CRC for Contamination Assessment and Remediation of the Environment
no. 14
technical
report
Contaminant bioavailability and bioaccessibility
Part 1: A scientific and technical review
J.C. Ng, A.L. Juhasz, E. Smith and R. Naidu
Petroleum Vapour Model Comparison: Interim Report for CRC CARE
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
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(Commonwealth) and subsequent amendments, no part of this publication may be reproduced,
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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
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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.
Table of contents
1. NEPM bioavailability review
1 1.1 Scope1 1.2 Specification
2 1.2.1 General
2 1.2.2 Specific outcomes
2 1.2.3 Presentation
2 2. Background
2.1 National and international approaches
2.1.1 Key findings
3. Bioavailability and bioaccessibility concepts
3.1 Bioavailability processes as defined by NRC (2003)
3.2 Human contaminant exposure
3.2.1 Ingestion and release of contaminants from the soil matrix
3.3 Methods for the determination of contaminant bioavailability
3 3 4 5 6 10 10 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
4. In vitro methods for assessing contaminant bioaccessibility
31 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
5.1 Correlation between in vivo bioavailability and in vitro bioaccessibility
42 47 6. Knowledge gaps
52 7. Conclusions
53 8. References
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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
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Figures
Figure 1.
Figure 2.
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
Correlation between in vivo bioavailability and in vitro
bioaccessibility for organic contaminants
7
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 arsenicspiked soils
49
Figure 4.
Schematic diagram for the determination of contaminant
bioavailability and bioaccessibility for human health risk
assessment
54
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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.
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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.
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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
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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.
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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
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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.
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Bound
contaminant
Association
A
Biological membrane
C
Dissociation
Released
contaminant
D
Absorbed
contaminant in
organism
E
Site of biological
response
B
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
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.
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‘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 nonlethal, 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
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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.
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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.
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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
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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.
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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 PAHcontaminated 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
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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/Fcontaminated 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.
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Table 2. Absorption/bioavailability of orally administered benzo[a]pyrene (adapted from Ramesh et al. 2004).
Benzo[a]pyrene
dose (mg kg-1)
Dosing vehicle
Absorption/
bioavailability (%)
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
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
2.5 x 106 Bq
Vegetable oil
5.5
Grova et al. 2002
50 µCi
Spiked milk
33
Laurent et al. 2002
Animal model
Reference
Rats
Wistar
Goat
Swine
West & Horton 1976
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: 0.51.7%; 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: 0.52.9%; 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 6170% while the relative bioavailability ranged from 8799%.
Pu et al. 2006
Sprague-Dawley
male rats (urine,
faeces, body fat)
PCBs
Soil was spiked with 0.25,-1.0 µCi kg-1
14
C-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,
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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.
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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
5 mg/kg
Oral
43.8
44.1
87.9
1 mg/kg
i.v.
83.9
4.0
87.9
Hamster
arsenic acid
Odanaka, Matano & Goto1980
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
0.4 mg As/kg
s.c.
86
6.4
92.4
0.4 mg As/kg
Oral
77
8.0
85
4.0 mg As/kg
Oral
89
6.1
95.1
0.4 mg As/kg
s.c.
73
3.8
76.8
0.4 mg As/kg
Oral
90
7.1
97.1
4.0 mg As/kg
Oral
65
9.1
74.1
0.00012 mg As/kg
Oral
65.0
16.5
81.5
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
0.050 mg As/kg
i.p.
75.7
9.9
85.6
Mouse
Mouse
Mouse
Rabbit
sodium arsenate
sodium arsenite
sodium arsenate
sodium arsenite
Vahter & Norin 1980
Vahter & Norin 1980
Hughes, Menache & Thompson 1994
Marafante et al. 1982
Table 5. Absolute (comparison to intravenous route) oral bioavailability of arsenic from soil in laboratory animals.
Species
Duration
(hours)
i.v. dose
Soil or sample
Soil dose
Bioavailability
(%, mean±SD)
Method
Reference
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 dosedependency
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)
Freeman et al. 1995
House Dust (same
location)
0.26 mg As/kg
19 (10)
AUC for urinary
excretion (AUC
for blood in
parentheses)
sodium arsenate
0.62 mg As/kg
68 (91)
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
0.5 mg As/kg
(n=6)
4.31-9.87*
1.27-2.98#
AUC for urinary
excretion
Ng et al. 1998
5 mg As/kg
(n=4)
1.02-1.96*
0.26-0.67#
AUC for urinary
excretion
Bruce 2004
AUC for blood
Bruce 2004
Immature swine
144
0.01 - 0.31 mg
As/kg
Australian data
Wistar rat
96
0.5 mg As3+/kg
Soil (Canberra,
Australia)
0.5 mg As5+/kg
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
Cattle
240
0.5 mg As3+/kg
0.5 mg As5+/kg
Gold mine tailings,
waste rock, heap
leach material;
Former arsenic
mine mixed waste
1.0 mg As/kg
(n=3)
7±0.34 - 58±6.49
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
(hours)
Oral gavage
Soil or sample
Soil dose
Bioavailability
(%, mean±SD)
Reference
International data
New Zealand
white rabbit
120
New Zealand
white rabbit
120
Cynomolgus
monkey
Cynomolgus
monkey
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
505.7
1.35 mg As5+/kg
Smelter impacted soil
(Anaconda, Montana USA)
1.25 mg As/kg
70
Freeman, Johnson,
Killinger et al. 1993
120
1.35 mg As5+/kg
Smelter impacted soil
(Anaconda, Montana USA)
1.25 mg As/kg
43.6
Freeman, Johnson,
Killinger et al. 1993
168
0.62 mg As5+/kg
Smelter impacted soil
(Anaconda, Montana USA)
0.62 mg As/kg
10-30
Freeman et al. 1995
0.26
10-30
Soil
0.04 - 0.24 mg
As/kg
78
slag
0.61 - 1.52 mg
As/kg
42
5 mg As/kg
8.14.0
Dust (same location)
Immature swine
144
0.01 - 0.31 mg As/kg
Freeman, Johnson,
Llao et al. 1993
US EPA 1996
Australian data
Wistar rat
96
Sodium arsenite:
5 mg As3+/kg
Soils (from 16 cattle dip
Calcium arsenite:
5 mg As3+/kg
sites, NSW, Australia)
Sodium arsenate:
5 mg As5+/kg
14.47.1
6032.4
Ng & Moore 1996
Table 6 (cont.) Relative bioavailability (comparison to oral gavage) of arsenic from soil in laboratory animals by comparison of blood concentrations
otherwise specified.
Species
Wistar rat
Duration
(hours)
Oral gavage
96
Sodium arsenite:
0.5 mg As3+/kg
Soil or sample
Soils (9 samples from one
CCA contaminated site)
Soil dose
0.5 mg As/kg
Bioavailability
(%, mean±SD)
13.04.5
Calcium arsenite:
0.5 mg As3+/kg
32.211.2
Sodium arsenate:
0.5 mg As5+/kg
3813.2
Reference
Ng & Moore 1996
Sprague Dawley
rat
168
Sodium arsenite:
5 mg As3+/kg
Sodium arsenate:
5 mg As5+/kg
Gold mine tailings, waste
rock, heap leach material;
Former arsenic mine mixed
waste
0.5 mg As/kg
10.28 - 203.09
(urine)
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
0.2-0.25 mg As/kg
Juhasz et al. 2008
Brown Chromosol soilB
0.2-0.5 mg As/kg
5333.9 24.19.5g
97.21.9 - 92.95.3
110.6 - 697.24
(20.3 - 2919)Liver
Bruce et al. 2003;
Bruce 2004
Cattle
266 days
(repeated
doses)
Sodium arsenite: 0.5 mg
As3+/kg
Sodium arsenate: 0.5 mg
As5+/kg
Gold mine tailings, waste
rock, heap leach material;
Former arsenic mine mixed
waste
0.5 mg As/kg
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
subjects
Dose and frequency
Time
duration
% Dose in
urine
Arsenic acid
6
0.01 µg
5 days
57.9
Arsenic acid
6
0.06 ng
7 days
62.3
Arsenic trioxide
1
700 µg
3 days
68.2
Sodium
Arsenite
3
500 µg
4 days
45.1
Sodium
metaarsenite
1
125 µg x 5 days
14 days
1
250 µg x 5 days
1
% of Total urinary metabolites
As5+
As3+
MMA
DMA
27.2
20.6
51.0
Not Determined
7.9
31.7
Reference
Tam et al. 1979
Pomroy et al. 1980
28.2
32.2
Yamauchi & Yamamura 1979
25
21.3
53.7
Buchet, Lauwerys & Roels 1981a
54
16
34
50
14 days
73
7
20
73
500 µg x 5 days
14 days
74
19
21
60
1
1000 µg x 5 days
14 days
64
26
32
42
Sodium monomethyl 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
Yamauchi & Yamamura 1979
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
CRC CARE Technical Report no. 14
Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
26
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
CRC CARE Technical Report no. 14
Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
Equation [5]
27
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.
CRC CARE Technical Report no. 14
Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
28
Table 8. Absolute (comparison to intravenous route) oral bioavailability of lead in laboratory animals.
Species
Duration
(hours)
i.v. dose
Soil or sample
Soil dose
Bioavailability
(%, mean±SD)
0.83-22.61 µg
Pb/kg/day
(Food)
41.5
0.08-26.2 mg
Pb/kg/day (lead
acetate spiked feed)
15±8.1(blood)
7.4±4.1(bone)
11±7.1(liver)
0.12-23.8 mg Pb/kg
day (mine waste)
2.7±1.5(blood)
0.4±0.16(bone)
0.55±0.68 (liver)
Method
Reference
International
data
Humans
Sprague Dawley
rat
72
0.02, 0.20,
2.0 mg/kg
b.w. for 29
days
29-30 days
Food
Lead acetate
Urinary & faecal
excretion
Ziegler et al. 1978*
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
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
Australian data
*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
(hours)
Oral gavage
Soil or sample
Soil dose
Bioavailability
(%, mean±SD)
0.075-0.625 mg
Pb/kg-day
0-105
Reference
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
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)
Sprague Dawley
rat
30 days
0.076-25.7 mg
Pb/kg-day
Mine waste soil
0.119-23.2 mg
Pb/kg-day
10 mg Pb/kg (lead
acetate)
Zinc concentrate
0.025-0.225 mg
Pb/kg (lead acetate)
5 soils from a domestic
incinerator site and
residential land developed
on quarry fill material
5 mg/kg (lead
acetate)
0.5 mg Pb/kg-d
(lead acetate)
56-58 (blood),
74-86 (liver),
68-74 (kidney),
68-72 (bone)
12.1-26.8 (blood),
4.8-13.3 (bone),
0.6-13.6 (liver)
US EPA 2007b
Casteel et al. 1997
Freeman et al. 1992
Australian data
Sprague Dawley
rat
Juvenile swine
168
5 days
Cattle
240
Cattle
266 days
10 mg Pb/kg
38±6.8
Bruce 2004
0.225-0.53 mg Pb/kg
7.4 ± 4.2 – 19.1 ±
1.9
Zinc concentrate
5 mg Pb/kg
80±13.2
Bruce 2004
Zinc concentrate
0.5 mg Pb/kg-d
48±3.2Liver
Bruce 2004
Juhasz, Weber et al.
2009
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.
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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.
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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.
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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
30.03 g glycine
1.5
1:100
1
Intestinal
1.75 g bile, 0.5 g pancreatin
7.0
1:100
4
Gastric
10 g pepsin, 8.77 g NaCl
1.8
1:150
1
Intestinal
3.5 g bile, 0.35 g pancreatin
5.5
1:150
1
Gastric
1.25 g pepsin, 0.5 g sodium malate,
0.5 g sodium citrate, 420 µl lactic acid,
500 µl acetic acid
2.5
1:100
1
Intestinal
1.75 g bile, 0.5 g pancreatin
7.0
1:100
4
Gastric
1 g pepsin, 3 g mucin, 2.9 g NaCl, 0.7
g KCl, 0.27 g KH2PO4
2.0
1:50
2
Intestinal
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
7.5
1:100
6
IVG
PBET
DIN
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Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
34
Table 12. Methods for the assessment of contaminant bioaccessibility which include esophageal, gastric and intestinal phases.
Phase
Parameter
Bioaccessibility method
A
Esophageal
Gastric
B
a
C
b
Wittsiepe et al. 2001c
Reference
TIM (Minekus et al. 1995)
Target Contaminant
-
PCBs, Lindane
PCDD/F
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 αamylase, 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
Composition (per litre)
Full details not available. Gastric juice
containing lipase and pepsin.
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)
Initial gastric pH decreasing from 5.0 to
3.5, 2.5 and 2.0 after 30, 60 and 90
minutes respectively.
Soil:solution ratio
BARGE; Oomen et al. 2000, 2001
1:25
1:37.5
1:30
d
pH
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.
Phase
Parameter
Bioaccessibility method
A
Intestinal
Duodenal juice : 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
-
1:97.5
1:60
pH
7.2 (NaHCO3)
>5.5
7.5
Mixing
peristaltic
60 rpm
200 rpm
Residence time
6 hr
2 hr
3 hr
Soil:solution ratio
b
Full details not available. Intestinal
juice containing porcine bile and
pancreatin.
The TIM bioaccessibility method developed by Minekus et al. 1995 is a dynamic system.
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.
b
C
d
Duodenal juice : 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
Composition (per litre)
3 sections, duodenum (pH 6.5),
jejunum (pH 6.8) and ileum (pH 7.2)
a
B
c
Table 13. Methods for the assessment of contaminant bioaccessibility which include gastric and intestinal phases.
Phase
Parameter
Bioaccessibility method
D
Gastric
Ruby et al. 2002
Hack & Selenka 1996
Wittsiepe et al. 2001
Target POP
Dioxins, furans
PAHs, PCBs
PCDD/F
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
Soil:solution ratio
1:100
1:120
1:105
pH
1.5 (HCl)
2.0 (HCl)
2.0 (HCl)
Mixing
30 rpm
200 rpm
200 rpm
Residence time
1 hr
2 hr
2 hr
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
Soil:solution ratio
1:100
1:120
1:120
pH
7.2 ± 0.2 (NaOH)
7.0 (NaHCO3)
7.0 (NaHCO3)
Mixing
30 rpm
200 rpm
200 rpm
Residence time
4 hr
6 hr
6 hr
a
a
F
Reference
a
Intestinal
E
Information in parenthesis indicates method of pH adjustment
Table 13 (cont.) Methods for the assessment of contaminant bioaccessibility which include gastric and intestinal phases.
Phase
Parameter
Bioaccessibility method
G
Reference
Holman 2000
Target POP
Gastric
Intestinal
a
H
a
PAHs
Composition (per litre)
-
I
DIN 2000
Van de Wiele, Verstraete & Siciliano
2004 (SHIME)
Organics
PAHs
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
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
Soil:solution ratio
1:50
1:100
1:15
b
pH
6.5-7.3
7.5 (NaHCO3)
6.3 (NaHCO3)
Mixing
20 rpm
200 rpm
150 rpm
Residence time
4 hr
6 hr
5 hr
Information in parenthesis indicates method of pH adjustment.
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.
.
e
The 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.
b
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
CRC CARE Technical Report no. 14
Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
39
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
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Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
40
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.
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Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
41
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
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42
(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.
CRC CARE Technical Report no. 14
Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
43
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: 4.67.4) 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 540%. 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: 4.67.4) 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 ITEQ.
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
a
See 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)
Herbicide (18)
Pesticide (13)
Mine waste
(4 composite
waste types from
60 samples)
Mine waste (9)
22-1345
39-3034
180±50 1340±720
Mine waste (8)
Geogenic (11)
606-12781
13-422
Industrial (2)
In vitro method
(see Tables 11 and
12)
SBRC Gastric
As
bioaccessibility
Reference
6-89%
9-89%
3.9±0.5 8.5±4.35%
Juhasz et al. 2007a
Diacomanolis, Ng &
Noller 2007
Bruce 2004
PBET
7.1±1.7 –
10.2±2.1%
5-35%
1-22%
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
PBET
0.5-42%
6.8-16.7%
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-gastricintestinal
10-34%
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
PBET
210-2570
1280-204500
Smith, Weber &
Juhasz 2009
0.6-61.1%
Button et al. 2009
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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
Mine waste (4)
1030-5820
PBET Gastric
PBET Intestinal
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
Mine waste (7)
1388-10230
PBET Gastric
PBET Intestinal
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
Mine waste (1)
3060
PBET
DIN
BARGE
SHIME
TIM
Denys et al. 2007
15-56%
5-25%
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)
612-6380
550±80 5450±2700
56, 91%
23, 40%
16, 31%
29,66%
1, 4%
4, 13%
10.2±1.5 –
13.4±2.6%
Oomen et al. 2002
Mine waste
(4 composite
waste types from
60 samples)
Mine waste (9)
SBET
DIN (unfed)
DIN (fed)
BARGE
SHIME
TIM
PBET
PBET
3.7±1.7 –
24.2±4.8%
Bruce 2004
170-12100
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Contaminant bioavailability and bioaccessibility. Part 1: A scientific and technical review
Diacomanolis, Ng
& Noller 2007
46
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).
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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
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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 arsenicspiked 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).
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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
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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.
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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.
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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 monooxygenases).
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.
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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.
Collect soil <2mm size
Contaminant analysis
Contaminant
concentration
> HIL?
No
Report no risk
Assume 100%
bioavailability
Yes
Yes
In vivo – in vitro
correlated method
available for estimating
contaminant
bioavailability?
In vivo
assessment
cost
prohibitive?
No
No
Yes
Sieve soil to <250 µm
Sieve soil to <250 µm
Contaminant analysis
Contaminant analysis
In vitro assessment
using validated method
In vivo assessment
Exposure assessment
Risk characterisation
Figure 4. Schematic diagram for the determination of contaminant bioavailability and
bioaccessibility for human health risk assessment.
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