Procedural Review of Health Reference
Values Established by enHealth for PFAS
Prepared by
Prof (Adj) Andrew Bartholomaeus
School of Pharmacy
Faculty of Health
University of Canberra
&
Therapeutic Research Centre
School of Medicine
University of Queensland
30 August 2016
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Contents
1
Terms of Reference ......................................................................................................................... 3
2
Background – The enHealth Process .............................................................................................. 4
3
Comments on Administrative Aspects of the enHealth Process .................................................... 8
4
EFSA and US EPA Assessments ....................................................................................................... 9
4.1
5
4.1.1
What are HRVs ................................................................................................................ 9
4.1.2
Balancing Risk.................................................................................................................. 9
4.1.3
Status of Current Human Health Reference Values (HRVs) .......................................... 10
4.1.4
The Basic Process for Conducting HHRAs ..................................................................... 11
4.1.5
Basic Issues in Assessment of Epidemiology Studies .................................................... 13
Sources of variation between enHealth Workshop, EFSA and US EPA Risk Assessments............ 16
5.1
Toxicology and Selection of the PoD..................................................................................... 16
5.1.1
PFOS .............................................................................................................................. 17
5.1.2
PFOA.............................................................................................................................. 20
5.2
Toxicokinetics ........................................................................................................................ 22
5.3
Mechanisms of Action........................................................................................................... 24
5.3.1
Pharmacokinetics .......................................................................................................... 24
5.3.2
Toxicity .......................................................................................................................... 24
5.3.3
Comment....................................................................................................................... 25
5.4
6
Preliminary Observations........................................................................................................ 9
Epidemiology......................................................................................................................... 26
5.4.1
Exposure........................................................................................................................ 26
5.4.2
Carcinogenicity .............................................................................................................. 26
5.4.3
Reproductive effects ..................................................................................................... 27
5.4.4
Other Effects ................................................................................................................. 29
Conclusions ................................................................................................................................... 30
6.1
Use of International Risk Assessments and derived HRVs .................................................... 31
6.2
Sources of Differences Between US EPA and EFSA Risk Assessments .................................. 31
6.3
Potential Public Health Consequences of the Choice of HRVs ............................................. 31
6.4
Balancing Risk Mitigation with Risk Generation ................................................................... 32
6.5
Overall Conclusion ................................................................................................................ 32
7
Recommendations ........................................................................................................................ 32
8
References .................................................................................................................................... 33
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1 TERMS OF REFERENCE
In early August 2016 the Department of Health commissioned this review in fulfilment of the
Government’s commitment to review the interim human health reference values (HRVs) for per- and
poly-fluorinated alkyl substances (PFAS) in drinking water. Delivery of the review was therefore
considered urgent and a period of one month was allocated.
The Terms of Reference endorsed by the Government for this independent review are as follows:
“The independent review will consider:
(1) Approaches and assumptions used by the European Food Safety Authority (EFSA), as
outlined in the reports Perfluorooctane sulfonate (PFOS), perfluorooctanoic acid (PFOA)
and their salts, Scientific Opinion of the Panel on Contaminants in the Food Chain (EFSA,
2008) and Perfluoroalkylated substances in food: occurrence and dietary exposure
(EFSA, 2012).
(2) Approaches and assumptions used by the United States Environmental Protection
Agency (US EPA), as outlined in the 2016 Health Effects Support Document for
Perfluorooctane Sulfonate (PFOS) (US EPA, 2016b) and the 2016 Health Effects Support
Document for Perfluorooctanoic Acid (PFOA) (US EPA, 2016a).
(3) The applicability and relevance of these approaches and assumptions in the Australian
context, having regard to existing Australian regulatory science policy as described in
such guidance materials as:
a. Australian Pesticide and Veterinary Medicines Authority (APVMA) Data guidelines
(http://apvma.gov.au/registrations-and-permits/data-guidelines) and Application of
science to regulatory risk assessment (http://apvma.gov.au/node/15486)
b. the enHealth Environmental Health Risk Assessment, Guidelines for Assessing
Human Health Risks from Environmental Hazards (enHealth, 2012);
c. the Food Standards Australia New Zealand (FSANZ) Risk Analysis in Food Regulation
publication: (http://www.foodstandards.gov.au/publications/riskanalysisfood
regulation/Pages/de fault.aspx (FSANZ)
d. the National Industrial Chemicals Notification and Assessment Scheme (NICNAS)
Handbook for notifiers: https://www.nicnas.gov.au/regulation-andcompljance/nicnas-handbook (NICNAS)
e. the National Health and Medical Research (NHMRC) Guidelines for Managing Risks in
Recreational Water (NHMRC, 2008) and NHMRC Australian Drinking Water
Guidelines (NHMRC, 2016).
Given the limited time available, and the considerable body of documentation, the primary focus of
this review is to identify the principle sources of variation between the US EPA and EFSA risk
assessments of PFAS and the resultant guidance values, to comment on the consistency of the
approaches taken with Australian guidance on, and practice of, health risk assessment and to form a
view on the suitability of the EFSA values selected by enHealth as an interim measure pending more
extensive consideration by FSANZ. In the time available for this review it is not possible to
definitively identify one or other of the approaches as “correct” and the other not. Rather, the
potential sources of strength and weakness in each assessment are examined together with a
consideration of the nature and significance of methodological deviations from general regulatory
approaches.
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2 BACKGROUND – THE ENHEALTH PROCESS
The Standing Committee on Environmental Health (enHealth) under the guidance of the Australian
Health Protection Principle Committee (AHPPC) provides nationally agreed environmental health
policy advice, based on the best available evidence and expertise, to the Australian Health Ministers
Advisory Council (AHMAC) through the AHPPC. The committee consists of representatives from key
Commonwealth departments, each of the States and Territories, and New Zealand. In addressing
specific public health issues enHealth draws on specialist scientific and medical expertise through
the establishment of working groups and or the programming of workshops where the issue can be
discussed in detail, applying a multidisciplinary approach. Unlike Australian regulators of chemicals
in food, FSANZ and APVMA, and international agencies such as the US EPA, US FDA or EFSA,
enHealth is not supported by a specialist scientific secretariat (that is a risk assessment group) and
therefore relies on members of its scientific workshops and the agencies of the enHealth
membership to prepare background papers for consideration by the medical and scientific experts of
its scientific workshops.
On 15 March 2016, the AHPPC endorsed the enHealth Guidance Statements on Perfluorinated
Chemicals (PFCs), which include an undertaking by enHealth to convene an expert group, in early
2016, to provide advice to the AHPPC on the development of an Australian interim HRV for
perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) for consistent use in the
undertaking of human health risk assessments and the management of contaminated sites across
Australia. In developing the interim HRV, the workshop was to consider relevant international
guidelines, as well as contemporary scientific and technical issues. This workshop was convened in
April 2016 and provided recommendations for the establishment of HRVs to support jurisdictional
responses to incidents of environmental contamination with PFAS and to set drinking water
guideline values for these substances. The enHealth committee was aware that a number of
international regulatory bodies and the OECD had previously considered the HRVs for PFAS based on
access to a larger data base of published and unpublished studies than was readily available in
Australia at short notice. EnHealth was also aware that there was an immediate need for HRVs to be
identified for use by State and Territory environment agencies in the management of PFAS
contamination of ground water and of food produced in contaminated areas. EnHealth therefore
undertook to review overseas Human Health Risk Assessments (HHRA) and standards rather than a
de novo assessment, and based on a consideration of these, determine temporary/interim
Australian HRVs and drinking water guideline values for PFOS, PFOA and related substances. This
view was reinforced by the knowledge that the US EPA values were then in draft, that it was
understood that EFSA intended that their values would be under review, and that a more detailed
review of HRVs for PFAS was due to commence in FSANZ.
The workshops comprised of recognised experts from a range of scientific disciplines of direct
relevance to the objectives of the workshop and involved a consideration of international
assessments with a principle focus on those of EFSA and the US EPA. Attendees included
toxicologists, members of the enHealth committee, representatives of the Cooperative Research
Centre for Contamination Assessment and Remediation of the Environment (CRC CARE), FSANZ, the
Australian Government Department of Health and Australian Government Department of the
Environment. Scientific and medical experts on the workshop were tasked with preparing and
presenting papers on the toxicology and health effects of PFAS, and the differences in approach of
the US EPA and EFSA for discussion, reflecting their respective expertise.
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The international PFOS and PFOA HRV reviews considered by the workshop included;
(1) European Food Safety Authority (EFSA, 2008);
(2) United States Environmental Protection Agency (USEPA);
a. Health Effects Support Document for PFOS 2014 (draft) (US EPA, 2014a);
b. Health Effects Support Document for PFOA 2014 (draft) (US EPA, 2014b);
c. Provisional Health Advisories for PFOA and PFOS 2009 (US EPA, 2009b);
d. Soil screening levels for PFOA and PFOS 2009 (US EPA , 2009a);
(3) United States Agency for Toxic Substances and Disease Registry (ATSDR, 2015) (draft);
(4) Danish Ministry of the Environment (Perfluoralkylated substances: PFOA, PFOS and
PFOSA: Evaluation of health hazards and proposal of a health based quality criterion for
drinking water, soil and groundwater., 2015);
(5) German Ministry of Health Drinking Water Commission and Federal Environment
Agency, 2006 (GDWC, 2006);
(6) Swedish Environmental Protection Agency, 2012 and 2014 (cited in Danish Ministry of
the Environment document);
(7) United Kingdom Committee on Toxicity of Chemicals in Food (COT) (COT, 2006a; COT,
2006b; COT, 2009; COT, 2014);
(8) Minnesota Department of Health (MDH, 2009a; MDH, 2009b); and
(9) CRC CARE 2016 (draft).
During consideration of the available reviews and established HRVs (Table 1), and the approach and
assumptions utilised by the various regulatory agencies, the workshop engaged in extensive
discussion of the relative merits and weaknesses of those assessments. Individual experts of the
workshop identified potential strengths and deficiencies in both the assessment by EFSA and that of
the US EPA. Although the workshop identified a number of issues impinging on the establishment of
appropriate HRVs and discussed specific recommendations for these values there was not a final
consensus on the treatment of uncertainties and the relative merits or otherwise of the approaches
of EFSA and the US EPA.
The time and resources available to the workshop to explore these issues was necessarily limited,
and enHealth were aware of the pending thorough review of PFAS HRVs by FSANZ. In this context
the workshop noted that;




Because of the exceptionally long half-life (time required for blood levels to decrease by half
once dosing has ceased) of PFAS in humans the systemic (i.e. internal) exposure to PFAS is
determined by oral (or other routes of) exposure over long periods of time. As lowering of
HRVs and drinking water guideline values cannot therefore affect internal exposures
meaningfully over the short to medium term, and given the steps already taken to reduce
exposure in affected communities, lowering the HRVs established by EFSA would have no
short term impact on public health.
The establishment of interim HRVs substantially lower than those of EFSA had the potential
to greatly constrain the FSANZ review.
The EFSA HRVs were therefore concluded to be adequately protective for short to medium
term exposures as a temporary measure.
The FSANZ review would have greater scope in terms of time and resources than the
enHealth workshop and could draw on the deliberations of the workshop to inform that
review.
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The enHealth committee therefore determined to utilise the EFSA HRVs as temporary (i.e. interim)
values pending the finalisation of the FSANZ review.
The enHealth committee, meeting on 6 April 2016, considered the outcome of the technical
workshop from the day before and agreed to have a workshop report prepared and, based on a
consideration of this, to make recommendations to AHPPC. On the 26th of May 2016 enHealth met
again by teleconference to consider a draft workshop report and consider their next steps. The
record of outcomes states:
“Members noted the draft Workshop report prepared by [one of the invited experts] and
that comments and suggested edits were in the process of being incorporated.
“Members noted that the workshop report may take longer than was desirable to finalise,
and so agreed that enHealth proceed instead with a short statement on recommended
interim human health reference values (the “enHealth statement”). A first draft of the
statement was provided by SA Health. The draft set out, as a summary, many of the issues
discussed at the workshop.”
EnHealth made the following decisions;
1) Adoption of Tolerable Daily Intake (TDI) values derived by the European Food Safety
Authority (2008).
Members agreed that the EFSA approach is acceptable. Members further agreed
that the US EPA health advisories on perfluorooctane sulfonate (PFOS) and
perfluorooctanoic acid (PFOA), released in May 2016, should be considered and a
response included in the statement.
2) Adoption of the same TDI value proposed for PFOS for perfluorohexane sulfonate (PFHxS).
Given the comparative toxicity, members agreed that the PFOS TDI value be
adopted also for PFHxS. The practical effect of this is that in applying the TDI for
PFOS, any PFHxS present will also need to be taken into account.
3) Adoption of interim drinking water guideline values for PFOS and PFOA.
Members agreed to adopt interim drinking water guideline values based on the EFSA
TDIs and application of the methodology used in the Australian Drinking Water
Guidelines (ADWG). These would be used for site-specific assessments. Members
discussed the desirability of the NHMRC undertaking a formal process for
establishing guideline values in the ADWG and agreed that this should be considered
further following the completion of the work FSANZ is undertaking to establish
health based guidance values for promulgation in the Australia New Zealand Food
Standards Code.
4) Adoption of interim guideline values for surface water (recreational water and fish
consumption).
Members agreed to adopt interim water quality guideline values for recreational
water based on the approach recommended by the NHMRC, effectively 10 times the
value of the drinking water guideline values.
Members agreed that the assessment of any health risks from contaminated
seafood would be based on the levels detected in the seafood. As such, there was
no requirement to set a level in water from which seafood is taken.
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5) Adoption of seafood screening guideline values.
Members agreed that the interim TDIs would form the basis of site –specific risk
assessments. As such, general seafood screening guideline values were not
required.
6) Updating the enHealth Guidance Statements and Factsheet.
Members agreed that the Guidance Statements and Factsheet would be updated to
include: (i) a reference to the enHealth statement on interim TDIs and water quality
guideline values; and (ii) information about PFHxS.
Members also agreed to develop a “Questions and Answers” guide to facilitate
consistent responses by jurisdictions to media enquiries about the interim reference
values and the inclusion of PFHxS.
7) Changing references from “perfluorinated chemicals (PFCs)” to “per- and poly-fluoroalkyl
substances (PFAS)”.
Members agreed to adopt the change in nomenclature for this group of chemicals.
8) Seek AHPPC endorsement of the enHealth statement on interim TDIs and updated Guidance
Statements and Factsheet.
Members discussed the next steps and agreed that the enHealth statement and
updated Guidance Statements and Factsheet should be provided to the AHPPC for
consideration and endorsement.
A finalised record of the outcomes of the workshop was not available at the time of this review,
however an uncirculated draft outcomes document was provided. The draft record of outcomes
indicates that the workshop gave detailed consideration to the principal studies, points of departure,
toxicological endpoints and uncertainty factors used to derive the HRVs of the different agencies
considered. As many agencies had based their own reviews on that of EFSA, ATSDR or US EPA the
workshop gave most attention to these. The workshop also carefully considered the different
modelling approaches used by the US EPA and EFSA, as discussed later in this report.
As the EFSA and US EPA HRVs represent the upper and lower bounds of the HRVs considered by
enHealth the focus of this review (and the terms of reference) is a comparison of those assessments
and the enHealth deliberations on them. The US EPA has subsequently finalised its assessment of
PFAS and the resultant HRVs. The terms of reference for this review directs attention to the finalised
US EPA reviews of 2016. Consequently, this review considers the finalised US EPA review in place of
the Draft available to enHealth at the time it completed its considerations.
Subsequent to the completion of the enHealth consideration, FSANZ has commenced a separate
review of PFAS HRVs and will consider whether Maximum Levels (i.e. permissible levels) or some
other guidance, should be set for PFAS in food. Additional information available subsequent to the
enHealth assessment, the outcomes of the enHealth workshop, and the recommendations of this
review will be available for consideration in the FSANZ review. FSANZ will also have the discretion to
utilise experts in specific aspects of the PFAS as required.
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Table 1 Health Reference Values for PFOS and PFOA from International Regulatory Agencies
EFSA 2008
USEPA 2016
ATSDR 2015
Danish EPA 2015
USEPA DWG 2009
Minnesota 2009
Germany 2006
PoD
mg/kg/d
0.03
0.00051#
0.00252#
0.033
0.03
0.0025#
0.025
PFOS
UF
200
30
90
1230
390
30
300
TDI/RfD
ng/kg/d
150
20
30
30
80
80
100
PoD
mg/kg/d
0.3
0.0045#
0.00154#
0.003#
0.46
0.0023#
0.1
PFOA
UF
200
300
90
30
2430
30
1000
TDI/RfD
ng/kg/d
1500
20
20
100
190
77
100
# Based on the human equivalent dose derived from theoretical pharmacokinetic modelling and incorporating a variety of
assumptions to compensate for data deficiencies.
3 COMMENTS ON ADMINISTRATIVE ASPECTS OF THE ENHEALTH PROCESS
EnHealth provides a highly valuable and appropriate consultative jurisdictional forum to support coordinated approaches to environmental health issues across the Commonwealth and New Zealand
(NZ). The use by enHealth of expert working groups to provide specialist medical and scientific
expertise to support its work is consistent with that of most major international regulatory agencies,
including EFSA and the US EPA, with the Australian Therapeutic Goods Administration (TGA), and to
a degree, that of FSANZ and the APVMA. Where the enHealth process deviates is that it does not
include a dedicated, specialist scientific secretariat (a risk assessment group) to support its work.
Most regulators utilise a scientific secretariat to prepare initial risk assessments that address the
available data, draw conclusions, make recommendations and pose specific scientific questions for
the expert committee to address. The expert committees then address the specific
recommendations and questions, critique the overall assessment and provide recommendations for
further work. This process ensures that expert committees have all the necessary information at
their disposal, and can then focus their individual attention and their available time on the specific
aspects of the assessment that their specialist expertise can most add value to. The advantage of
this approach is that the decision making process is considerably more transparent, the available
expert resources are utilised efficiently, the scientific line of reasoning is documented in detail,
approaches to HHRA are consistent, and where appropriate the engagement of external
stakeholders is facilitated through public consultation processes such as those of EFSA, US EPA and
FSANZ.
The absence of a finalised, ratified outcome report from the expert workshop at the time the current
review of the process commenced is notable. The draft report provided is also notably brief given
the complexity of the issues and the detail in which they were addressed by the workshop
participants, and clearly reflects the absence of the support of a suitably capable expert scientific
secretariat. The approach taken by enHealth in this specific case however, recognised the clear need
to provide urgent guidance to the various State and Territory environment agencies to support their
ongoing remediation and mitigation efforts. The draft report indicates that the workshop gave
careful consideration to the sources of the variations between EFSA and US EPA reference values
and considered the strengths and weaknesses of the approach taken by each agency. Additionally,
the principle and central outcome of immediate importance from the workshop, that the EFSA
values were appropriate as a temporary measure, was agreed by the participants of the workshop
on the 5 April. A clear basis for the decision of enHealth to utilise the EFSA HRVs was provided by
the experts and enHealth committee members constituting the workshop of 5th April 2016.
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Thus, although there are aspects of the process that could be improved these do not reflect on the
suitability or otherwise of the EFSA HRVs for PFAS as a temporary/interim measure pending formal
review by FSANZ.
4 EFSA AND US EPA ASSESSMENTS
4.1 PRELIMINARY OBSERVATIONS
In order to understand the different approaches applied by the US EPA and EFSA to their risk
assessment of PFAS, and more specifically the HRVs they established, some general principles and
background information first need to be discussed.
4.1.1 What are HRVs
Health Reference Values, or HRVs, come in many forms and are intended to cover a range of
exposure scenarios. In the context of the considerations by enHealth of PFAS, the most common
HRVs are Tolerable Daily Intakes (TDIs), Acceptable Daily Intakes (ADIs) and Acute Reference Doses
(ARfDs). ADIs and TDIs are established to represent the maximum intake of a substance, whether
naturally occurring or synthetic, that can be ingested by the population every day of their entire
lifetime without appreciable risk. An ADI is used for substances intentionally added to food and a
TDI is used for contaminants that may be naturally present in the agricultural environment or water
source, or are anthropogenic contaminants. Because the ADI or TDI are maximum average daily
intakes an ARfD may be established for acutely toxic substances to represent the maximum amount
of a substance that can be safely ingested in a single day or a single meal/drink.
These values are not, and are not intended to be, bright lines between safety and risk, but rather
represent the limit of confidence in the safe intake level. The greater the long term exposure of an
individual exceeds the TDI, the more likely that some risk will be associated with that exposure.
Because TDIs specify safe daily intakes for a lifetime of exposure, even quite substantial exceedance
of the value for short periods is not generally associated with increased risk. If realistically
achievable short term exposures above the TDI are considered likely to present a risk, an ARfD is
established to place an upper boundary on the safe daily intake.
In circumstances where data are incomplete or a significant degree of uncertainty applies to the
derivation of a TDI, a provisional TDI or PTDI may be established to reflect this.
4.1.2 Balancing Risk
Human Health Risk Assessment (HHRA) and the establishment of HRVs involves elements of both
science and policy. As a matter of policy the process assumes that any uncertainties between the
available data and their relevance to human health result in the general human population being
more sensitive than the species or population from which the data are derived unless there is strong
evidence to the contrary. This is a precautionary approach, and there is no evidence that humans
are routinely more sensitive than experimental animals, for example. Indeed, because of the
extreme conditions of exposure and a range of physiological differences between experimental
animals and the general human population, humans are frequently less sensitive to toxicological
effects. As a consequence, the HRVs established through these processes are frequently likely to be
lower than the true tolerable intake level. Nonetheless, for chemicals intentionally added to food
either during production or processing, this cautious approach provides additional protection
without generating appreciable collateral health risks. This is not always the case however.
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There are circumstances where overly conservative or precautionary processes for establishing HRVs
can generate health risks greater and more probable than those initially intended to be avoided. For
human therapeutics an excessively cautious approach may preclude the availability of life saving
medicines for example, and consequently the safety assessment adopts a risk benefit or Margin of
Exposure approach. For natural contaminants in food, such as some heavy metals, an excessively
cautious approach may impact food security and nutritional adequacy, and care is therefore taken to
balance those risks within the HHRA through progressive refinement of underlying assumptions.
For anthropogenic contaminants such as PFAS which cannot readily be removed from the
environment the establishment of values that are, with respect to the overall weight of evidence,
disproportionately low, has the potential to result in a range of adverse health outcomes which may
be greater than the toxicological risks intended to be avoided. Such outcomes may include
prolonged unwarranted stress in exposed populations, the recommendation, or seeking out, of
unnecessary medical interventions with their attendant risks, or interventions in pregnancy and
avoidance of breast feeding to the detriment of the foetus and neonate. Other, economic impacts
although likely, are beyond the scope of a HHRA and of this review. Simplistic selection of the
lowest international HRV is therefore not necessarily optimal for the overall protection of public
health. Determination of suitable HRVs for PFAS requires a careful consideration of the strengths
and weaknesses of the approaches taken by international agencies that have had access to the
underlying data and a considered selection of the most appropriate approach/values within the
context of the exposure patterns in Australia. A suitably precautionary approach to public health
requires a balancing of risk prevention against the potential for risk generation.
4.1.3 Status of Current Human Health Reference Values (HRVs)
For any substance, Human Health Risk Assessments (HHRAs) are an ongoing, iterative process. As
new data become available they are incorporated into the risk assessment and may over time alter
the established HRVs. These values are as likely to increase as to decrease depending on the nature
of the data generated and their impact on the magnitude and direction of uncertainties. For PFAS
the EFSA and US EPA reviews are now finalised but EFSA has indicated an intention to review their
values in the near future and FSANZ is currently in the early stages of a review of the HRVs and the
need for permitted levels appropriate for food and bottled water. FSANZ will have the opportunity
to consider the recommendations of this review, the deliberations of the expert workshop convened
by enHealth, together with any new data that become available, and any revisions of HRVs and
standards by international regulators.
The importation and use of PFAS has been progressively reduced in Australia over the past decade
and exposure of the general community will also progressively decrease. For communities around
point sources of contamination, such as defence bases where PFAS containing fire-fighting foam has
been used, an expanding plume of contaminated ground water may result in a transient increase in
exposure where ground water is used domestically or for food production. As such plumes move
progressively further from the point source, concentrations in ground water decreases through
dilution and consequent exposure also decreases. Provision of uncontaminated water for domestic
use in these areas will greatly decrease exposure.
Internationally where the use of PFASs has become restricted, a general trend towards progressively
lower PFAS serum levels has been observed reflecting a progressive reduction in exposure and a
slow but continued elimination of PFAS.
Because PFAS have an exceptionally long half-life in human blood, the primary determinant of
ongoing exposure is the existing blood level and not the daily intake (other than the unlikely scenario
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of intake of aberrantly high levels of PFAS through non environmental sources). Consequently, even
quite substantial differences in permitted levels in drinking water for example, do not have a
significant short term impact on systemic (i.e. internal) exposure.
4.1.4 The Basic Process for Conducting HHRAs
The basic process for human health risk assessment is essentially the same across every major
chemicals regulator in Australia and internationally. The process involves elements of policy,
convention and science and is designed to be inherently risk averse or precautionary. All aspects of
the process however, are subordinate to the best available science and new approaches to
maximising the scientific basis for decision making, and new understandings of mechanisms affecting
cross species extrapolation, in HHRAs are continuously being refined. Consequently, deviation from
a predefined process in order to add new understandings or a more physiologically based method of
reducing uncertainty is not inappropriate simply because it might deviate from that process,
provided the validity of the process is sufficiently robust that it does not simply replace one source
of uncertainty with another.
Toxicology data obtained in animals or in humans through direct experimental testing or
epidemiology studies is analysed to identify the most sensitive toxicological effect in the most
sensitive species. The dose immediately below the dose at which this effect is observed is then the
Point of Departure (PoD), the dose point at which the process departs from analysis of the toxicology
data to an estimation of safe exposure levels for the general human population. The study or
studies producing the PoD is often referred to as the pivotal study or studies. The lowest dose
producing an adverse effect is called the Lowest Observed Adverse Effect Level (LOAEL) and the dose
immediately below that dose is called the No Observed Adverse Effect Level (NOAEL).
There are two key sources of variability in this portion of the process. The first of these is in
determination of the adversity of an observed effect in an animal. And the second is identification
of effects in animals that are species specific and therefore not, or unlikely to be, relevant to humans
where the physiology is different and the mechanism of toxicity does not apply.
Various physiological adaptive mechanisms may be stimulated by exposures that are either not
adverse, or are not adverse at low levels of stimulation. Similarly, although experimental animals
share the majority of physiological processes with humans there are important biochemical and
anatomical differences that can render an effect in an animal irrelevant for human risk assessment,
or indicate that studies in a different species that is a better model for human responses, is more
appropriate as the basis for identifying a PoD for that particular process. The conclusions from a
consideration of these issues depend on a number of factors, both policy and science based,
including the expertise and experience of the evaluator(s), the availability of data informing the
consideration and the balancing of the weight of evidence. For natural and anthropogenic
environmental contaminants integration of epidemiology and toxicology may provide indications of
the likely relevance of observations in animals to the general population, or indicate that the animal
studies have over predicted the likely human sensitivity to the contaminant.
Various agencies make more or less use of various modelling techniques such as Benchmark Dose
Modelling (BMD) to assist in identification of the PoD, although this is generally confined to
circumstances where a clear NOAEL is not obtained (uncommon for OECD test guideline compliant
studies). Where a study does not provide a NOAEL, because effects are seen at every dose, and
there are 2 or preferably 3 doses producing an effect and defining a well characterised dose
response curve, the BMD approach may constitute a useful addition to the risk assessment process.
In other circumstances the BMD approach uses predefined, somewhat arbitrary, dose response
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curves, tends to give values not dis-similar to the NOAEL/LOAEL approach, and frequently adds little
of value to the process, whilst reducing the transparency of the assessment for audiences other than
experienced toxicologists. Differences in the value of the PoD, generally small, may arise where a
BMD approach is used by one agency and the NOAEL by another.
Where a PoD is taken from a human epidemiology study the uncertainty inherent in extrapolating
from experimental animals to humans is removed, but often greater uncertainty arises in the
estimation of the exposures (doses) of the studied population, and the introduction of various
confounding factors that can hinder or preclude meaningful interpretation. The estimates of
exposures and the identification and management of confounding factors can vary substantially.
Once a PoD has been identified the risk assessment gives consideration to the sources and extent of
uncertainties, inherent in any form of extrapolation, and attempts to quantify those uncertainties
into Uncertainty Factors (also often called safety factors or adjustment factors). As a matter of
precautionary policy, rather than science, the uncertainty is always assumed to work to make
humans more sensitive than the most sensitive animal species and the most sensitive individuals to
be substantially more sensitive than the average. The internationally accepted default values are 10
for differences between animals and humans and 10 for the difference between the average and
most sensitive individuals. Additional uncertainty factors may be added to adjust for other sources
of uncertainty such as specific study types that are not available. Similarly, the default uncertainty
factors may be reduced where data are available that indicates that animals are more sensitive than
humans for a specific toxicological effect or data that reduce the uncertainty of extrapolation such as
comparative toxicokinetics between humans and animals. This component is a substantial potential
source of variability in the determination of HRVs between regulatory agencies and is a significant
contributor to the variation in HRVs for PFAS.
A further source of variance between PoDs identified by different agencies is the use of
pharmacokinetic modelling to determine the Human Equivalent Dose (HED). This approach is
common or routine for the US EPA but an exception for most other regulatory agencies. The HED is
the dose that would need to be given to a human to achieve the same blood level as that in the
experimental animals in the pivotal study used as the basis for the PoD. Where comprehensive data
are available for both humans and the experimental animals, this approach can remove substantial
uncertainty in the cross species extrapolation and has considerable potential for improved risk
assessment outcomes. Unfortunately, data are generally incomplete and most modelling must
incorporate a range of assumptions which may create uncertainties equal to or exceeding those
initially intended to be reduced.
Having identified the PoD and the appropriate uncertainty factors, the HRV is determined by dividing
the PoD by the uncertainty factors and expressing the HRV as a tolerable (or acceptable, permissible
etc) daily intake value (TDI) in weight units per kg of body weight per day.
The final step is setting permissible levels of a substance in drinking water or food. This process
involves an estimate of the daily intake of water and food for high consumers from the most
sensitive portion of the population, and a calculation of the highest permissible level that would not
result in an individual exceeding the TDI on an average daily basis over a period of a year. The
principle source of variation in the determination of an acceptable residue level in water is the
determination of the HRV. Estimates of intake for water can vary depending on the climate of the
target population but generally has a small impact on the permitted residue levels.
12 | P a g e
Estimation of permitted levels of a contaminant in food can be substantially influenced by both the
dietary patterns of the target population and by the HRV. For agricultural and veterinary chemicals,
the permitted level (the Maximum Residue Level or MRL) is set at the lowest value consistent with
good agricultural practice (GAP) and is essentially an “As Low As Reasonably Achievable Approach”
(ALARA), where the chemical is used at the minimum level to achieve the required effect. If the
residue level of the chemical when used in accordance with GAP is too high to be safe, then the use
is not permitted.
Drinking water guidelines are generally viewed as limit values rather than target values and the
principles of ALARA are equally applicable. Climate, water sources and other factors will affect what
constitutes “Reasonably Achievable” at any given time or location.
As the principle source of variation in permitted levels in water derived by EFSA and the US EPA is
the derivation of the HRVs, rather than the exposure assessments, the latter are not further
addressed in this review.
4.1.5 Basic Issues in Assessment of Epidemiology Studies
Epidemiology studies provide a valuable source of information for use in HHRAs. Because data are
derived from humans the uncertainty inherent in extrapolating results in experimental animals to
humans is avoided. Conversely however, the experimental conditions of an epidemiological study
cannot generally be controlled and manipulated to the extent possible in toxicology studies and a
range of uncertainties related to the study design, estimation of exposures, confounding by coincident exposures, life style and other factors frequently create uncertainty of similar magnitude to
that in cross species extrapolation. Additionally, a large proportion of epidemiology studies involve a
substantial number of comparisons which generates the probability that statistically significant
associations between a presumed level of exposure and a disease outcome will arise purely by
spontaneous random variation and not through a cause and effect mechanism. The generally
accepted basis for interpreting the results of epidemiology studies in terms of causation are the
criteria first put forward by Bradford-Hill (The Environment and Disease: Association or Causation,
1965).
Epidemiology studies fall in to two broad categories: observational and experimental, or
intervention, studies. For the purposes of HHRA of environmental contaminants the study design
most commonly available is the observational analytical study, because intervention studies where
exposures are manipulated are generally not ethical. An observational study is one where the
researcher has no control over circumstances within which events occur. These studies are further
divided into descriptive and analytical study types:
a. In a descriptive study the researcher collects data to describe or characterise the
disease, pathology, event or condition of interest in terms such as time, location,
population, and progression;
b. In analytical epidemiology the researcher will seek to identify risk factors or causes of a
particular pattern of outcomes, such as disease, by comparing different groups.
Although this review cannot provide a comprehensive discussion of the basis for interpretation and
critique of epidemiology studies, a consideration of both the Bradford Hill criteria and the statistical
insights of Ioannidis (2005), discussed further below, illustrates the key issues of relevance to PFAS
risk assessment that underpin the EFSA and US EPA assessments.
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4.1.5.1
Assessing causality – the Bradford-Hill criteria
The most common errors occurring when non-epidemiologists such as the media interpret or
comment on epidemiology studies is confusion of association with causation and not considering the
available studies collectively. In considering the extent to which a study or data set supports or
indicates a causal relationship between an exposure (or activity or other factor) and a specific
outcome of interest, the risk assessment first needs to exclude non-causal explanations for the
apparent association such as investigator or other sources of bias, potential sources of confounding
and the potential for random or chance occurrence. If no obvious non-causal explanations are
identified, consideration of standardised guidelines for causal inference guides the assessment. The
foundations of this approach were first put forward by Sir Austin Bradford Hill in a speech to the
Occupational Medicine Section of the Royal Society for Medicine in 1965, which are now known as
the Bradford Hill criteria for causation. These criteria, with some examples taken from the report of
the original speech, are:
1)
2)
3)
Strength of Association
a)
Where an association is exceptionally strong such as the many hundred-fold increase in
the incidence of scrotal cancer amongst chimney sweeps observed in the 18th century
by Percival Pott, even the weakest epidemiology study design may be sufficient to
reliably assign causation.
b)
Similarly, the 10-fold to 30-fold increased incidence of lung cancer amongst moderate
to heavy smokers does not require a particularly sophisticated study design in order to
attribute causation. Conversely, the two-fold increase in coronary thrombosis in
smokers is not sufficiently strong, in isolation, to be causally attributed to smoking from
the weaker study designs such as cross sectional studies.
c)
Where the strength of association is low, as is frequently the case in the current day,
better designed studies and or other types of data that allow consideration of the
remaining criteria will be required to form a judgement about causation.
Consistency of Association
a)
Has the association been observed by other investigators in other locations and at other
times and under other circumstances. In essence this criterion is about reproducibility
of the association by other researchers.
b)
The reproducibility of an outcome may however simply reflect a methodological or
confounding factor common to multiple studies by multiple investigators and again
therefore is not sufficient grounds, in isolation, on which to conclude causation.
Specificity of association
a)
4)
Specificity refers to a one to one relationship between a single causal agent and a single
effect. The relationship between chicken pox and the virus which causes it is an
example of specificity. Although a lack of specificity does not necessarily negate a
conclusion of causation, where a single agent causes effects at multiple sites or tissue
types for example, where it is present, it provides solid evidence of causation.
Temporality
a)
In essence this criterion is concerned with the sequence of exposure and effect. For this
criterion to be satisfied, exposure must occur prior to the observed effects and in the
case of diseases with long evolution times, it must occur sufficiently in advance of the
disease, and for a sufficient period, to plausibly have resulted in the disease
development.
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5)
Dose response
6)
a)
The increase in lung cancer death rates with increased cigarette use, for example,
provides strong evidence in support of a cause and effect relationship.
b)
The absence of a dose response in the presence of good evidence of substantial
variation of exposure, or dosage, raises significant doubt over a causal relationship.
c)
However, if the internal dose or the critical dose or concentration at the target affected
by the substance is not directly related to the external or administered dose, a dose
response relationship may not be readily apparent if the administered dose is the only
source of information.
Plausibility
a)
7)
Based on the information we have available, is the relationship between exposure and
effect biologically plausible? Although a lack of knowledge about a plausible
mechanism for a relationship does not necessarily prevent a conclusion of causation,
where that relationship has had very limited investigation for example, if a substantial
body of knowledge about the toxicological mechanism(s) of a substance does exist, the
lack of a plausible mechanistic relationship between the proposed cause and the effect
will weigh against a conclusion of causation. Thus, information sources other than
toxicological and epidemiological sources may need to be sourced.
Coherence
a)
8)
Is the proposed causal relationship consistent with the broader knowledge of the
natural history and biology of the disease?
Experimental evidence
9)
a)
This criterion primarily refers to the availability of experimental/intervention studies
where the causal relationship has been investigated, and supported, experimentally.
b)
Intervention studies that demonstrate a reduction of a disease outcome from removal
or reduction of a postulated causal factor provide strong evidence of a causal
relationship, for example, the reduction in the incidence of cholera after Dr John Snow
removed the pump handle from the Broad Street pump in the St James’s parish in 1854
was strong evidence the contaminated water was causing the disease.
Analogy
a)
4.1.5.2
What do we know about similar/related substances and similar disease or pathology
outcomes that might add or subtract support to the proposed causal relationship?
Design Limitations in Epidemiology
Ioannidis (2005) has discussed the statistical basis for the poor reproducibility of many types of
epidemiology studies in the literature and proposed a number of criteria on which to judge the likely
reliability of the results of such studies. These criteria, in conjunction with those of Bradford Hill,
also provide a useful basis for considering the suitability of an epidemiology study for inclusion in a
HHRA, and the weight that should be given to the results of these studies in forming risk assessment
conclusions. Ioannidis observes that the design factors that tend to lead to unreliable results
include:
1)
small study size,
2)
small effect sizes in relation to background variability,
3)
large numbers of variables being tested that are unrelated to a specific prior hypothesis,
15 | P a g e
4)
a high degree of flexibility in study designs, definitions, outcomes, and analytical modes,
5)
conflicts of interest including; financial, ideological, philosophical, reputational.
There is a degree of overlap between the criteria of Bradford Hill and those of Ioannidis. Strength of
association for example is related to study sample sizes and the size of the effect in comparison to
background variation. The larger the sample and effect sizes the stronger the statistical significance
and apparent association.
Epidemiological studies conducted in the absence of a specific (that is a precise/targeted) scientific
hypothesis will generally employ a wide range of investigative variables and conduct multiple
comparisons in search of “significance”. Studies of this type will frequently yield random statistically
significant findings that are not reproducible in subsequent studies.
5 SOURCES OF VARIATION BETWEEN ENHEALTH WORKSHOP, EFSA AND
US EPA RISK ASSESSMENTS
Primary sources of variation between the EFSA and US EPA assessments include elements of
toxicology, particularly the selection of the PoD, approaches to address the differences in
toxicokinetics, variations in selection and use of uncertainty factors, considerations of the
mechanism of action, conclusions on the epidemiology studies and the use of modelling techniques.
5.1 TOXICOLOGY AND SELECTION OF THE POD
The risk assessments of EFSA and the US EPA each conform to the general approach described
above. As the enHealth process endorsed the EFSA values for temporary use in Australia rather than
a de novo assessment of the data, pending the FSANZ review, the following discussion focuses on
the EFSA and US EPA reviews. The principle sources of variation arise from various aspects of the
interpretation of the available evidence and in the approach to managing uncertainty. The nature of
both the variations and the consistencies between these assessments are discussed below in terms
of the key data sets supporting the determination of an HRV;




animal toxicology studies
toxicokinetics
mechanisms of toxicity
human epidemiology.
A reappraisal of all the individual toxicology studies reviewed by EFSA and the US EPA is beyond the
scope of and time available for, this review. However, a general consideration of the studies
selected for identification of the PoD is important in understanding the source of the different HRVs
derived by the two agencies. Additional studies that investigate the potential for, and mechanism
of, higher (or lower) sensitivity in experimental animals than humans inform the selection of
uncertainty factors and provide some indication of the likely magnitude and direction of
conservatism within the derivation of an HRV.
Importantly the differences between the EFSA and US EPA assessments are not due to new data or
information available to the US EPA that was not available to EFSA. Although there are a small
number of new studies reviewed in the US EPA assessment they have not affected the choice of
pivotal studies for the determination of the PoD.
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Hepatic and developmental toxicity are the most sensitive toxic effects of PFAS in animals and these
effects dominate the pivotal data sets used by both the US EPA and EFSA to derive their respective
PoDs for their HRVs.
In selecting which study or studies should provide the PoD there are two ways to think of dose,
administered dose and internal or systemic dose. For practical reasons related to accessibility of
tissue for sampling, the key comparative measure of systemic exposure is the Area Under the
(plasma level versus time) Curve or AUC. The AUC provides the time weighted average plasma level
achieved from repeat dosing. For most compounds the AUC is directly proportional to the daily
administered dose once plasma steady state has been reached (that is after a period of stabilisation
related to the balance between dose administered, frequency of administration and the rate at
which the body clears the administered substance). For compounds such as PFAS which have
exceptionally long half-lives, and great variability between species, comparing doses on the basis of
the amount administered each day may provide a misleading picture, because plasma steady state
may take many months or years to reach. If data on actual plasma/serum levels is available, a
comparison across species may be more robust if these are used rather than the daily administered
dose. The liver is a possible exception due to its more direct exposure to administered dose rather
than average systemic blood levels, as discussed further later in this review. A further step is to
convert the serum levels in experimental animals to the (theoretically determined) equivalent oral
dose that would need to be given to humans to reach that blood level – called the Human Equivalent
Dose or HED.
As illustrated for PFOS in Table 2, and PFOA in Table 4, the US EPA has used serum level
determinations from experimental animals, in Physiologically Based Pharmacokinetic Modelling
(PBPK) as the basis for a calculation of the HED at the NOAEL and LOAEL. The validity of the
calculated HED is entirely dependent on the validity of the PBPK model and the assumptions built
into those models as discussed later in this review under toxicokinetics. These issues are explored
further for each of PFOS and PFOA below. Identification of the pivotal study or studies to derive the
PoD is a separate and distinct step to the subsequent conversion to a HED. More specifically,
conversion to the HED is not a necessary step in the identification of the PoD.
5.1.1 PFOS
In the Australian context, PFOS is the predominant contaminant of concern. In considering sources
of variation between the HHRAs of EFSA and the US EPA the first step is a comparison of the studies
considered for the derivation of the PoD.
For the purpose of identifying the most sensitive toxicological effect in the most sensitive species the
US EPA ranking of toxicity studies by modelling average serum levels based on measured levels at
termination, at the NOAEL/LOAEL is the correct approach (for effects other than those in the liver) as
it compares studies on the basis of systemic (that is internal dose) rather than the administered
dose, and thereby removes a substantial proportion of the uncertainty in cross species comparison
derived from very different pharmacokinetic behaviours in different species. With the exception of
the liver and GIT tract, the concentration of PFAS in target tissues will be proportional to the serum
levels. For reproduction and developmental studies in particular the evidence indicates that the
ratio between maternal serum and both milk and cord blood are comparable across species inclusive
of humans. Consequently, the “dose” experienced by the target tissue (or foetus/neonate) is
proportional to the maternal serum and not necessarily to the maternal oral dose (except over very
long periods). In this respect the US EPA approach to ranking studies to identify the PoD may be the
more appropriate. As noted by the enHealth workshop however
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“ … to predict the area under the serum concentration time curve (AUC). AUC was divided
by study duration (presumably the same as dosing duration) to provide an average serum
concentration over the study period. The model is based on Andersen et al. (2006), but
further developed by US EPA. It is an empirical model – i.e. a model that includes some
understanding of the processes involved but the estimated values for the parameters which
drive the model are derived by Bayesian statistical techniques (using Markov Chain Monte
Carlo) to optimise the fit of predicted serum concentrations (model outputs) to measured
concentrations in a specific study or set of studies, rather than being determined from a
mechanistic understanding of the kinetic processes. This model has 3 compartments
amongst which the PFOS entering the body can move. Time and concentration dependency
for transfer between compartments was required to replicate experimental serum levels.
This model was developed using data from a monkey study rather than a study in rats. The
US EPA did consider the model parameters relevant for rats and used them during the
modelling.
The Workshop noted it is difficult for an independent third party to replicate the US EPA
PBPK modelling for estimating the average serum concentration in an animal experiment.
Other jurisdictions use serum concentrations actually measured during the experiment. “
Additionally the US EPA might be considered to have given insufficient weight to evidence
supporting the importance of PPAR alpha in mediating developmental effects in rodents, as
discussed later in this review.
For some studies the US EPA choice of NOAEL is questionable. As the enHealth workshop notes the
NOAEL set for the Butenhoff et al study of 0.3 mg/kg bw/day is based on effects seen only on postnatal day 17 but not on days 13, 21 or 61 indicating the effect is unlikely to be treatment related and
not suitable for setting a NOAEL. Although this study was used as the pivotal study in the initial US
EPA draft assessment, in the final report it is only supportive of the Luebker et al study.
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Table 2 EPA Human Equivalent PFOS Doses Derived from the Modelled Animal Average Serum Values
Table 3 Identification of the PoD for PFOS by EFSA and the USEPA
Study
EFSA
Pivotal study
Seacat et al (2002)
US EPA
Seacat et al 2002
Seacat et al. (2003)
Luebker et al (2005b)
Luebker et al (2005a)
Luebkaer et al 2005 a
Butenhoff et al (2009)
Lau et al (2003)
Species, route, Duration
NOAEL, LOAEL
Mg/kg bw/day
Ava Serum level NOAEL,
LOAEL, µg/mL
Monkey, gavage, 183 days
0.03, 0.15
7.8, 38
Monkey, gavage, 186 days
Rat (m), 98
Rat reproduction, 84
Rat reproduction, 63
Rat Reproduction, 63
Rat developmental, gavage,
41
Rat developmental, gavage,
19
0.15, 0.75,
0.34, 1.33
0.1, 0.4
, 0.4
0.4, 0.8
0.3, 1.0
38, 157
16.5, 64.6
6.26, 25
, 25
19.9, 39.7
10.4, 34.6
1.0, 2.0
17.5, 35.1
a
the US EPA calculated the average serum level based on the final level measured at sacrifice and using PBPK modelling.
EFSA used the final serum level. Values for the EFSA NOAEL for their choice of the PoD of 0.03 mg/kg bw/day in the Seacat
et al 2002 monkey study is calculated from the US EPA modelled serum levels at 0.15 mg/kg bw/day (their choice of NOAEL
for this study) and assuming dose proportionality between 0.03 and 0.15 mg/kg bw/day.
Importantly, EFSA have concluded that the NOAEL in a 183-day monkey study reviewed by both
agencies and by the ATSDR, is lower than that identified by the US EPA and ATSDR for this study. If
the serum level for the EFSA NOAEL of 0.03 mg/kg bw/day is calculated from the modelled US EPA
values at 0.15 mg/kg bw/day in the same study, and dose proportionality is assumed across this
range, then the EFSA NOAEL is arguably the appropriate PoD, because the average serum level is the
lowest (or comparable to the lowest) and the monkey is likely to be a better model for human risk
19 | P a g e
assessment than the rat (due to more similar pharmacokinetics and biochemistry). The study cited
by US EPA (Luebker, et al., 2005) which had a NOAEL of 6.26 µg/ml, was reviewed by EFSA but not
considered in selection of a PoD. The EFSA discussion of this study is insufficiently detailed to
identify their reasoning for this. Nevertheless, in terms of internal dose based on average serum
levels of PFOS, there are no material differences in the PoD identified by EFSA compared to the US
EPA. Differences between the two agencies assessments therefore hinge on the validity of the
US EPA PBPK modelling of the human equivalent dose and of the Uncertainty Factors applied by each
agency.
5.1.2 PFOA
The key studies considered by the US EPA and by EFSA for derivation of the PoD for PFOA are
provided in tables 4 and 5.
Table 4. EPA Human Equivalent PFOA Doses Derived from the Modelled Animal Average Serum
Values
The US EPA modelled the HED for each of the studies in Table 5 then applied the relevant
uncertainty factors. Those studies without a NOAEL accrued an additional 10-fold safety factor. On
the basis of these calculations the EPA determined that the pivotal studies were those of Dewitt, Lau
& Butenhoff all of which yielded the same HRV. EFSA did not have the studies by Macon et al 2011
or DeWitt et al 2008 as those became available subsequent to completion of their assessment.
Conversely the US EPA did not have access to the study by Sibinski (1987) reviewed by EFSA. These
differences in data assessed do not affect the derivation of the PoD however, as those studies were
not pivotal, in isolation, to the US EPA decision and in fact the US EPA did not use the Macon study in
its PoD modelling. Rather than apply arbitrary uncertainty factors for studies without a NOAEL, EFSA
utilised benchmark dose modelling to calculate a dose equivalent to a 10 % increase the incidence of
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the effect driving the LOAEL, over control values, and then used a value equal to the lower bound of
the 95% confidence interval. This value is then the BMDL10. For consistency a BMDL10 was
determined for each of the studies considered for derivation of the POD. Although this approach is
not necessarily routine, it is entirely consistent with international practice, appropriate to the
specific dataset available and suitably precautionary, giving values lower than the NOAELs where
those were available.
Table 5 Summary of studies on PFOA considered by EFSA for the PoD compared to those of the EPA
Study
EFSA
Lau et al (2006)
Perkins et al (2004)
Sibinsky (1987)b
Butenhoff et al (2004a)
US EPA
DeWitt et al (DeWitt,
Copeland, Strynar, &
Luebke, 2008)
Lau et al 2006
Perkins et al 2004
Wolf et al 2007
Butenhoff et al 2004a
Macon et al (2011)
Species, route, Duration
NOAEL, LOAEL
mg/kg bw/day
Mice Developmental, Gavage,
GD 1-17
Rat, diet, 13 weeks,
Ava Serum level
NOAEL, LOAEL
µg/mL or BMDL10
-, 1
BMDL10 0.46
0.06, 0.64
BMDL10 0.44
Rat carcinogenicity, diet, 104
1.3, 14.2
weeks (lifetime)
BMDL10 0.74
Rat reproduction, oral gavage, 2 1, 10
generation
BMDL10 0.31
17.5
Mice, gavage, 15 days
1.88, 3.75
38.2, 61.9
Mice developmental, oral
gavage, GD 1-17
Rat, diet, 13 weeks,
Mice developmental, oral
gavage, GD 1-17
Rat reproduction, oral gavage, 2
generation
Mice developmental, oral
gavage, GD 1-17
-, 1
-, 38
0.64, 1.94
-, 3
31.6, 1.94
77.9
-, 1
-, 45.9
-,0.3
-,12.4
26
14
a
the US EPA calculated the average serum level based on the final level measured at sacrifice and using PBPK modelling.
Because all the studies considered for the PoD were in rodents EFSA compared the studies based on oral doses
administered. EFSA however modelled the BMDL10 for each study and used these values in place of the NOAEL for
comparisons. b this report was not available to the US EPA.
The lowest BMDL10 was 0.31 mg/kg bw/day from the Butenhoff study equivalent to an average
serum level of 14 µg/mL based on the US EPA calculations for the same study. In terms of internal
dose expressed as average serum levels of PFOA, EFSA therefore established a PoD comparable to
that of the US EPA.
As for PFOS the differences between the US EPA and EFSA derived HRVs for PFOA is predominantly
dependent on the use of PBPK modelling by the US EPA and the selection of uncertainty factors by
each agency, but with the added complexity of the use of BMD modelling by EFSA to identify the dose
for the PoD.
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5.2 TOXICOKINETICS
The most substantial difference in the US EPA versus EFSA derivations of HRVs is in the use of
Physiologically Based Pharmacokinetic Modelling (PBPK) by the US EPA to determine the HED for
doses used in the animal studies. The ATSDR describe the key aspects of this process as follows:
“PBPK models for a particular substance require estimates of the chemical substance-specific
physicochemical parameters, and species-specific physiological and biological parameters.
The numerical estimates of these model parameters are incorporated within a set of
differential and algebraic equations that describe the pharmacokinetic processes. Solving
these differential and algebraic equations provides the predictions of tissue dose. Computers
then provide process simulations based on these solutions.
The structure and mathematical expressions used in PBPK models significantly simplify the
true complexities of biological systems. If the uptake and disposition of the chemical
substance(s) are adequately described, however, this simplification is desirable because data
are often unavailable for many biological processes. A simplified scheme reduces the
magnitude of cumulative uncertainty. The adequacy of the model is, therefore, of great
importance, and model validation is essential to the use of PBPK models in risk assessment.
Importantly, the accuracy and utility of these models are dependent on the validity of assumptions
that are made and the quality of the data providing the values for key parameters used in the
equations. In considering the utility of one of these models the ATSDR makes the observation that;
”The human model was calibrated to predict limitation half-times estimated for human
populations (e.g. 2.3 or 3.8 years for PFOA, 5.4 years for PFOS). As a result, comparisons
made between observed and predicted serum concentrations evaluate whether or not the
populations actually exhibit the half-times to which the model was calibrated, and not the
validity of the model to predict the internal distribution of PFOA or PFOS. It is not currently
possible to assess with confidence whether the human model can accurately predict doses to
liver or any other tissues”
The draft record of outcomes from the enHealth workshop also notes several difficulties in adopting
reference values based on the US EPA PBPK modelling due to weaknesses or lack of transparency in
those models.
“There are some issues with the level of understanding of the toxicokinetics of PFOS and
PFOA within and between species that make the approaches adopted by USEPA and ATSDR
somewhat problematic.
Volume of distribution – the volume of distribution is the apparent volume of the body
within which a chemical distributes once it enters the body. Reaching steady state
concentrations (those where intake and elimination are balanced) requires a large
proportion of the storage locations to be filled.
For PFCs it is known they are highly bound to serum albumin, they are therefore confined
primarily to extracellular fluid and have limited distribution into other tissues. However the
distribution mechanisms are not fully understood, may be different in different organisms,
and the volume of distribution is difficult to determine. Hence the reliability of values used
for the volume of distribution in extrapolating no effect serum concentrations in animals to
an equivalent human dose is uncertain.
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The workshop attendees agreed the variability in the toxicokinetics for these chemicals
mean that it would be appropriate to use toxicokinetic modelling in determining a tolerable
daily intake if there was sufficient evidence that the parameter values used were robust and
reliable. However the workshop attendees agreed there was not sufficient understanding
of the parameterisation of the modelling approach taken by USEPA or ATSDR for it to be
adopted prior to the assessment to be undertaken by FSANZ.
At plasma steady state, the ratio of tissue levels of PFAS to that in a given tissue for a specific species
will remain relatively constant regardless of the species under consideration or the differences in
other pharmacokinetic parameters between species. The comparative values for those relationships
however will differ markedly between species if other pharmacokinetic parameters vary between
the species being compared. Critically, where the HED approach is used to determine the PoD for
liver effects seen in rats this approach is likely to greatly overestimate the dose to the liver in
humans. Where dramatic differences in t½ (or more precisely Clearance and Volume of distribution)
exist as for PFAS in humans compared to rats, the ratio between serum levels and liver levels,
particularly during the absorption phase, will also be very different. As the liver receives 75% of its
blood flow from the hepatic portal vein which drains the intestinal mesentery and only 25% from the
systemic blood through the hepatic artery, the dominant determinant of the pattern and extent of
liver exposure is the concentration in the hepatic venous blood which is proportional to the
intestinal concentration during the absorption phase (i.e. the oral dose and the time course of
absorption) rather than the systemic blood concentration. This is particularly true where rodents
are dosed in the food and allowed to eat ad libitum. Because clearance in humans is markedly
slower, resulting in a half-life of years compared to a few days in rats, the serum concentration in
humans for a given oral dose will be higher than in rats but the liver concentration during the
absorption phase is likely to be more similar to that in rats for the same dose. This is true even given
the observation that post mortem liver to serum ratios in rats are similar to post mortem ratios in
humans as the ratio post mortem reflects equilibration subsequent to the absorption phase. Rats
are generally fasted prior to termination and equilibration between liver and serum may occur
subsequent to the absorption phase.
Consequently, the use of HED for liver effects in experimental animals with markedly shorter t½ than
humans, particularly rats, may greatly overestimate the potential liver exposure, and therefore may
not necessarily decrease uncertainty in dose extrapolation, and is arguably inappropriate at least for
liver effects.
The PBPK approach does however have potential value for modelling the HED for other toxicological
effects. The ratio between maternal serum levels of PFAS and that of breast milk and placental cord
blood are similar between humans and rats. A cross species dose comparison based on serum PFAS
rather than oral dose administered to rats, does therefore provide a potentially more robust basis
for identifying the PoD provided assumptions used in the PBPK modelling are robust and grounded in
adequate data for each species and humans modelled.
In this last respect a recent review by Dong et al (2016), of the PBPK approach of the US EPA is
pertinent and reflects the concerns of the enHealth workshop (attended by at least two of the
authors of the paper). These authors, while acknowledging the potential benefit of PBPK modelling
and HEDs, have concluded that the approach of the US EPA to toxicokinetic modelling may be
compromised by “systematic fitting residual errors across species”, that the current approaches may
lead to unnecessarily conservative reference doses and that the basis of modelling used by the US
EPA may benefit from further refinement. A reconciliation of the various viewpoints on the validity
23 | P a g e
and utility of PBPK modelling in the specific instance of PFAS risk assessment is beyond the scope of
this review but will need to be a key consideration in the FSANZ review of HRVs for PFAS.
5.3 MECHANISMS OF ACTION
Where the data will allow, a careful consideration of the mechanisms of pharmacokinetic variability
and toxicological action are essential components of any risk assessment. Interspecies variations in
responses to a toxicant are determined by one or both of these. How and to what extent these
mechanisms are incorporated into a risk assessment and the extent and direction to which they
moderate the degree of uncertainty and the uncertainty factors applied, is a common source of
variation between risk assessments of the same toxicant.
For PFAS there are two distinct sources of interspecies variation in sensitivity. Firstly, there is a clear
and, very unusually, dramatic difference in pharmacokinetics between humans and all experimental
animals so far examined, with non-human primates having the most similar kinetics. This variance
acts to increase the blood levels of PFAS in humans for any given dose (in mg/kg bw/d) compared to
experimental animals, because although humans and rats appear to absorb PFAS to a similar extent,
humans excrete the compounds extremely slowly by comparison (Clearance in humans 0.03 mL/kg
bw/day, female rats 666 mL/kg bw/day). Regardless of the risk assessment process applied, the
derivation of HRVs must incorporate adjustments to accommodate this known difference.
5.3.1 Pharmacokinetics
Ingested PFOA and PFOS are essentially fully absorbed over approximately an hour. The mechanism
of absorption has not been fully elucidated but, as PFAS do not have the characteristics necessary for
ready passive absorption, is likely to involve elements of active transport by organic anion
transporters (OAT), and/or absorption in conjunction with lipids.
PFAS bind to serum albumin and various other plasma proteins including gamma-globulin, alphaglobulin, alpha-2-macroglobulin, transferrin, and beta-lipoproteins in both rat and human plasma
but the affinity in rats for albumin binding is an order of magnitude greater than for humans. The
albumin -PFOA dissociation constant is 0.4 mM for human serum albumin and 0.36 nM for rat serum
albumin and involves 6–9 binding sites with noncovalent binding apparently at the same sites as
fatty acids.
Absorbed PFAS distribute widely from plasma into the soft tissues. The highest concentrations are
found in the liver. As for absorption from the GIT the mechanism of absorption into the liver is not
fully understood but, again, is likely to involve active transport by OAT involved in movement of fatty
acids and other organic anions. There is evidence that PFOA is a ligand for OAT in the luminal and
basolateral membranes of renal tubular epithelial cells, which transport PFOA in the glomerular
filtrate back into the tubular cells. The difference in plasma half-lives of the PFAS between species is
therefore likely to be attributable to differences in the nature, density and activity of renal tubular
OAT between species and differences in their affinity for PFAS.
5.3.2
Toxicity
The most sensitive toxic effects of PFAS in animals are hepatic and developmental effects which are
both mediated to some extent by activation of the peroxisome proliferator-activated receptor-α
(PPARα). PPARα is a member of the nuclear receptor superfamily that mediates a broad range of
biological responses including lipid metabolism, energy homeostasis, and cell differentiation.
Marked differences in the sensitivity of species to PPARα activating toxicants have been observed.
24 | P a g e
Humans, non-human primates and guinea pigs substantially resistant to PPARα mediated toxicity
and rodents (rats and mice) the most sensitive.
Differences in response to PPARα agonists across species are likely to be due to a combination of;
differences in the ability of PPARα to be activated by peroxisome proliferators, differences in the
inducibility (increase in production) of PPARα after exposure to peroxisome proliferators and
differences in pattern and levels of tissue-specific expression of PPARα. Notably the level of
expression of PPARα in human liver is about 1–10% of that in the rat and mouse liver.
Although the critical biological target pathways leading to developmental effects in rodents have not
been established, the observation in a number of studies that developmental toxicity of PFOA/PFOS
and other ligands for PPARα, is significantly dependent on the expression of PPARα, indicates that
rodents are more likely to over-predict human developmental risk for such substances than underpredict. A study of PFOA in pregnant mice for example, used wild-type, PPARα-null, and PPARαhumanized (expressing human PPARα) mice and demonstrated lower postnatal survival in wild-type
mice, as predicted by rat studies, but no effect in null or humanized mice. Results indicate that
PPARα mediates at least some of the developmental effects in mice, and that species differences
exist between mice and humans.
The normal role of PPARα is the regulation of lipid homeostasis through the modulation of
expression of genes involved in fatty acid uptake, activation, and oxidation. PPARα receptor
activation by toxicants in rats and mice initiates a sequence of biochemical events that include
marked hepatocellular hypertrophy due to an increase in number and size of peroxisomes, a large
increase in peroxisomal fatty acid β-oxidation, CYP450 induction and alterations in lipid metabolism.
Both PFOA and PFOS alter the expression in rats of genes associated with lipid homeostasis and
down- regulate genes that control cholesterol biosynthesis. In comparison with naturally occurring
long-chain fatty acids such as linoleic and α-linoleic acids, PFOA and PFOS are relatively weak ligands
for PPARα.
5.3.3 Comment
The marked differences between the pharmacokinetics of PFAS in experimental animals and humans
has been inadequately addressed in the EFSA assessment for effects other than liver toxicity (where
a MOE approach was used for PFOS based on a monkey study). EFSA have used standardised
uncertainty factors to account for the differences in pharmacokinetics whereas the US EPA has taken
a different approach to managing this uncertainty, using PBPK modelling incorporating many
assumptions. The use of HED or high uncertainty factors for liver toxicity may not be appropriate as
it may greatly over-predict risk. For developmental toxicity use of average serum levels as the dose
comparator across species may be the most appropriate approach as the ratio between maternal
serum levels and foetal/neonate exposure is sufficiently similar in rats and humans. The
extrapolation may need to be tempered by the likely higher sensitivity of rodents to developmental
effects due to differences in PPARα between species.
The available evidence indicates PPARα plays a key role in the toxicology of PFAS and that rodents in
particular are likely to be considerably more sensitive than humans. This area requires further
consideration in terms of the appropriate uncertainty factor to be applied for interspecies
extrapolation.
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5.4 EPIDEMIOLOGY
Epidemiology studies have been conducted on populations exposed occupationally, those exposed
to, relatively, high levels in contaminated drinking water through leakage from nearby
manufacturing facilities, and the general population. Blood levels in occupationally exposed
populations are 2 to 3 orders of magnitude (100 – 1000 times) higher than the general population
and provide a worst case exposure. As the occupationally exposed population is predominantly
male it can only provide limited evidence regarding potential female reproductive toxicity. The
following discussion compares the conclusions of EFSA and the US EPA/ATSDR reviews regarding the
findings from these studies to highlight consistencies and differences in their conclusions. A detailed
consideration of individual studies is beyond the scope of, and time available for this review.
Epidemiology studies available for PFAS exposed populations have found a range of associations
between exposure and various diseases outcomes including cancer. When analysed as outlined
earlier, these studies have, except as noted, generally failed to provide substantive evidence of
adverse outcomes in exposed human populations, including those occupationally exposed where
serum levels are 2-3 orders of magnitude greater than that likely in exposed populations in Australia.
The conclusions of EFSA, the US EPA and the ATSDR reflect the considerations discussed above and
for the most part are largely concordant, with greater or lesser degrees of caution and nuancing of
those conclusions. There is a trend for the US EPA to conclude evidence of association where both
the ATSDR and EFSA (together with most other international regulatory assessments) have
concluded the deficiencies in or inconsistencies between studies preclude such conclusions. These
differences are likely to reflect the input of the various review participants rather than a
fundamental difference in interpretation by the expert epidemiology evaluators.
A reconciliation of these differences is beyond the scope of this review and should be included in the
FSANZ HHRA.
5.4.1 Exposure
Serum levels of PFOA and PFOS in occupationally exposed subjects are of the order of 1 to 2 µg/mL
for each. In the highly exposed population around one US facility serum levels for PFOA averaged
0.423 µg/mL in 2004-2005. For the US population as a whole serum levels were 0.00392 µg/mL in
2004-2005.
5.4.2 Carcinogenicity
Studies in occupationally exposed populations who have the highest blood levels of PFAS and highly
exposed populations around PFAS manufacturing facilities have not, collectively, provided a basis for
concluding that PFAS cause cancer. The overall conclusions of the ATSDR, EFSA and US EPA are
largely concordant, with the exception of the US EPA conclusion regarding PFOA.
The ATSDR summarised the evidence as follows:
Although several studies have found significant increases in cancer risk, the results should be
interpreted cautiously since most studies did not control for potential confounding variables
(particularly smoking), the number of cancer cases was low, and a causal relationship between
perfluoroalkyls and cancer cannot be established from these studies. Additionally, the lack of
consistency across facilities may be suggestive of a causative agent other than PFOA or PFOS.
EFSA conclusions on epidemiology are:
Epidemiological studies in PFOA-exposed workers do not indicate an increased cancer risk.
26 | P a g e
Epidemiological studies in PFOS exposed workers have not shown convincing evidence of
increased cancer risk.
US EPA conclusions on Epidemiology are:
PFOS
Human epidemiology studies did not find a direct correlation between PFOS exposure and the
incidence of carcinogenicity in worker-based populations. Although one worker cohort found
an increase in bladder cancer, smoking was a major confounding factor, and the standardized
incidence ratios were not significantly different from the general population. Other worker and
general population studies found no statistically-significant trends for any cancer type.
PFOA
Under EPA’s Guidelines for Carcinogen Risk Assessment (USEPA 2005a), there is
“suggestive evidence of carcinogenic potential” for PFOA. Epidemiology studies
demonstrate an association of serum PFOA with kidney and testicular tumors among highly
exposed members of the general population.
Although the overall conclusion of the US EPA regarding the association between PFOA exposure
and cancer varies somewhat from the other two assessments, in their discussion of the overall
conclusions from the human cancer epidemiology studies they note that:
A group of independent toxicologists and epidemiologists critically reviewed the
epidemiological evidence for cancer based on 18 studies of occupational exposure to
PFOA and general population exposure with or without co-exposure to PFOS. The project
was funded by 3M, but the company was not involved in the preparation or approval of
the report. The authors evaluated the published studies based on the study design,
subjects, exposure assessment, outcome assessment, control for confounding, and
sources of bias. They followed the Bradford Hill guidelines on the strength of the
association, consistency, plausibility, and biological gradient in reaching their conclusion.
They found a lack of concordance between community exposures and occupational
exposures one or two magnitudes higher than those for the general population. The
discrepant findings across the study populations were described as likely due to chance,
confounding, and/or bias (Chang et al. 2014).
The discordance within the US EPA assessment likely reflects the nature of the agencies
guidance rather than an issue of scientific interpretation. Notably the available epidemiology
studies do not appear to have had a bearing on the selection of a PoD for human health risk
assessment by any of the agencies whose reviews were considered by enHealth.
A more detailed review of the epidemiological assessments regarding carcinogenesis than is
possible in this review is an appropriate inclusion in the FSANZ consideration of HRVs for PFAS for
Australia.
5.4.3 Reproductive effects
A number of studies have found an association between PFAS exposure and lower, but not low, birth
weight. “Low birth weight” is a clinically defined condition associated with adverse outcomes in
neonates. A lower average birth weight, that remains within the clinically normal range, does not
constitute “low birth weight” in the clinical context.
27 | P a g e
This observation however does not equal a conclusion that lower average birth weights in exposed
populations are acceptable or appropriate, a consideration requiring further attention in the FSANZ
review preferably with clinical input.
ATSDR
There is evidence to suggest that high serum PFOA or PFOS levels are associated with lower
birth weights. The significant associations have come from general population studies and a
study of highly exposed residents. Studies of populations with lower serum PFOA or PFOS
levels have not found significant associations for birth weight. Although significant
associations were found, decreases in birth weight were small and may not be biologically
relevant. No studies found an increased risk of low birth weight in infants (<2,500 g) in
highly exposed residents.
EFSA
PFOA
In two recent studies, PFOA exposure of pregnant women, measured by maternal and/or cord
serum levels was associated with reduced birth weight. The Panel noted that these
observations could be due to chance, or to factors other than PFOA.
PFOS
The very few epidemiological data available for the general population do not indicate a risk
of reduced birth weight or gestational age.
US EPA
PFOA
Studies in the high-exposure community reported an association between serum PFOA
and risk of pregnancy-related hypertension or preeclampsia, conditions related to renal
function during pregnancy; this outcome has not been examined in other populations. An
inverse association between maternal PFOA (measured during the second or third
trimester) or cord blood PFOA concentrations and birth weight was seen in several
studies.
The epidemiology studies did not find associations between PFOA and
neurodevelopmental effects, or preterm birth and other complications of pregnancy.
PFOS
….the available information indicates that the association between PFOS exposure and
birth weight for the general population cannot be ruled out.
Although there remains some concern over the possibility of reverse causation explaining
some previous study results, these collective findings indicate a consistent association
with fertility and fecundity measures and PFOS exposures.
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5.4.4
Other Effects
ATSDR
… consistent findings were found for association of serum PFOA and PFOS with
increases in serum lipid levels, increases in uric acid levels, and alterations in
biomarkers of liver damage. Although other effects have been reported, they have
not been consistently found in similar types of studies, have only been examined in a
single study, or were only found in general population studies.
EFSA
PFOA
Epidemiological studies in PFOA-exposed workers do not indicate an increased
cancer risk. Some have shown associations with elevated cholesterol and
triglycerides, or with changes in thyroid hormones, but overall there is no consistent
pattern of changes.
PFOS
Epidemiological studies in PFOS exposed workers have not shown convincing
evidence of increased cancer risk. An increase in serum T3 and triglyceride levels was
observed, which is the opposite direction to the findings in rodents and monkeys.
US EPA
PFOA
….epidemiology studies have generally found positive associations between serum PFOA
concentration and total cholesterol (TC) in the PFOA-exposed workers and the highexposure community (i.e. increasing lipid level with increasing PFOA); similar patterns
are seen with low-density lipoproteins (LDLs) but not with high-density lipoproteins
(HDLs). These associations were seen in most of the general population studies, but
similar results also were seen with PFOS, and the studies did not always adjust for these
correlations. Associations between serum PFOA concentrations and elevations in serum
levels of alanine aminotransferase (ALT) and gamma-glutamyl transpeptidase (GGT)
were consistently observed in occupational cohorts, the high-exposure community, and
the U.S. general population. The associations are not large in magnitude, but indicate
the potential for PFOA to affect liver function. Diagnosed thyroid disease in females and
female children was increased both in the high-exposure C8 study population and in
females with background exposure; thyroid hormones are not consistently associated
with PFOA concentration. Associations between PFOA exposure and risk of infectious
diseases (as a marker of immune suppression) were not identified, but a decreased
response to vaccines in relation to PFOA exposure was reported in studies in adults in
the high-exposure community population and in studies of children in the general
population; in the latter studies, it is difficult to distinguish associations with PFOA from
those of other correlated PFASs.
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PFOS
The strongest associations are related to serum lipids with increased total cholesterol
and high density lipoproteins (HDLs). The associations for most epidemiology endpoints
are mixed. While mean serum values are presented in the human studies, actual
estimates of PFOS exposure (i.e., doses/duration) are not currently available. Thus, the
serum level at which the effects were first manifest and whether the serum had
achieved steady state at the point the effect occurred cannot be determined. It is likely
that some of the human exposures that contribute to serum PFOS values come from
PFOS derivatives or precursors that break down metabolically to PFOS. These
compounds may originate from PFOS in diet and materials used in the home, thus, there
is potential for confounding. Additionally, most of the subjects of the epidemiology
studies have many perfluoroalkyl substances (PFAS), other contaminants, or both in
their blood. Taken together, the weight of evidence for human studies supports the
conclusion that PFOS exposure is a human health hazard.
6 CONCLUSIONS
The primary focus of this review has been to identify the principle sources of variation between the
US EPA and EFSA risk assessments of PFAS and the resultant guidance values, and to form a view on
the suitability of the EFSA values selected by enHealth as an interim measure pending more
extensive consideration by FSANZ. In the time available for this review it is not possible to
definitively state that either one of the reviews is “correct” and the other not. There are potential
strengths and potential weaknesses in each assessment and both contain value judgements that are
as much policy or convention based as they are science based. Deviations by each agency from what
are the general routine approaches to risk assessment are not inappropriate simply because they are
not routine. A judgement to that effect requires a more detailed analysis of the reasoning and of the
science underlying the respective approaches. Equally, while there is considerable room for
improvement in the enHealth process and particularly the level of documentation supporting the
decision making process, that observation does not impact the validity of the decisions themselves.
The draft enHealth workshop report, and background documents utilised by the workshop, indicate
that the workshop gave careful consideration to the sources of the variations between EFSA and US
EPA reference values and considered the strengths and weaknesses of the approach taken by each
agency. In particular, the workshop discussed the use of PBPK modelling by the US EPA in some
detail and noted significant concerns around the nature and range of assumptions required to
support the model.
Some weaknesses in the enHealth process have been identified, most particularly around the lack of
scientific support provided to expert working groups and the enHealth committee, and a consequent
lack of adequate detail and transparency in records of the decision making process and rationale.
The draft record of the deliberations of the expert workshop is notably brief given the extent of the
work conducted by the experts attending, and the extensive discussion of the various points raised.
That the draft workshop report was not finalised, and had not been circulated to the attendees for
comment, as at the completion of this review is notable.
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6.1 USE OF INTERNATIONAL RISK ASSESSMENTS AND DERIVED HRVS
The use by Australian regulatory agencies of human health reference values, and the associated risk
assessments, derived by international agencies such as the US EPA or EFSA, as interim measures to
support immediate actions or deliberations is not unusual and is an appropriate mechanism for
responding to environmental contamination issues in a timely manner.
6.2 SOURCES OF DIFFERENCES BETWEEN US EPA AND EFSA RISK ASSESSMENTS
As with all toxicological risk assessments, there are various sources of uncertainties, strengths and
weaknesses in both the EFSA and US EPA derivations of their respective HRVs.
Although the US EPA assessment is more recent, the differences between the two assessments and
the resultant HRVs are not due to differences in the studies available for assessment by the two
agencies. The key sources of variation relate to the use of PBPK modelling by the US EPA and ATSDR
uniquely, and differences in selection of uncertainty factors by EFSA and the US EPA. PBPK
modelling is not a routine aspect of risk assessment methodologies for major international agencies
other than the US EPA. That PBPK modelling is not a normal aspect of risk assessment in Australia or
internationally is not however, of itself, a basis for rejecting the approach, particularly where
pharmacokinetics is a pivotal source of interspecies differences in response to PFAS. The core
consideration, as identified by the enHealth workshop, is the validity and utility of the model. Thus
although the US EPA assessment utilised ostensibly sophisticated PBPK modelling the necessary
range of assumptions, between humans and the respective experimental animals, means that the
use of PBPK has potentially replaced one set of uncertainties with another.
The US EPA use of HEDs based on PBPK modelling of serum levels of PFAS is not likely to be
appropriate for liver effects because liver exposure for the same serum PFAS levels will be higher in
rats than in humans, at the least during the absorption phase. Actual administered dose in mg/kg bw
is likely to be a better basis for determining the PoD for liver effects.
The use of average or actual final serum levels, by both the US EPA and EFSA, to compare internal
exposures/doses across experimental animals to determine the PoD is appropriate for compounds
with high variation in pharmacokinetic parameters. The limitations of the model used to convert
final serum level to average levels by the US EPA PBPK modelling, which is not a routine aspect of
risk assessment methodologies for major international agencies other than the US EPA (and ATSDR),
is noted however.
There are indications in the data set that, at least for a number of toxicological effects used as the
basis for determination of the PoD, humans are likely to be less sensitive than animals due to both
pharmacokinetic (liver) and pharmacodynamics (developmental effects) differences. Further
consideration of the appropriate adjustment (reduction) of uncertainty factors to recognise this
observation are warranted.
6.3 POTENTIAL PUBLIC HEALTH CONSEQUENCES OF THE CHOICE OF HRVS
Because of the exceptionally long half-life of PFAS in humans, the systemic (i.e. internal) exposure to
PFAS is determined by oral (or other routes of) exposure over long periods of time. As a
consequence, even an order of magnitude reduction in levels in drinking water would not
significantly impact blood levels for a protracted period and the issue of whether the EFSA or US EPA
values are more appropriate is largely esoteric over the short to medium term.
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As the phasing out of PFAS for most uses has resulted in declining exposures and a progressive
decline in serum levels where this has been monitored, the choice of EFSA HRVs over those of the US
EPA has no substantive impact on public health in the short to medium term.
6.4 BALANCING RISK MITIGATION WITH RISK GENERATION
For anthropogenic contaminants such as PFAS which cannot readily be removed from the
environment, the establishment of values that are, with respect to the overall weight of evidence,
disproportionately low, has the potential to result in a range of adverse health outcomes which may
be greater than the toxicological risks intended to be avoided. Such outcomes may include
prolonged unwarranted stress in exposed populations, the recommendation, or seeking out, of
unnecessary medical interventions with their attendant risks, or interventions in pregnancy and
avoidance of breast feeding to the detriment of the foetus and neonate. Other, economic impacts
although likely, are beyond the scope of a HHRA and of this review. Simplistic selection of the
lowest international HRV is therefore not necessarily optimal for the overall protection of public
health, regardless of the time sequence of the available assessments. Determination of suitable
HRVs for PFAS requires a careful consideration of the strengths and weaknesses of the approaches
taken by international agencies that have had access to the underlying data, and a considered
selection of the most appropriate approach/values within the context of the exposure patterns in
Australia. A suitably precautionary approach to public health requires a balancing of risk prevention
against the potential for risk generation.
6.5 OVERALL CONCLUSION
The adoption of the EFSA HRVs as a temporary (i.e. interim) measure, pending a formal more
extensive review by FSANZ, is appropriate and is protective of public health.
7 RECOMMENDATIONS
Based on the findings of this review the following recommendations are made;

The adoption of the EFSA health reference values (TDI) and their use to derive Australian
drinking water guideline values, as an interim measure pending FSANZ review, can be
concluded to be appropriate, to be based on the expert consideration of the strengths and
weaknesses of the available risk assessments from international agencies, and to be consistent
with current risk assessment practices both in Australia and internationally.

Consideration should be given to the need for the responsibility for setting HRVs, and particularly
for contaminants that are also present in food and water, to be formally supported in future by an
existing agency with;
 experience and expertise in setting, and documenting, these values,
 appropriate consultation mechanisms in place to ensure the highest possible degree of
transparency,
 the capacity to provide expert scientific support to expert working groups and decision making
committees,

The scientific literature on, and international regulatory assessments of the HRVs for PFAS
should be monitored on an ongoing basis by FSANZ in conjunction with enHealth and adjusted
up or down as indicated by the emerging data.
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
Australia, through FSANZ or another suitable agency, should consider whether there is value in
seeking to initiate an international consultative review of HRVs for PFAS through the
CODEX/JECFA mechanism to establish consistent international HRVs for these substances.

As identified in both this review and by the enHealth workshop, pivotal issues that FSANZ should
address and consider seeking specialist expert advice on, include;




the strengths, weaknesses and validity of the PBPK approach to HED calculation for PFAS,
the relative merits of the interpretation of the epidemiology data by the US EPA compared
to most other international agencies’
the clinical relevance of the observed lower birth weights and elevated cholesterol levels in
highly exposed populations,
the significance of the recent US National Toxicology Program (NTP) review of the immunetoxicity potential of PFAS.
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