating procedures for these in vivo models are currently unavailable. However, the US EPA have developed a guidance document for evaluating the bioavailability of metals in soil for use in human health risk assessment (US EPA 2007a), while Rees et al. (2009) detail an in vivo swine assay for the determination of relative arsenic bioavailability in contaminated soil and plant matrices. The use of juvenile swine for the assessment of RBA is prescribed by the US EPA, however, other in vivo models (e.g. rodents, primates) may be utilised if deemed suitable for the contaminant of interest. Bioavailability endpoints may include the determination of the contaminant in blood, organs, fatty tissue, urine and faeces, urinary metabolites, DNA adducts and enzyme induction (e.g. cytochrome P450 mono-oxygenases).

Several in vitro methods have been developed for the prediction of contaminant relative bioavailability. These in vitro methodologies do not attempt to replicate conditions found in vivo, but mimic key processes such as contaminant dissolution. In vitro assays have the potential to overcome the time and expense limitations of in vivo studies, thereby providing a surrogate measurement of bioavailability that is quick and inexpensive compared to animal models. However, in order for an in vitro bioaccessibility test system to be useful in predicting the in vivo relative bioavailability of a test material, it is necessary to establish empirically that a strong correlation exists between the in vivo and in vitro results across a variety of sample types.

A limited number of studies have established the relationship between in vivo RBA and in vitro bioaccessibility. For inorganic contaminants, these studies have been limited to arsenic (Basta et al. 2007; Juhasz et al., 2007b, Juhasz, Smith et al. 2009; Rodriguez et al. 1999) and lead (Drexler & Brattin 2007; Juhasz, Weber et al. 2009; Schroder et al. 2004; US EPA 2007b) with only the US EPA (2007b) study gaining regulatory acceptance. Data is emerging for other inorganic contaminants (e.g. cadmium, nickel), however the correlation between RBA and bioaccessibility is lacking or limited information is available, which precludes confidence in the determination of in vivo – in vitro relationships.


A number of in vitro methodologies have been developed for the determination of organic contaminant bioaccessibility, however the correlation between bioaccessibility and RBA is lacking or limited information is available, which precludes confidence in the determination of in vivo – in vitro relationships. Consequently, in vitro assays should not be used as a surrogate assay for predicting relative bioavailability for organic contaminants.

Figure 4 illustrates when and how bioavailability and bioaccessibility can be incorporated into the risk assessment framework for the incidental soil ingestion exposure pathway. The recommended decision framework is intended for the collection of data to inform site-specific risk-based decisions. The decision framework prevents the use of invalidated models for site-specific risk assessment.

Figure 4. Schematic diagram for the determination of contaminant bioavailability and bioaccessibility for human health risk assessment.

Risk characterisation

No In vivo

assessment cost

prohibitive?

Collect soil <2mm size

Contaminant analysis

Contaminant concentration

> HIL?

Report no riskNo

In vivo – in vitro correlated method

available for estimating contaminant

bioavailability?

Yes

Yes

Yes

No

Assume 100% bioavailability

Sieve soil to <250 µm


In vitro assessment using validated method

Sieve soil to <250 µm


In vivo assessment

Exposure assessment


8. References

Aharoni, C & Sparks, DL 1991, ‘Kinetics of soil chemical reactions – A theoretical treatment’, in DL Sparks, & DL Suarez (eds), Rates of soil chemical processes, SSSA, Madison, WI., pp 1-19.

Alexander, FW, Clayton, BE & Delves, HT 1974, ‘Mineral and trace-metal balances in children receiving normal and synthetic diets’, Quarterly Journal of Medicine, New Series, XLIII, no. 169, pp. 89-111.

Alexander, M 1995, ‘How toxic are toxic chemicals in soil?’, Environmental Science and Technology vol. 29, no. 11, pp. 2713-2717.

Alexander, M 2000 ‘Aging, bioavailability, and overestimation of risk from environmental pollutants’, Environmental Science and Technology vol. 34, no. 20, pp. 4259-4265.

Alvaro, D, Cantafora, A, Attili, AF, Ginanni Corradini S & De Luca, C 1986, ‘Relationships between bile salts, hydrophobicity and phospholipid composition in bile of various animal species’, Comparative Biochemistry and Physiology Part B vol. 83, iss. 3, pp. 551-554.

ATSDR 1995, Toxicological profile for polycyclic aromatic hydrocarbons (PAHs), Agency for Toxic Substances and Disease Registry, US Department of Health and Human Services, US public health service, Atlanta, GA, USA, available at <http://www.atsdr.cdc.gov/>.

ATSDR 2000, Toxicological profile for polychlorinated biphenyls. Agency for Toxic Substances and Disease Registry, US Department of Health and Human Services, US public health service, Atlanta, GA, USA, available at <http://www.atsdr.cdc.gov/>.

Basta, NT, Rodriguez, RR & Casteel, SW 2001, ‘Bioavailability and risk of arsenic exposure by the soil ingestion pathway’, in WT Frankenberger (ed.), Environmental Chemistry of Arsenic, Marcel Dekker Inc., New York, NY.

Basta, NT, Foster, JN, Dayton, EA, Rodriguez, RR & Casteel, SW 2007, ‘The effect of dosing vehicle on arsenic bioaccessibility in smelter-contaminated soils’, Journal of Environmental Science and Health Part A 42: pp. 1275-1281.

Bauer, E, Guo, Z, Ueng, Y.-F, Bell, LC, Zeldin, D & Guengerich, FP 1995, ‘Oxidation of benzo(a)pyrene by recombinant human cytochrome P450 enzymes’, Chemical Research in Toxicology vol. 8, iss. 1, pp. 136-142.

Bordelon, NR, Donnelly, KC, King, LC, Wolf, DC, Reeves, WR & George, SE 2000, ‘Bioavailability of the Genotoxic Components in Coal Tar Contaminated Soils in Fischer 344 Rats’, Toxicological Sciences vol. 56, no. 1, pp. 37-48.

Bouchard, M & Viau, C 1997, ‘Urinary excretion of benzo(a)pyrene metabolites following intravenous, oral, and cutaneous benzo(a)pyrene administration’, Canadian Journal of Physiology and Pharmacology vol. 75, pp. 185-192.

Bridges, CC & Zalpus RK 2005, ‘Molecular and ionic mimicry and the transport of toxic metals’, Toxicology and Applied Pharmacology vol. 204, iss. 3, pp. 274-308.

Bruce, SL 2004, ‘Development of a Risk Assessment Tool to Minimise the Impact of Arsenic and Lead Toxicity from Mine Tailings’, PhD thesis, The University of Queensland.


Bruce, SL, Noller, BN, Grigg, AH, Mullen, BF, Mulligan, DR, Ritchie, PJ, Currey, N & Ng JC 2003, ‘A field study conducted at Kidston Gold Mine, to evaluate the impact of arsenic and zinc from mine tailing to grazing cattle’, Toxicology Letters vol. 137, issues 1-2, pp. 23-34.

Buchet, JP, Lauwerys, R & Roels H 1981a, ‘Comparison of the urinary excretion of arsenic metabolites after a single oral dose of sodium arsenite, monomethylarsonate, or dimethylarsinate in man’, International Archives of Occupational and Environmental Health vol. 48, no. 1, pp. 71-79.

Buchet, JP, Lauwerys, R & Roels, H 1981b, ‘Urinary excretion of inorganic arsenic and its metabolites after repeated ingestion of sodium metaarsenite by volunteers’, International Archives of Occupational and Environmental Health vol. 48, no. 2, pp. 111-118.

Buckley, TJ & Lioy PJ 1992, ‘An examination of the time course from human dietary exposure to polycyclic aromatic hydrocarbon to urinary elimination of 1-hydroxypyrene’, British Journal of Industrial Medicine vol. 49, no. 2, pp. 113-124.

Budinsky, RA, Rowlands, JC, Casteel, S, Fent, G, Cushing, CA, Newsted, J, Giesy, JP, Ruby, MV & Aylward, LL 2007, ‘A pilot study of oral bioavailability of dioxins and furans from contaminated soils: Impact of differential hepatic enzyme activity and species differences’, Chemosphere vol. 70, issue 10, pp. 1774-1786.

Button, M, Watts, MJ, Cave, MR, Harrington, CF & Jenkin, GT 2009, ‘Earthworms and in vitro physiologically-based extraction tests: complementary tools for a holistic approach towards understanding risk at arsenic-contaminated sites’, Environmental Geochemistry and Health vol. 31, no. 2, pp. 273-282.

Carrizales, L, Razo, I, Téllez-Hernández, JI, Torres-Nerio, R, Torres, A, Batres, LE, Cubillas, A-C & Díaz-Barriga, F 2006, ‘Exposure to arsenic and lead of children living near a copper-smelter in San Luis Potosi, Mexico: Importance of soil contamination for exposure of children’, Environmental Research vol. 101, iss. 1, pp. 1-10.

Casteel, SW, Cowart, RP, Weis, CP, Henningsen, GM, Hoffman, E, Brattin, WJ, Guzman, RE, Starost, MF, Payne, JT, Stockham, SL, Becker, SV, Drexler, JW & Turk, JR 1997, ‘Bioavailability of lead to juvenile swine dosed with soil from the Smuggler Mountain NPL site of Aspen, Colorado’, Fundamental and Applied Toxicology vol. 36, no. 2, pp.177-187.

Cave, MR, Wragg, J, Palumbo, B & Klinch, BA 2003, Measurement of the bioaccessibility of arsenic in UK soils, ISBN 1 84432 128 2, Environment Agency, Bristol, UK.

Cavret, S, Laurent, C, Feidt, C, Laurent, F & Rychen, G 2003, ‘Intestinal absorption of 14C from 14C-phenanthrene, 14C-benzo(a)pyrene and 14C-tetrachlorodibenzo-para-dioxin: Approaches with the Caco-2 cell line and portal absorption measurements in growing pigs’, Reproduction Nutrition Development vol. 43, no.2, pp. 145-154.

Chaloupka, K, Steinberg, M, Santostefano, M, Rodriguez, LV, Goldstein, L & Safe, S 1995, ‘Induction of CYP1A-1 and CYP1A-2 gene expression by a reconstituted mixture of polynuclear aromatic hydrocarbons in B6C3F1 mice’, Chemico-Biological Interactions vol. 96, iss. 3, pp. 207-221.

Charman, WN, Porter, CJH, Mithani, S & Dressman, JB 1997, Physicochemical and physiological mechanisms for the effects of food on drug absorption: the role of lipids and pH’, Journal of Pharmaceutical Sciences vol 86, no. 3, 269-282.

Chiou, CT 1998, ‘Soil sorption of organic pollutants and pesticides’, in RA Meyers (ed.), Encyclopedia of environmental analysis and remediation, John Wiley and Sons, New York, NY.


Daugherty, AL & Mrsny, RJ 1999, ‘Transcellular uptake mechanisms of the intestinal epithelial barrier. Part one’, Pharmaceutical Science & Technology Today vol. 2, iss. 4, pp. 144-151.

Davis, A, Ruby, MV & Bergstrom PD 1992, ‘Bioavailability of arsenic and lead in soils from the Butte, Montana, mining district’, Environmental Science and Technology vol. 26, no. 3, pp. 461-468.

Dean, JR & Ma R 2007, ‘Approaches to assess the oral bioaccessibility of persistent organic pollutants: A critical review’, Chemosphere vol. 68, issue 8, pp. 1399-1407.

Degan, LP & Philips, SF 1996, ‘Variability of gastrointestinal transit in healthy women and men’, Gut vol. 39, no. 2, pp. 299-305.

Denys, S, Caboche, J, Tack, K & Delalain P 2007, ‘Bioaccessibility of lead in high carbonate soils’, Journal of Environmental Science and Health, Part A vol. 42, iss. 9, pp. 1331-1339.

Devesa-Rey, R, Paradelo, R, Díaz-Fierros, F & Barral, MT 2008,’ Fractionation and bioavailability of arsenic in the bed sediments of the Anllóns River (NW Spain)’, Water, Air and Soil Pollution vol. 195, no. 1-4, pp. 189-198.

Diacomanolis, V, Ng, JC & Noller, BN 2007, ‘Development of mine site close-out criteria for arsenic and lead using a health risk approach’, in A Fourie, M Tibbett & J Wiertz (eds), Mine Closure 2007: Proceedings of the Second International Seminar on Mine Closure, Sandiago, Chile, 16-19 October, pp. 191-198.

DIN, Deutsches Institut fur Normung e. V 2000, Soil quality – Absorption availability of organic and inorganic pollutants from contaminated soil material, DIN E 19738.

Drexler, JW & Brattin, WJ 2007, ‘An in vitro procedure for estimation of lead relative bioavailability: with validation’, Human and Ecological Risk Assessment vol. 13, iss. 2, pp. 383-401.

Ellickson, KM, Meeker, RJ, Gallo, MA, Buckley, BT & Lioy, PJ 2001, ‘Oral bioavailability of lead and arsenic from a NIST standard reference soil material’, Archives of Environmental Contamination and Toxicology vol. 40, no. 1, pp. 128-135.

Environment Agency 2007, In-vitro bioaccessibility testing: Current science and way forward, Environment Agency Science Update 2, Environment Agency UK.

EPA 1996, Bioavailability of arsenic and lead in environmental substrates: results of an oral dosing study of immature swine, EPA 910/R-96-002, Seattle, WA: EPA Region 10.

Fendorf, S, La Force, MJ & Li, GC 2004, ‘Temporal changes in soil partitioning and bioaccessibility of arsenic, chromium, and lead’, Journal of Environmental Quality vol. 33, no. 6, pp. 2049-2055.

Fouchécourt, MO, Arnold, M, Berny, P, Videmann, B, Rether, B & Rivière, JL 1999, ‘Assessment of the bioavailability of PAHs in rats exposed to a polluted soil by natural routes: Induction of EROD activity and DNA adducts and PAH burden in both liver and lung’, Environmental Research vol. 80, iss. 4, pp. 330-339.

Freeman, GB, Johnson, JD, Killinger, JM, Liao, SC, Feder, PI, Davis, AO, Ruby, MV, Chaney, RL, Lovre, SC & Bergstrom PD 1992, ‘Relative bioavailability of lead from mining waste soil in rats’, Fundamental and Applied Toxicology vol. 19, no. 3, pp. 388-398.

Freeman, GB, Johnson, JD, Killinger, JM, Liao, SC, Davis, AO, Ruby, MV, Chaney, RL, Lovre, SC & Bergstrom, PD 1993, ‘Bioavailability of arsenic in soil impacted by smelter activities following oral administration in rabbits’, Fundamental and Applied Toxicology vol. 21, no. 1, pp. 83-88.


Freeman, GB, Johnson, JD, Llao, SC, Schoof, RA & Bergstrom, PD 1993, ‘Pilot study of absolute bioavailability of arsenic in soil impacted by smelter activities following oral administration in rabbits and monkeys’, in Proceedings: International Conference on Arsenic Exposure and Health Effects, New Orleans, July 28-30, 1993, SEGH, 1-3.

Freeman, GB, Schoof, RA, Ruby, MV, Davis, AO, Dill, JA, Liao, SC, Lapin, CA & Bergstrom, PD 1995, ‘Bioavailability of arsenic in soil and house dust impacted by smelter activities following oral administration in cynomolgus monkeys’, Fundamental and Applied Toxicology vol. 28, no. 2, pp. 215–222.

Friedman, HI & Nylund B 1980, ‘Intestinal fat digestion, absorption, and transport. A review’, American Journal of Clinical Nutrition vol. 33, iss. 5, pp. 1108-1139.

Fries, GF, Marrow, GS & Somich, CJ 1989, ‘Oral bioavailability of aged polychlorinated biphenyl residues contained in soil’, Bulletin of Environmental Contamination and Toxicology vol. 43, no. 5, pp. 683-690.

Girouard, E & Zagury, GJ 2009, ‘Arsenic bioaccessibility in CCA-contaminated soils: Influence of soil properties, arsenic fractionation and particle-size fraction’, Science of the Total Environment vol. 407, iss. 8, pp. 2576-2585.

Gorelick, FS & Jamieson, JD 1994, ‘The pancreatic acinar cell: structure-functional relationship’, in LR Johnson (ed.), Physiology of the Gastrointestinal Tract, vol. 2, Raven Press, New York, NY.

Groen, K, Vaessen, HAMG, Kliest, JJG, Deboer, JLM, Vanooik, T, Timmerman, A & Vlug RF 1994, ‘Bioavailability of inorganic arsenic from bog containing soil in the dog’, Environmental Health Perspectives vol. 102, no. 2, pp. 182-184.

Grøn, C, Oomen, A, Weyand, E, & Wittsiepe, J 2007, ‘Bioaccessibility of PAH from Danish soils’ Journal of Environmental Science and Health Part A vol. 42, issue 9, pp. 1233-1239.

Grova, N, Feidt, C, Laurent, C & Rychem, G 2002, ‘[14C]-Milk, urine and faeces excreation kinetics in lactating goats after an oral administration of [14C] polycyclic aromatic hydrocarbons’, International Dairy Journal vol. 12, issue 12, pp. 1025–1031.

Guyton, AC 1991, Textbook of Medical Physiology, WB Saunders, Philadelphia, PA.

Hack, A & Selenka, F 1996, ‘Mobilization of PAH and PCB from contaminated soil using a digestive tract model’, Toxicology Letters vol. 88, issues 1-3, pp. 199-210.

Hamel, SC, Buckley, B & Lioy, PJ 1998, ‘Bioaccessibility of metals in soils for different liquid to solid ratios in synthetic gastric fluid’, Environmental Science and Technology vol. 32, no. 3, pp. 358-362.

Hecht, SS, Grabowski, W & Groth, K 1979, ‘Analysis of feces for benzo[a]pyrene after consumption of charcoal-broiled beef by rats and humans’, Food and Cosmetics Toxicology vol. 17, issue 3, pp. 223-227.

Hillgren, KM, Kato, A & Borchardt, RT 1995, ‘In vitro systems for studying intestinal drug absorption’, Medicinal Research Reviews vol. 15, issue 2, pp. 83-109.

Holman, HYN 2000, In vitro gastrointestinal mimetic protocol for measuring bioavailable contaminants, US Patent 6040188, University of California, USA.

Hughes, MF, Menache, M & Thompson, DJ 1994, ‘Dose-dependent disposition of sodium arsenate in mice following acute oral exposure’, Fundamental and Applied Toxicology vol. 22, iss. 1, pp. 80-89.

Hursh, JB & Suomela, J 1968, ‘Absorption of 212Pb from the gastrointestinal tract of man’, Acta Oncologica vol. 7, issue 2, pp. 108-120.


IARC 1983, Polynuclear aromatic compounds, Part I, Chemical, environmental and experimental data, IARC monographs on the evaluation of the carcinogenic risk of chemicals to humans, vol. 32, pp. 35-91, International Agency for Research on Cancer, France.

IPCS 1995, A review of selected persistant organic pollutants: DDT-Aldrin-Dieldrin-Endrin-Chlordane, Heptachlor-Hexachlorobenzene-Mïrex-Toxaphene, Polychlorinated biphenyls, Dioxins and Furans, PCS/95.39, International Programme on Chemical Safety (IPCS), World Health Organization, Canada, pp. 1-145.

Johnson, LR 2001, ‘Gastric Secretion’, in LR Johnson (ed.), Gastrointestinal Physiology, sixth edition, Mosby Publishing, St. Louis, Missouri.

Jongeneelen, FJ, Anzion, RBM, Scheepers, PTJ, Bos, RP, Henderson, PT, Nijenhuis, EH, Veenstra, SJ, Brouns, RME & Winkes, A 1988, ‘1-hydroxypyrene in urine as a biological indicator of exposure to polycyclic aromatic hydrocarbons in several work environments’, Annals of Occupational Hygiene vol. 32, no. 1, pp. 35-43.

Juhasz, AL, Mallavarapu, M & Naidu, R 2000, ‘Bioavailability: The major challenge (constraint) to bioremediation of organically contaminated soils’, in DL Wise, DJ Trantolo, EJ Cichon, HI Inyang & U Stottmeister (eds), Remediation of Hazardous Waste Contaminated Soils, 2nd edition, vol. 1, Engineering Considerations and Remediation Strategies, section 1-1 Engineering Issues in Waste Remediation, Marcel Dekker, New York, NY.

Juhasz, AL, Smith, E & Naidu, R 2003, ‘Estimation of human bioavailability of arsenic in contaminated soil’, in A Langley, M Gilbet & B Kennedy (eds), Proceedings of the Fifth National Workshop on Assessment of Site Contamination, National Environmental Protection Council Service Corporation, 183-194.

Juhasz, A, Smith, E, Weber, J, Naidu, R, Rees, M, Rofe, A, Kuchel, T & Sansom, L 2008, ‘Effect of soil ageing on in vivo arsenic bioavailability in two dissimilar soils’, Chemosphere vol. 71, issue 11, pp. 2180–2186.

Juhasz, AL, Smith, E, Weber, J, Naidu, R, Rees, M, Rofe, A, Kuchel, T & Sansom, L 2009, ‘Assessment of Four Commonly Employed In Vitro Arsenic Bioaccessibility Assays for Predicting In Vivo Arsenic Bioavailability in Contaminated Soils’, Environmental Science and Technology vol. 43, no. 24, pp. 9487-9494.

Juhasz, AL, Smith, E, Weber, J, Rees, M, Rofe, A, Kuchel, T, Sansom, L & Naidu, R 2006, ‘In vivo assessment of arsenic bioavailability in rice and its significance for human health risk assessment’, Environmental Health Perspectives vol. 114, no. 12, pp. 1826-1831.

Juhasz, AL, Smith, E, Weber, J, Rees, M, Rofe, A, Kuchel, T, Sansom, L & Naidu R 2007a, ‘In vitro assessment of arsenic bioaccessibility in contaminated (anthropogenic and geogenic) soils’, Chemosphere vol. 69, issue 1, pp. 69-78.

Juhasz, AL, Smith, E, Weber, J, Rees, M, Rofe, A, Kuchel, T, Sansom, L & Naidu, R 2007b, ‘Comparison of in vivo and in vitro methodologies for the assessment of arsenic bioavailability in contaminated soils’, Chemosphere vol. 69, iss. 6, pp. 961-966.

Juhasz, AL., Weber, J, Smith, E, Naidu, R, Marschner, B, Rees, M, Rofe, A, Kuchel, T & Sansom, L 2009, ‘Evaluation of SBRC-gastric and SBRC-intestinal methods for the prediction of in vivo relative lead bioavailability in contaminated soils’, Environmental Science and Technology vol. 43 no. 12, pp. 4503-4509.

Kasprzak, KS 2002, ‘Oxidative DNA and protein damage in metal-induced toxicity and carcinogenesis’, Free Radical Biology and Medicine vol. 32, issue 10, pp. 958-967.


Kenyon, EM, Hughes, MF & Levander, OA 1997, ‘Influence of dietary selenium on the disposition of arsenate in the female B6C3F1 mouse’, Journal of Toxicology and Environmental Health, Part A vol. 51, iss. 3, pp. 279-299.

Koganti, A, Spina, DA, Rozett, K, Ma, B-L & Weyand, EH 1998, ‘Studies on the applicability of biomarkers in estimating the systemic bioavailability of polynuclear aromatic hydrocarbons from manufactured gas plant tar contaminated soils’, Environmental Science and Technology vol. 32, no. 20, pp. 3104-3112.

Krishnamohan, M 2008, ‘Carcinogenicity of monomethylarsonous acid (MMAIII) and sodium arsenate (AsV) and identification of early warning biomarkers’, PhD thesis, The University of Queensland.

Krishnamohan, M, Qi, L, Lam, PKS, Moore, MR & Ng JC 2007a, ‘Urinary arsenic and porphyrin profile in C57BL/6J mice chronically exposed to monomethylarsonous acid (MMAIII) for two years’, Toxicology and Applied Pharmacology vol. 224, iss. 1, pp. 89-97.

Krishnamohan, M, Wu, HJ, Huang, SH, Maddelena, R, Lam, PKS, Moore, MR & Ng, JC 2007b, ‘Urinary arsenic methylation and porphyrin profile of C57Bl/6J mice chronically exposed to sodium arsenate’, Science of the Total Environment vol. 379, issues 2-3, pp. 235-243.

Krishnamohan, M, Wu, HJ, Huang, SH, Maddelena, R, Lam, PKS, Ng, JC & Moore, MR 2006, ‘Use of urinary porphyrin profiles as an early warning biomarker for monomethylarsonous acid (MMAIII) exposure’, in Environmental Toxicology 10, WIT Transactions on Biomedicine and Health Book Series, WIT Press Southhampton, UK, pp. 193-201.

Laurent, C, Feidt, C, Grova, N, Mpassi, D, Lichtfouse, E, Laurent, F & Rychen, G 2002, ‘Portal absorption of 14C after ingestion of spiked milk with 14C-phenanthrene, 14C-benzo(a)pyrene or 14C-TCDD in growing pigs’, Chemosphere vol. 48, issue 8, pp. 843-848.

Linz, DG & Nakles, DV (eds) 1997, Environmentally acceptable endpoints in soil: risk-based approach to contaminated site management based on availability of chemicals in soil, American Academy of Environmental Engineers, 630 p.

Luthy, RG, Allen, GR, Brusseau, ML, Cunningham, SD, Gschwend, PM, Pignatello, JJ, Reinhard, M, Traina, SJ, Weber, WJ & Westall, JC 1997, ‘Sequestration of hydrophobic organic contaminant by geosorbants’, Environmental Science and Technology vol. 31, no. 12, pp. 3341-3347.

Madara, JL & Trier, JS 1994, ‘The functional morphology of the mucosa of the small intestine’, in LR Johnson (ed.), Physiology of Gastrointestinal Tract, 3rd edition, vol. 2, Raven Press, New York, N.Y.

Marafante, E, Bertolero, F, Edel, J, Pietra, R & Sabbioni, E 1982, ‘Intracellular interaction and biotransformation of arsenite in rats and rabbits’, Science of the Total Environment vol. 24, iss.1, pp. 27-39.

Marschner, B, Welge, P, Hack, A, Wittsiepe, J & Wilhelm, M 2006 ‘Comparison of soil Pb in vitro bioaccessibility and in vivo bioavailability with Pb pools from a sequential soil extraction’, Environmental Science and Technology vol. 40, no. 8, pp. 2812-2818.

Minekus, M, Marteau, P, Havenaar, R & Huis in’t Veld, JHJ 1995, ‘A multicompartmental dynamic computer-controlled model simulating the stomach and small intestine’, Alternatives to Laboratory Animals vol. 23, no. 2, pp. 197-209.


Mirvish, SS, Ghadirian, P, Wallcave, L, Raha, C, Bronczyk, S & Sams, J P 1981, ‘Effect of diet on fecal excretion and gastrointestinal tract distribution of unmetabolized benzo(a)pyrene and 3-methylcholanthrene when these compounds are administered orally to hamsters’, Cancer Research vol. 41, no. 6, pp. 2289-2293.

Molly, K, Van de Woestyne, M & Verstraete, W 1993, ‘Development of a 5-step multi-chamber reactor as a simulation of the human intestinal microbial ecosystem’, Applied Microbiology and Biotechnology vol. 39, no. 2, pp. 254-258.

Moore, MR, McColl, KEL, Rimington, C & Goldberg, A 1987, Disorders of Porphyrin metabolism, Plenum Medical Book Company, New York & London, pp. 1-374.

Navarro, MC, Pérez-Sirvent, C, Martínez-Sánchez, MJ, Vidal, V & Marimón, J 2009, ‘Lead, cadmium and arsenic bioavailability in the abandoned mine site of Cabezo Rajao (Murcia, SE Spain)’, Chemosphere vol. 63, iss. 3, pp. 484-489.

NEPI 2000, Assessing the bioavailability of metals in soil for use in human health risk assessment, Washington, DC: National Environmental Policy Institute.

NEPI (National Environmental Policy Institute) 2000, Assessing the bioavailability of organic chemicals in soil for use in human health risk assessments, National Environmental Policy Institute, Washington, D.C.

Ng, JC & Moore, MR 1996, ‘Bioavailability of arsenic in soils from contaminated sites using a 96 hour rat blood model’ in A. Langley, B. Markey and H. Hill (eds), The Health Risk Assessment and Management of Contaminated Sites, Commonwealth Department of Human Services and Health and the Environmental Protection Agency, Contaminated Sites Monograph Series, No. 5, 355-363, South Australian Health Commission, Adelaide.

Ng, JC, Bruce, SL & Noller, BN 2003, ‘Chapter 14: Laboratory and field evaluation of potential arsenic exposure from mine tailings to grazing cattle’, in WR Chappell, CO Abernathy & RL Calderon (eds), Arsenic: Exposure and Health Effects, Elsevier Science, Oxford, pp. 179-193.

Ng, JC, Kratzmann, SM, Qi, L, Crawley, H., Chiswell, B & Moore, MR 1998, ‘Speciation and bioavailability: risk assessment of arsenic contaminated sites in a residential suburb in Canberra’, The Analyst vol. 123, pp. 889-892.

Ng, JC, Noller, BN, Bruce, SL & Moore, M 2003, ‘Bioavailability of metals and arsenic at contaminated sites from cattle dips, mined land and naturally occurring mineralisation origins’, in A Langley, M Gilbey & B Kennedy (eds), Health and Environmental Assessment of Site Contamination, 5th edition, NEPC Service Corporation, Adelaide, pp. 163-181.

Ng, JC, Qi, L, Wang, JP, Xiao, XL, Shahin, M, Moore, MR & Prakash, AS 2001, ‘Mutations in C57Bl/6J and metallothionein knock-out mice ingested sodium arsenate in drinking water for over two years’, in WR Chappell, CO Abernathy & RL Calderon (eds) Arsenic: Exposure and Health Effects, Elsevier Science, Oxford, pp. 231-242.

Ng, JC, Wang, JP, Zheng, BS, Zhai, C, Maddalena, R, Liu, F & Moore, MR 2005, ‘Urinary porphyrins as biomarkers for arsenic exposure among susceptible populations in Guizhou province, China’, Toxicology and Applied Pharmacology vol. 206, iss. 2, pp. 176-184.

NRC (National Research Council) 2003, Bioavailability of contaminants in soils and sediments: Processes, tools and applications, The National Academies Press, Washington, DC.

Odanaka, Y, Matano, O & Goto, S 1980, ‘Biomethylation of inorganic arsenic by the rat and some laboratory animals. Bulletin of Environmental Contamination and Toxicology vol. 24, no. 1, pp. 452-459.


Oomen, AG 2000, Determinants of oral bioavailability of soil-borne contaminants, Dutch National Institute of Public Health and the Environment, Bilthoven, Netherlands.

Oomen, AG, Hack, A, Minekus, M, Zeydner, E, Cornelis, C, Schoeters, G, Verstraete, W, Van de Wiele, T, Wragg, J, Rompelberg, CJ M, Sips, AJAM & Van Wynen, JH 2002, ‘Comparison of five in vitro digestion models to study the bioaccessibility of soil contaminants’, Environmental Science and Technology vol. 36, no. 15, pp. 3326-3334.

Oomen, AG, Rompelberg, CJM, Bruil, MA, Dobbe, CJG, Pereboom, DPKH & Sips, AJAM 2003, ‘Development of an in vitro digestion model for estimation of bioaccessibility of soil contaminants’, Archives of Environmental Contamination and Toxicology vol. 44, no. 3, pp. 281-287.

Oomen, AG, Rompelberg, CJM, Van de Kemp, E, Pereboom, DPKH, De Zwart, LL & Sips, AJAM 2004, ‘Effect of bile type on the bioaccessibility of soil contaminants in an In vitro digestion model’, Archives of Environmental Contamination and Toxicology vol. 46, no. 2, pp. 183-188.

Oomen, AG, Sips, AJAM, Groten, JP, Sijm, DTHM & Tolls, J 2000, ‘Mobilization of PCBs and lindane from soil during in vitro digestion and their distribution among bile salt micelles and proteins of human digestive fluid and the soil’, Environmental Science and Technology vol. 34, no. 2, pp. 297-303.

Oomen, AG, Tolls, J, Kruidenier, M, Bosgra, SSD, Sips, AJAM & Groten, JP 2001, ‘Availability of polychlorinated biphenyls (PCBs) and lindane for uptake by intestinal Caco-2 cells’, Environmental Health Perspectives vol. 109, no. 7, pp. 731-737.

Owen, BA 1990, ‘Literature-derived absorption coefficients for 39 chemicals via oral and inhalation routes of exposure’, Regulatory Toxicology and Pharmacology vol. 11, iss. 3, pp. 237-252.

Palumbo-Roe, B & Klinck, B 2007, ‘Bioaccessibility of arsenic in mine waste-contaminated soils: A case study for an abandoned arsenic mine in SW England’, Journal of Environmental Science and Health, Part A vol. 42, iss. 9, pp. 1251-1261.

Pomroy, C, Charbonneau, SM, McCullough, RS & Tam, GKH 1980, ‘Human retention studies with 74As’, Toxicology and Applied Pharmacology vol. 53, iss. 3, pp. 550-556.

Pouschat, P & Zagury, GJ 2006, ‘In vitro gastrointestinal bioavailability of arsenic in soils collected near CCA-treated utility poles’, Environmental Science and Technology vol. 40, no. 13, pp. 4317-4323.

Pu, X, Lee, LS, Galinsky, RE & Carlson, GP 2004, ‘Evaluation of a rat model versus a physiologically based extraction test for assessing phenanthrene bioavailability from soils’, Toxicological Sciences vol. 79, no. 1, pp. 10-17.

Pu, X, Lee, LS, Galinsky, RE & Carlson, GP 2006, ‘Bioavailability of 2,3',4,4',5-pentachlorobiphenyl (PCB118) and 2,2',5,5'-tetrachlorobiphenyl (PCB52) from soils using a rat model and a physiologically based extraction test’, Toxicology vol. 217, iss. 1, pp. 14-21.

Rabache, M, Billaud, C & Adrian, J 1985, ‘Evolution du benzo(a)pyrene dans le tube digestif. 1. Sa disparition’ International Journal for Vitamin and Nutrition Research vol. 55, no. 1, pp. 113-117.

Rabinowitz, MB, Wetherill, GW & Kopple, JD 1976, ‘Kinetic analysis of lead metabolism in healthy humans’, Journal of Clinical Investigation vol. 58, iss. 2, pp. 260-270.

Ramesh, A, Inyang, F, Hood, DB, Archibong, AE, Knuckles, ME & Nyanda, AM 2001, ‘Metabolism, bioavailability, and toxicokinetics of benzo[a]pyrene [B(a)P] in F-344 rats following oral administration’, Experimental and Toxicologic Pathology vol. 53, iss. 4, pp. 275-290.


Ramesh, A, Walker, SA, Hood, DB, Guillén, MD, Schneider, K & Weyand, EH 2004, ‘Bioavailability and risk assessment of orally ingested polycyclic aromatic hydrocarbons’, International Journal of Toxicology vol. 23, iss. 5, pp. 301-333.

Rees, M, Sansom, L, Rofe, A, Juhasz, AL, Smith, E, Weber, J, Naidu, R & Kuchel, T 2009, ‘Principles and application of an in vivo swine assay for the determination of arsenic bioavailability in contaminated matrices’, Environmental Geochemistry and Health vol. 31, suppl. 1, pp. 167-177.

Reeves, WR, Mcdonald, TJ, Bordelon, NR, George, SE & Donnelly, KC 2001, ‘Impacts of Aging on In Vivo and In Vitro Measurements of Soil-Bound Polycyclic Aromatic Hydrocarbon Availability’, Environmental Science and Technology vol. 35, no. 8, pp. 1637-1643.

Roberts, SM, Weimar, WR, Vinson, JRT, Munson, JW, Bergeron, RJ, 2002, ‘Measurement of arsenic bioavailability in soil using a primate model’, Toxicological Sciences vol. 67, no. 2, pp. 303-310.

Robinson, JR, Felton, JS, Levitt, RC, Thorgeirsson, SS & Neberts, DW 1975, ‘Relationship between aromatic hydrocarbon responsiveness and the survival times in mice treated with various drugs and environmental compounds’, Molecular Pharmacology vol. 11, no. 6, pp. 850-865.

Rodriguez, RR & Basta, NT 1999, ‘An in vitro gastrointestinal method to estimate bioavailable arsenic in contaminated soils and solid media’, Environmental Science and Technology vol. 33, no. 4, pp. 642-649.

Roos, PH 2002, ‘Differential induction of CYP1A1 in duodenum, liver and kidney of rats after oral intake of soil containing polycyclic aromatic hydrocarbons’, Archives of Toxicology vol. 76, no. 2, pp. 75-82.

Rotard, W, Christmann, W, Knoth, W & Mailahn, W 1995, Bestimmung der resorptionsverfügbaren PCDD/PCDF aus Kieselrot. UWSF-Z Umweltchem Ökotox 7:3-9.

Ruby, MV, Davis, A, Kempton, JH, Drexler, JW & Bergstrom, PD 1992, ‘Lead bioavailability – dissolution kinetics under simulated gastric conditions’, Environmental Science and Technology vol. 26, no. 6, pp. 1242-1248.

Ruby, MV, Davis, A, Link, TE, Schoof, R, Chaney, RL, Freeman, GB & Bergstrom, P 1993, ‘Development of an in vitro screening test to evaluate the in vivo bioaccessibility of ingested mine-waste lead’, Environmental Science and Technology vol. 27, no. 13, pp. 2870-2877.

Ruby, MV, Davis, A, Schoof, R, Eberle, S & Sellstone, CM 1996, ‘Estimation of lead and arsenic bioavailability using a physiologically based extraction test’, Environmental Science and Technology vol. 30, no. 2, pp. 422-430.

Ruby, MV, Fehling, KA, Paustenbach, DJ, Landenberger, BD & Holsapple, MP 2002, ‘Oral bioaccessibility of dioxins/furans at low concentrations (50–350 ppt toxicity equivalent) in soil’, Environmental Science and Technology vol. 36, iss. 22, pp. 4905-4911.

Ruby, MV, Schoof, R, Brattin, W, Goldade, M, Post, G, Harnois, M, Mosby, DE, Casteel, S, Berti, W, Carpenter, M, Edwards, D, Cragin, D & Chappell, W 1999, ‘Advances in evaluating the oral bioavailability of inorganics in soil for use in human health risk assessment’, Environmental Science and Technology vol. 33, no. 21, pp. 3697-3705.


Sarkar, D, Makris, KC, Parra-Noonan, MT & Datta, R 2007, ‘Effect of soil properties on arsenic fractionation and bioaccessibility in cattle and sheep dipping vat sites’, Environment International vol. 33, iss. 2, pp. 164-169.

Schaider, LA, Senn, DB, Brabander, DJ, McCarthy, KD & Shine, JP 2007, ‘Characterization of zinc, lead, and cadmium in mine waste: implications for transport, exposure and bioavailability’, Environmental Science and Technology vol. 41, no. 11, pp. 4164-4171.

Schlebaum, W, Badora, A, Schraa, G & van Riemsdijk, WH 1998, ‘Interactions between a hydrophobic organic chemical and natural organic matter: equilibrium and kinetic studies’, Environmental Science and Technology vol. 32, iss. 15, pp. 2273-2277.

Schoket, B, Phillips, DH, Poitier, MC & Vincze, I 1993, ‘DNA adducts in peripheral blood lymphocytes from aluminium plant production workers determined by 32P-postlabelling and enzyme linked immunosorbent assay’, Environmental Health Perspectives vol. 99, pp. 307-309.

Schricker, BR, Miller, DD, Rasmusen, RR & van Campen, D 1981, ‘A comparison of in vivo and in vitro methods for determining availability of iron from meals’, American Journal of Clinical Nutrition vol. 34, pp. 2257-2263.

Schroder, JL, Basta, NT, Casteel, SW, Evans, TJ, Payton, ME & Si, J 2004, ‘Validation of the in vitro gastrointestinal (IVG) method to estimate relative bioavailable lead in contaminated soils’, Journal of Environmental Quality vol. 33, pp. 513-521.

Scott, WC & Dean, JR 2005, ‘An assessment of the bioavailability of persistent organic pollutants from contaminated soil’, Journal of Environmental Monitoring vol. 7, iss. 7, pp. 710-715.

Semple, KT, Morriss, AWJ & Paton, GI 2003, ‘Bioavailability of hydrophobic organic contaminants in soils: fundamental concepts and techniques for analysis’, European Journal of Soil Science vol. 54, iss. 4, pp. 809-818.

Seydel, JK & Schaper, KJ 1982, ‘Quantitative structure-pharmacokinetic relationships and drug design’, Pharmacology and Therapeutics vol. 15, iss. 2, pp. 131-182.

Simon, GL & Gorbach, SL 1995, ‘National alimentary tract microflora’, in MJ Blaser (ed.), Infections of the Gastrointestinal Tract, Raven Press, New York, NY.

Sips, AJAM, Bruil, MA, Dobbe, CJG, van de Kamp, E, Oomen, AG, Pereboom, DPKH, Rompelberg, CJM & Zeilmaker, MJ 2001, Bioaccessibility of contaminants from ingested soil in humans. Method and research on the bioaccessibility of lead and benzo[a]pyrene, Bilthoven, National Institute of Public Health and the Environment, RIVM report 711701012.

Smith, E, Weber J & Juhasz, AL 2009, ‘Arsenic distribution and bioaccessibility across particle size fractions in historically contaminated soils’, Environmental Geochemistry and Health vol. 31, suppl. 1, pp. 85-92.

Steele, MJ, Beck, BD, Murphy, BL, & Strauss, HS 1990, ‘Assessing the contribution from lead in mining wastes to blood’, Regulatory Toxicology and Pharmacology vol. 11, iss. 2, pp. 158-190.

Stroo, HF, Jensen, R, Nakles, DV & Liban, CB 2000, ‘Assessing environmental risks at a former manufactured gas plant site using environmentally acceptable endpoints’, Environmental Science and Technology vol. 34, iss. 18, pp. 3831-3836.

Swindle, MM & Smith, AC 1998, ‘Comparative anatomy and physiology of the pig’, Scandinavian Journal of Laboratory Animal Science Supplement vol. 25, pp. 11-22.


Tam, GKH, Charbonneau, SM, Bryce, F, Pomroy, C & Sandi, E 1979, ‘Metabolism of inorganic arsenic (74As) in humans following oral ingestion’, Toxicology and Applied Pharmacology vol. 50, is. 2, pp. 319-322.

Tang, X-Y, Tang, L, Zhu, Y-G, Xing, B-S, Duan, J & Zheng, M-H 2006, ‘Assessment of the bioaccessibilitry of polycyclic aromatic hydrocarbons in soils from Beijing using an in vitro test’, Environmental Pollution vol. 140, iss. 2, pp. 279-285.

Thomas, PE, Reik, LM, Ryan, DE & Levin, W 1983, ‘Induction of two immunochemically related rat liver cytochrome P-450 isozymes, cytochrome P-450c and P-450d, by structurally diverse xenobiotics’, The Journal of Biological Chemistry vol. 258, no. 7, pp. 4590-4598.

Tran, HP, Prakash, AS, Barnard, R & Ng, JC 2002, ‘Arsenic inhibits the repair of DNA damage induced by benzo(a)pyrene’, Toxicology Letters, vol. 133, iss. 1, pp. 59-67.

Tso, P 1994, ‘Intestinal Lipid Absorption’, in LR Johnson (ed.), Physiology of the Gastrointestinal Tract, 3rd edition, vol. 2, Raven Press, New York, NY.

US EPA 1994, Guidance manual for the integrated exposure uptake biokinetic model for lead in children, OSWER 9285.7-15-1, EPA/540/R-93/081, Office of Solid Waste and Emergency Response, United States Environmental Protection Agency, Washington, DC 20460.

US EPA 1996, Bioavailability of arsenic and lead in environmental substrates: Results of an oral dosing study of immature swine, EPA 910/R-96-002, Superfund, United States Environmental Protection Agency, Washington, DC 20460

US EPA 2007a, Guidance for evaluating the oral bioavailability of metals in soils for use in human health risk assessment, OSWER 9285.7-80, Office of Solid Waste and Emergency Response, United States Environment Protection Agency, Washington, DC 20460.

US EPA 2007b, Estimation of relative bioavailability of lead in soil and soil-like materials using in vivo and in vitro methods, OSWER 9285.7-77, Office of Solid Waste and Emergency Response, United States Environmental Protection Agency, Washington, DC 20460.

US EPA 2008, Standard operating procedure for an in vitro bioaccessibility assay for lead in soil, EPA 9200.1-86, Office of Solid Waste and Emergency Response, United States Environment Protection Agency, Washington, DC 20460.

US EPA 2009, Validation assessment of in vitro lead bioaccessibility assay for predicting relative bioavailability of lead in soils and soil-like materials at superfund sites, OSWER 9200.3-51, United States Environment Protection Agency, Washington, DC 20460.

Vahter, M & Norin, H 1980, ‘Metabolism of 74As-labeled trivalent and pentavalent inorganic arsenic in mice’, Environmental Research vol. 21, iss. 2, pp. 446-457.

Vahter, M Couch, R, Nermell, B & Nilsson, R 1995, ‘Lack of methylation of inorganic arsenic in the chimpanzee’, Toxicology and Applied Pharmacology vol. 133, iss. 2, pp. 262-268.

Van de Wiel, JAG, Fijneman, PHS, Duijf, CMP, Anzion, RBM, Theuws, JLG & Bos, RP 1993, ‘Excretion of benzo(a)pyrene and metabolites in urine and feces of rats: Influence of route of administration, sex and long-term ethanol treatment’, Toxicology vol. 80, iss. 2-3, pp. 103-115.


Van de Wiele, TR, Oomen, AG, Wragg, J, Cave, M, Minekus, M, Hack, A, Cornelis, C, Rompelberg, CJ, De Zwart, LL, Klinck, B, Van Wijnen, J, Verstraete, W & Sips, AJ 2007, ‘Comparison of five in vitro digestion models to in vivo experimental results: lead bioaccessibility in the human gastrointestinal tract’, Journal of Environmental Science and Health Part A vol. 42, iss.9, pp. 1203-1211.

Van de Wiele, TR, Verstraete, W & Siciliano, SD 2004, ‘Polycyclic aromatic hydrocarbon release from a soil matrix in the in vitro gastrointestinal tract’, Journal of Environmental Quality vol. 33, pp. 1343-1353.

Van Schooten, FJ, Moonene, EJ, van der Wal, L & Kleinjans, JCS 1997, ‘Determination of polycyclic aromatic hydrocarbons (PAH) and their metabolites in blood, feces, and urine of rats orally exposed to PAH contaminated soils’, Archives of Environmental Contamination and Toxicology vol. 33, no. 3, pp. 317-322.

Wall, KL, Gao, W, Koppele, JM, Kwei, GY, Kauffman, FC & Thurman, RG 1991, ‘The liver plays a central role in the mechanism of chemical carcinogenesis due to polycyclic aromatic hydrocarbons’, Carcinogenesis vol. 12, no. 5, pp. 783-786.

Walsh, JH 1994, ‘Gastrointestinal hormones’, in LR Johnson (ed.), Physiology of the Gastrointestinal Tract, vol. 1, Raven Press, New York, NY.

Wang, JP, Maddalena, R, Zheng, B, Zai, C, Liu, F & Ng JC 2009, ‘Arsenicosis status and urinary malondialdehyde (MDA) in people exposed to arsenic contaminated-coal in China’, Environment International vol. 35, iss. 3, pp. 502-506.

Weis, CP & La Velle, JM 1991, ‘Characteristics to consider when choosing an animal model for the study if lead bioavailability’, Chemical Speciation and Bioavailability vol. 3, pp. 113-119.

West, CE & Horton, BJ 1976, ‘Transfer of polycyclic hydrocarbons from diet to milk in rats, rabbits and sheep’, Life Sciences vol. 19, iss. 10, pp. 1543-1552.

Wildgrube, HJ, Stockhausen, H, Petri, J, Füssel, U & Lauer, H 1986, ‘Naturally occurring conjugated bile acids measured by high-performance liquid chromatography in human, dog and rabbit bile’, Journal of Chromatography A vol. 353, pp. 207-213.

Williams, TM, Rawlins, BG, Smith, B & Breward, N 1998, ‘In-vitro determination of arsenic bioavailability in contaminated soil and mineral beneficiation waste from Ron Phibun, Southern Thailand: A basis for improved human risk assessment’, Environmental Geochemistry and Health vol. 20, no. 4, pp. 169-177.

Wittsiepe, J, Erlenkämper, B, Welge, P, Hack, A & Wilhelm, M 2007, ‘Bioavailability of PCDD/F from contaminated soil in young Goettingen minipigs’, Chemosphere vol. 67, iss. 9, pp. S355-S364.

Wittsiepe, J, Schrey, P, Hack, A, Selenka, F & Wilhelm, M 2001, ‘Comparison of different digestive tract models for estimating bioaccessibility of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/F) from red slag ‘Kieselrot’’, International Journal of Hygiene and Environmental Health vol. 203, iss. 3, pp. 263-273.

Wragg, J, Cave, M & Nathanail, P 2007, ‘A study of the relationship between arsenic bioaccessibility and its solid-phase distribution in soils from Wellingborough, UK’, Journal of Environmental Science and Health Part A vol. 42, iss. 9, pp. 1303-1315.

Xing, B & Pignatello, JJ 1997, ‘Dual-mode sorption of low-polarity compounds in glassy poly(vinyl chloride) and soil organic matter’, Environmental Science and Technology vol. 31, iss. 3, pp. 792-799.

Yamauchi, H & Yamamura, Y 1979, ‘Dynamic change of inorganic arsenic and methylarsenic compounds in human urine after oral intake as arsenic trioxide’, Industrial Health, no. 17, pp. 79-83


Yamauchi, H & Yamamura, Y 1985, ‘Metabolism and excretion of orally administered arsenic trioxide in the hamster’, Toxicology vol. 34, iss. 2, pp. 113-121.

Yamazaki, H & Kakiuchi, Y 1989, ‘The uptake and distribution of benzo(a)pyrene in rat after continuous oral administration’, Toxicological and Environmental Chemistry vol. 24, iss. 1, pp. 95-104.

Yang, J, Barnett, MO, Jardine, PM, Basta, NT & Casteel, SW 2002, ‘Adsorption, sequestration, and bioaccessibility of As(V) in soils’, Environmental Science and Technology vol. 36, no. 21, pp. 4562-4569.

Yu, CH, Yin, L-M & Lioy, PJ 2006, ‘The bioaccessibility of lead (Pb) from vacuumed house dust on carpets in urban residences’, Risk Analysis vol. 26, iss. 1, pp. 125-134.

Zakharyan, RA, Wildfang, E & Aposhian, HV 1996, ‘Enzymatic methylation of arsenic compounds. III. The marmoset and tamarin, but not the rhesus, monkeys are deficient in methyltransferases that methylate inorganic arsenic’, Toxicology and Applied Pharmacology vol. 140, iss. 1, pp. 77-84.

Zhang, QY, Wikoff, J, Dunbar, D, Fasco, M & Kaminsky, L 1997a, ‘Regulation of cytochrome P4501A1 expression in rat small intestine’, Drug Metabolism and Disposition vol. 25, no. 1, pp. 21-26.

Zhang, XJ, Thomas, PT, Yarborough, A, Ma, B-L & Weyand, EH 1997b, ‘CYP1A enzymes as biomarkers for assessment of exposure to polycyclic aromatic hydrocarbons in mice’, The Toxicologists, Society of Toxicology Annual Meeting, vol. 36, pp. 690-698.

Zhitkovich, A 2005, ‘Importance of chromium-DNA adducts in mutagenicity and toxicity of chromium (VI)’, Chemical Research in Toxicology vol. 18, iss. 1, pp. 3-11.

Ziegler, EE, Edwards, BB, Jensen, RL, Mahaffy, KR & Fomon, S 1978, ‘Absorption and retention of lead by infants’, Pediatric Research vol. 12, iss. 1, pp. 29-34.

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Contaminant bioavailability and bioaccessibility Part 1: A ... · Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review Figures Figure 1. Bioavailability