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Cooperative Research Centre for Contamination Assessment and Remediation of the Environment

Assessment, management and remediation for PFOS and PFOAPart 2: health screening levels

TechnicAl RePORT nO. 38

Cooperative Research Centre for Contamination Assessment and Remediation of the Environment, Technical Report series, no. 38 January 2017 Copyright © CRC CARE Pty Ltd, 2017 This book is copyright. Except as permitted under the Australian Copyright Act 1968 (Commonwealth) and subsequent amendments, no part of this publication may be reproduced, stored or transmitted in any form or by any means, electronic or otherwise, without the specific written permission of the copyright owner. ISBN: 978-1-921431-54-8

Enquiries and additional copies: CRC CARE, C/- Newcastle University LPO, PO Box 18, Callaghan NSW, Australia 2308 Tel: +61 (0) 2 4985 4941 Fax: +61 (0) 8 8302 3124 [email protected] www.crccare.com

This report should be cited as: CRC CARE 2017, Assessment, management and remediation guidance for perfluorooctanesulfonate (PFOS) and perfluorooctanoic acid (PFOA) – Part 2: health screening levels, CRC CARE Technical Report no. 38, CRC for Contamination Assessment and Remediation of the Environment, Newcastle, Australia.

Disclaimer: This publication is provided for the purpose of disseminating information relating to scientific and technical matters. Participating organisations of CRC CARE do not accept liability for any loss and/or damage, including financial loss, resulting from the reliance upon any information, advice or recommendations contained in this publication. The contents of this publication should not necessarily be taken to represent the views of the participating organisations.

CRC for Contamination Assessment and Remediation of the Environment

Technical Report no. 38b

Assessment, management and remediation guidance for perfluorooctanesulfonate (PFOS) and

perfluorooctanoic acid (PFOA)

Part 2 – health screening levels

March 2017

CRC CARE Technical Report no. 38b i Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

Important note to readers

This guidance is to be regarded as both draft and interim, particularly in relation to health based screening values.

The following caveats and limitations apply to this guidance:

• the health based values in this guidance are based on the interim TDI values endorsed by enHealth in 2016 for PFOS (+PFHxS) and PFOA, and

• should revised TDI values be endorsed by enHealth or another major national health based agency in Australia, then the health derived values in this guidance will be updated accordingly, and the guidance re-issued.

It is recognised that there is much on-going research on the impacts of PFOS and PFOA on human health and the environment, the outcomes of which may improve the assessment, remediation and management of PFOS and PFOA. These developments will be monitored and this guidance will be updated, as appropriate.

It is also recognised that PFAS compounds other than PFOS and PFOA may contribute to the impacts of PFAS on human health and the environment. When further research results for other PFAS compounds (or classes thereof) of sufficient robustness to be utilised in the formulation of guidance for those compounds (or classes thereof) become available, then that information will be used to extend the scope of this guidance.

This guidance should therefore be regarded as both draft and interim.

March 2017

CRC CARE Technical Report no. 38b ii Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

Preamble

CRC CARE has milestones in its Agreement with the Commonwealth Government relating to the development of guidance for contaminants of emerging concern in collaboration with end users. Priority contaminants were identified for CRC CARE through regulators and end users who met at a Forum held in February 2012. These contaminants included perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA), which belong to a large group of compounds called per- and polyfluoroalkyl substances (PFAS). CRC CARE has undertaken the development of human health screening levels (HSLs) and ecological screening levels (ESLs) for PFOS and PFOA, These values provide a collective view of the available science and application of Australian approaches on the development of human health and ecological based screening levels.

The derivation of the HSLs and ESLs has followed the methodologies outlined in the National Environment Protection (Assessment of Site Contamination) Measure 1999, as revised in 2013 (NEPM). HSLs and ESLs are subject to the assumptions, uncertainties and limitations outlined in the relevant parts of the guidance – deviations from the assumptions used in their derivation may mean that a site specific assessment would be required. Further, HSLs for some exposures are calculated using a 100% relative source contribution. If more than one of these sources is relevant, use of the screening values would underestimate risks, and in this case, cumulative assessment of all relevant exposures must be made.

The NEPM emphasises the importance of formulating a conceptual site model on which to base the assessment of soil and groundwater contamination, and guidance has been provided on the development of a conceptual site model for PFOS and PFOA.

It is intended that the HSLs and ESLs that have been developed for PFOS and PFOA should be considered similarly to the NEPM HILs/HSLs and EILs/ESLs in forming generic screening levels which, if exceeded, would indicate that further more detailed investigation is required. The HSLs and ESLs and the information used in their derivation can assist in undertaking this more detailed investigation.

It is emphasised that exceedance of the HSLs and ESLs does not necessarily imply that the contamination poses an unacceptable risk, and the HSLs and ESLs should not be used as remediation targets, as this could result in unnecessary remediation.

The following limitations should be noted:

• The assessment and management of PFAS contamination in soil and groundwater is an emerging science in Australia and internationally and significant work continues to be undertaken. It is recommended that when using this guideline, more recent information is considered.

• Uncertainties and limitations in the guidance values and their application are noted in the guidance document, and these should be considered when using the values. Some of the screening levels have assumed the occurrence of bioaccumulation and biomagnification based on international work - there is uncertainty as to the levels that will occur in Australian organisms. Where significant exceedances of HSLs and ESLs occur, consideration should be given to direct monitoring of

CRC CARE Technical Report no. 38b iii Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

potentially affected organisms to determine if there are effects that can be distinguished. Further assessment would be would be required where bioaccumulation is an issue.

• This review has focussed on PFOS and PFOA. PFAS include a large number of compounds and analytical methods are reporting an increasing number of PFAS compounds. As this guidance document was being finalised, enHealth (2016) has advised that perfluorohexane sulfonate (PFHxS) should be considered to be additive to PFOS. The use of the HSLs and ESLs for PFOS and PFOA should recognise that other PFAS may contribute to potential effects and where present, these will need to be evaluated for potential cumulative risks in accordance with available information.

Because of the evolving nature of the science relating to PFAS, these guidelines are considered to be Interim Guidance, and it is recommended that a review of this guidance is considered following any change in recommendations relating to guideline levels for PFAS by enHealth or FSANZ.

This guidance is intended for a variety of users within the contaminated sites industry, including site owners, proponents of works, contaminated land professionals, and regulators. It is assumed that readers are familiar with the NEPM 2013. While the aim of this guideline is to provide a resource that can be used at PFAS-contaminated sites across Australia, it does not replace specific laws, regulations and guidance provided at a local level.

CRC CARE Technical Report no. 38b iv Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

Abbreviations

AFFF Aqueous film forming foam

BAF

BCF

Bioaccumulation factor

Bioconcentration factor

BMF Biomagnification factor

BSAF

CDI

Biota-sediment accumulation factor

Chronic daily intake

CSM Conceptual site model

CRC CARE Cooperative Research Centre for Contamination Assessment and Remediation of the Environment

ESL Ecological screening levels

HED Human equivalent dose

HIL Health investigation level

HRV Health reference value

HSL Health screening levels

MPC Maximum permissible concentration

NEPM National Environment Protection (Assessment of Site Contamination) Measure 1999, as revised in 2013

NRF National Remediation Framework

PFAS Per- and polyfluoroalkyl substances

PFHxS Perfluorohexane sulfonate

PFOA Perfluorooctanoic acid

PFOS Perfluorooctane sulfonate

RfD Reference dose

TDI Tolerable daily intake

TRV Toxicity reference value

UF Uncertainty factor

CRC CARE Technical Report no. 38b v Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

Table of contents

Executive summary i

1. Introduction 1

1.1 Project background 1

1.2 Purpose of this document 1

1.3 Basis for development of human health screening levels for PFAS and selection of parameters 2

1.4 Conceptual site model 5

2. PFOS and PFOA toxicity assessment 6

2.1 Overview 6

2.2 Summary of health effects 6

2.2.1 Acute toxicity 6

2.2.2 Chronic exposure 6

2.3 Toxicokinetics 7

2.4 Background levels 8

2.5 Summary of toxicity reference values 8

2.6 Interim guidance published by enHealth 10

2.7 Procedural review of the health reference values established by enHealth 11

3. Risk characterisation 13

3.1 Overview 13

3.2 Threshold toxicology 13

4. Exposure assessment and development of HSLs 15

4.1 Overview 15

4.2 Fate and transport 15

4.2.1 Chemical stability 15

4.2.2 Groundwater migration potential and characteristics 15

4.2.3 Biomagnification in food chain 15

4.3 Soil receptor populations 16

4.3.1 Residential land (HSL A and HSL B) 17

4.3.2 Public open space (HSL C) 17

4.3.3 Commercial/industrial (HSL D) 17

4.3.4 Consumers of home grown produce and animal products 18

4.3.5 Surface water receptors – consumption of seafood 18

4.4 Groundwater receptor populations 18

CRC CARE Technical Report no. 38b vi Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

4.4.1 Drinking water 18

4.4.2 Primary contact recreation 18

4.4.3 Irrigation/agriculture 19

4.4.4 Stock water 19

4.4.5 Consumption of seafood 20

4.4.6 Aquaculture 20

4.5 Exposure pathway assessment 20

4.5.1 Oral exposure 20

4.5.2 Dermal absorption 21

4.5.3 Inhalation of dust 21

4.5.4 Intakes from other sources 22

4.6 Exposure estimation – soil 22

4.6.1 Exposure parameters 22

4.6.2 Summary of derived soil HSLs 23

4.7 Exposure estimation – water (fish consumption) 26

4.7.1 Fish consumption rates 26

4.7.2 Maximum permissible concentration in fish 27

4.7.3 RIVM approach to assessing bioaccumulation 28

4.7.4 Application of the RIVM approach to assessing bioaccumulation in Australia 29

4.7.5 Water and sediment HSLs for seafood consumption 30

4.7.6 Summary of derived surface water HSLs for human consumption of seafood 31

4.8 Limitations, uncertainty and application of HSLs 32

4.8.1 Conceptual site model 32

4.8.2 Soil HSLs for PFOS and PFOA 33

4.8.3 Surface water HSLs for PFOS and PFOA for consumption of seafood 34

4.8.4 Interim drinking water quality and recreational water quality guidelines 35

4.8.5 Multiple pathway exposure and cumulative risk 35

5. References 37

Appendices

Appendix A. Human health toxicity review 44

Appendix B. HSL calculations 84

CRC CARE Technical Report no. 38b vii Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

Tables

Table 1. Comparison of TDIs adopted internationally 9

Table 2. enHealth interim toxicity reference values 11

Table 3. Exposure parameters – soil exposure 25

Table 4. Summary of soil HSLs for PFOS (+PFHxS) and PFOA (mg/kg) 25

Table 5. Example international soil (residential) criteria for PFOS and PFOA (mg/kg)

26

Table 6. Average seafood consumption rates (g/day) 27

Table 7. US estimates for recreational marine fish intake (g/day) 27

Table 8. Summary of derived water and sediment HSLs protection of human health consuming seafood

31

Figures

Figure 1. Health risk assessment process 3

Figure 2. Risk assessment methodology 4

CRC CARE Technical Report no. 38b 1 Assessment, management and remediation for PFOS and PFOA. Part 2 – health screening levels.

1. Introduction

1.1 Project background

Perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) have been identified as contaminants of emerging concern in Australia (CRC CARE 2014a). These contaminants belong to the broad group of chemicals referred to as per- and poly-fluoroalkyl substances (PFAS).

PFAS are highly persistent and potentially toxic to humans and the environment, and some are bioaccumulative (US EPA 2014, CONCAWE 2016). Due to historical production, PFOS and PFOA are the most widely studied of the PFAS in terms of breakdown of precursor compounds. PFAS have been detected at concentrations of potential concern at a number of sites, particularly where there has been use of firefighting foams (Seow 2013). Awareness of PFOS and PFOA is growing rapidly. Media discussion of PFOS contamination has raised concern regarding the risk that the contamination may pose to the health of exposed persons and the environment.

In general, there is limited and incomplete information surrounding the occurrence, fate and toxicity of PFAS in the Australian environment. Australian guidelines for protecting human health and ecological systems are needed to assess the level of risk that the contamination poses.

This guidance provides a practicable approach to the risk-based assessment, management and remediation of PFAS contamination. It comprises five stand-alone sections:

• Part 1: Background • Part 2: Human health screening levels (this report) • Part 3: Ecological screening levels • Part 4: Application of human health and ecological screening levels • Part 5: Risk based management and remediation of PFOS and PFOA

This section, part 2, details screening levels for the protection of human health and provides guidance on their application and limitations. More detailed guidance on the application of the health screening levels is provided in part 4.

1.2 Purpose of this document

The purpose of part 2 of the PFOS and PFOA guidance is to:

• Review international and Australian literature to evaluate the toxicokinetics (involving toxicity, exposure pathways, bioavailability, bioaccumulation and background exposure) and dose-response of PFOS and PFOA in humans and determine the most appropriate model for assessment (i.e. modified daily intake or toxicokinetics).

• Following recognised Australian processes, use the tolerable daily intake (TDI) to derive screening levels protective of human health for soil (NEPM) and drinking water (NHMRC).

• Discuss important considerations in the application of the screening levels.


Part 2 should be read in conjunction with parts 1, 4 and 5 of the PFOS and PFOA guidance document.

1.3 Basis for development of human health screening levels for PFAS and selection of parameters

Soil, groundwater, surface water and sediment human health screening levels (HSLs) have been derived using a risk-based approach. In Australia, risk assessment procedures for contaminated land should be based on the methodologies outlined in Schedule B4 of the National Environment Protection (Assessment of Site Contamination) Measure 1999, as amended in 2013 (NEPC 2013) (NEPM). This approach is summarised in figure 1.

LIMITATIONS – this report was completed late 2016 and includes a comprehensive review of information identified and available in the latter half of 2015. Consideration has also been given to new information published since that time which has been considered to be particularly important.

The literature review of human health effects (Appendix A) has been separately prepared by CRC CARE and has undergone peer review. In mid-2016 while this report was being finalised enHealth published toxicological and drinking water criteria, and these have been adopted in this guidance.

There is considerable work being undertaken overseas, resulting in new information relevant to the assessment of PFAS. In addition, in its 2016 Statement, enHealth recommended that a review of the EFSA (2008) TDIs be undertaken by FSANZ, and until then the published TDIs and Australian Drinking Water Quality Guideline values for PFOS (+PFHxS) and PFOA should be considered as Interim. Accordingly, this report is considered as Interim, and should be reviewed and updated following any change to enHealth’s and/or FSANZ’s findings regarding PFAS.


Figure 1 Health risk assessment process (refer to NEPM 2013, Schedule B4)

Issues identification

•Why are we doing this assessment? • Is a risk assessment the right type of

decision-making tool? •Who and what are stakeholder

objectives? •What are we trying to find out? •What are the sources and hazards? •What exposure pathways should be

investigated? •What decisions need to be made

and when (urgency of answers)? •Problem formulation

Health risk assessment

Environmental site assessment

Preliminary conceptual site model

Data collection and evaluation

•Collection and analysis of relevant site data

•Development of conceptual site model •Evaluate uncertainties

Working conceptual site model

Other assessments (e.g. ecological, groundwater,

geotechnical)

Toxicity assessment

•Review qualitative and quantitative toxicity information (relevant to reference values)

•Determine appropriate dose-response relationships

• Identify most appropriate quantitative toxicity reference values

•Evaluate uncertainties

Exposure assessment

•Analysis of contaminant releases

• Identification of potential exposure pathways

•Estimation of exposure concentrations for each pathway

•Estimation of contaminant intake for each pathway

•Evaluate uncertainties

Risk characterisation

•Characterise potential for adverse health effects to occur

•Evaluate uncertainty •Undertake sensitivity analysis •Summarise risk information

and evaluation

Risk communication and management

Refined conceptual site model

Engage with stakeholders, risk communication and community engagement


The risk assessment process can be undertaken two ways:

• For an identified source, evaluate the potential health risk (forward calculation), or • For an agreed maximum health risk, estimate a source concentration (reverse

calculation). This source concentration may then be used as a criterion for screening purposes in site contamination investigations.

This process is illustrated in figure 2.

Figure 2 Risk assessment methodology (refer to NEPM 2013, Schedule B4)

The derivation and application of the health screening levels (HSLs) has followed the methodologies for assessment outlined in the National Environment Protection (Assessment of Site Contamination) Measure 1999, as revised in 2013. The NEPM emphasises the importance of formulating a conceptual site model on which to base the assessment of soil and groundwater contamination, and guidance has been provided on the development of a conceptual site model for PFOS and PFOA.

It is intended that the HSLs that have been developed for PFOS and PFOA should be considered similarly to the NEPM HILs/HSLs in forming generic screening levels which, if exceeded, would indicate that further more detailed investigation is required. The HSLs and the information used in their derivation can assist in undertaking this more detailed investigation. Such work should only be undertaken by qualified personnel.

It is emphasised that exceedance of the HSLs does not necessarily imply that the contamination poses an unacceptable risk, and the HSLs should not be used as remediation targets, as this could result in unnecessary remediation.

SourceConcentration

Toxicity Assessment

Exposure Assessment Risk

HSL Toxicity Assessment

Exposure Assessment

Acceptable Risk

FORWARD CALCULATION

REVERSE CALCULATION


1.4 Conceptual site model

A conceptual site model (CSM) is a representation of site-related information regarding contamination sources, receptors and exposure pathways between those sources and receptors. The development of a CSM is an essential part of all site assessments and provides the framework for identifying how the site became contaminated and how potential receptors may be exposed to contamination in either the present or the future.

The development of the CSM is an evolving process, where the understanding of the CSM is updated as new information on the site is obtained and understanding of the issues at the site is refined. As well as being the fundamental basis for assessing health risks, the CSM is also important for use in identifying data gaps, and in situations where targeted investigations may be required.

The development of CSMs should follow the approach presented in Schedule B2 of the NEPM. Further discussion regarding the development of a CSM for PFOS and PFOA contaminated sites is provided in part 1 and 5 of this guidance.


2. PFOS and PFOA toxicity assessment

2.1 Overview

This section provides an overview of the toxicity of PFOS and PFOA to humans. A detailed literature review of the effects of human exposure to PFOS and PFOA is presented in Appendix A (Expert review: human health criteria), with key findings summarised below.

2.2 Summary of health effects

A range of studies has been undertaken into the possible health effects of PFOS and PFOA in humans. Most human studies have looked for a relationship between levels of PFAS in the blood and a health effect, but results have been inconsistent, with some studies finding associations, but others investigating the same health effect finding no association. Additional difficulties arise when seeking to extrapolate from animal to human studies, as humans and animals have been found to react differently to PFOS and PFOA, with profound differences in the toxicokinetics observed (ATSDR 2015).

2.2.1 Acute toxicity

The limited data available for humans indicates that acute toxicity is not observed following high exposure to PFOS or PFOA through inhalation, ingestion, dermal or ocular contact (ATSDR 2015). Animal studies have found a moderate acute oral toxicity, with effects on the liver and gastrointestinal tract (PHE 2009) and PFOS being more toxic than PFOA in fresh water organisms (Ji et al. 2008, Li 2009).

2.2.2 Chronic exposure

Animal studies have indicated that chronic exposure of mice, rats and monkeys to PFOS can result in increased liver weight, liver cell hypertrophy, histopathological changes to lungs, decreased hormone level, decreased reproductive outcome, and development delays. Chronic exposure to PFOA resulted in increased liver weight, and reduced immunoglobulin M antibody titres. A summary of these studies and findings is presented in table 6 and table 7 of Appendix A.

Prior to 2014, limited information was available regarding the effects of chronic exposure to PFOS or PFOA in humans, with epidemiological studies having focused on people occupationally exposed to PFOS/PFOA. However, the number of human studies addressing potential PFOS/PFOA toxicity has been growing recently, with more than 50 epidemiological studies published from 2015−16 onwards. A number of epidemiological studies have also been conducted as cross-sectional or longitudinal analyses of routine medical surveillance at PFOS/PFOA production facilities, focused on occupational exposure. More recently epidemiological studies have also been conducted on general populations, attempting to assess the possible correlations between PFAS concentrations in the human body and various health endpoints. Based on these studies the following observations have been made (see appendix A: tables 8 and 9):

• Modest positive associations of PFOS/PFOA concentrations in serum with cholesterol (Frisbee et al. 2010).


• Higher serum levels of PFOA are associated with elevated levels of uric acid in the blood (which can be a risk factor for hypertension). For PFOS similar but less pronounced trends are found (Steenland et al. 2010).

• No significant association was observed with blood cell counts or thyroid hormones (Olsen et al. 2003).

• There is some, but much less consistent, evidence of a modest positive correlation with increased liver enzymes in the blood (Butenhoff et al. 2002, Lau et al. 2006).

• Mixed information is available regarding the carcinogenicity of PFOS and PFOA. No significant correlation between PFOS exposure and increased risk of cancer has been reported. However, higher PFOA serum levels have been associated with kidney and testicular cancer, and PFOA is classified by the IARC Monographs as ‘possibly carcinogenic to humans’ (Group 2B) (IARC 2015).

• No significant associations reported between serum PFOS/PFOA concentrations and reproductive or developmental outcomes, with the exception of one study by Fei et al. (2009) which found that plasma levels of PFOS/PFOA may reduce fecundity in the general population.

• PFOS and PFOA are not considered genotoxic (EFSA 2008).

2.3 Toxicokinetics

PFOS and PFOA uptake occurs through oral, inhalation and dermal exposure. Ingestion of food or contaminated drinking water is considered the primary route of human exposure, with inhalation also an exposure pathway of concern (ATSDR 2015, US EPA 2016a, 2016b). Uptake via dermal exposure generally appears to be less of a concern, though information is limited (Kudo & Kawashima 2003, ATSDR 2015). Infants may be exposed through breast milk, and young children through hand-to-mouth activities from treated carpets (ATSDR 2015).

PFOS and PFOA preferentially bind to proteins and accumulate primarily in the liver, blood, serum and kidneys (Stahl et al. 2011). They are not metabolised in mammals, and are therefore removed only by excretion (Stahl et al. 2011). In mammals, biological half-lives differ among species and between genders due to differences in renal clearance rates (Kudo & Kawashima 2003). More recently, excretion via milk in lactating mammals has also been indicated to be a major elimination pathway. For humans the reported mean half-lives for PFOS and PFOA are 5.5 years and 3.2 years respectively, with the following ranges reported:

• PFOS − Animals: 100 days in male and female rats (Gibson & Johnson 1979,

Chang et al. 2008) to 150 days in male and female monkeys (Perkins et al. 2004).

− Humans: 4.3 years in US cross-sectional study (age 10 to 69) (Olsen et al. 2012) to 8.7 years for primiparous (giving, or having given birth for the first time) Swedish women (Gebbink et al. 2015).

• PFOA − Animals: 2 to 4 hours in female rats (Gibson & Johnson 1979,

Chang et al. 2008) to 30 days in female monkeys (Perkins et al. 2004) − Humans: 2.1 years in Chinese adults (Olsen et al. 2007) to 4.1 years in US

infants (Spliethoff et al. 2008).


2.4 Background levels

A national survey in Australia showed PFOS and PFOA detected in drinking water at levels of 0–15.6 ng/L and 0–9.6 ng/L with corresponding uptake rates estimated to be 0–11 ng PFOS per day and 0–13 ng PFOA per day (Thompson et al. 2011). Limited information regarding other uptake pathways, such as consumption of fish and aquatic species is available in Australia. A study conducted in 2002–2003 across Australia of 3802 serum samples collected from both genders, in five age groups found mean blood concentrations of 20.8 ng/mL and 7.6 ng/mL for PFOS and PFOA respectively (Kärrman et al. 2006). A further three Australian studies were identified encompassing 2006–2007 (Toms et al. 2009), 2008–2009 and 2010–2011 (Toms et al. 2014) (refer to table 1 in appendix A for details). Data from these studies show that serum PFOS and PFOA concentrations in the general Australian population appear to have decreased from 2002 to 2011, with mean serum levels in the 2010–11 study of 10.2 ng/mL PFOS and 4.5 ng/mL PFOA. This general decrease in the concentrations of PFOS and PFOA in blood serum is consistent with trends observed the USA (ATSDR 2015), reflecting the reduction in production and use of products containing PFOS and PFOA.

2.5 Summary of toxicity reference values

For chemicals where there is an exposure level below which no toxic effect occurs (threshold toxicity), the estimated daily intake can be compared with a threshold toxicity reference value (TRV). The NEPM defines the TRV as a value representing a dose of a chemical that will not cause adverse effect over a lifetime of exposure. The TRV can refer to any appropriate measures of tolerable daily intake, and can include doses derived for different applications. References may be tolerable daily intakes (TDI), as for drinking water guidelines, or acceptable daily intakes (ADI) which are typically used in food/drug guidance, or USEPA reference doses (RfD) (NEPM).

Threshold TRVs are specific to the route of exposure (i.e. ingestion, inhalation, dermal), and can be quite different for different exposure routes. They are typically available for ingestion and inhalation exposure routes, and occasionally for dermal exposure.

Two types of data may be used to derive a human TDI – human data and animal data. Epidemiological studies have indicated a link between PFOS/PFOA exposure and public health. However, most studies only show the odds ratio between control and exposure groups, or the potential association between exposure and risk. Thus far, epidemiological studies have concluded there is a weak positive association between serum PFOS/PFOA and increased serum cholesterol and uric acid levels (appendix A, table 8 and table 9). In addition, a few but inconsistent positive correlations have been observed showing toxicity to liver enzymes. Although recent studies associate exposure to PFAS with adverse health outcomes, most studies are cross-sectional analyses and therefore the data are insufficient to draw unambiguous conclusions about the effects of PFOS/PFOA in the progress of a particular disease. Therefore, it is difficult at this stage to establish a TDI based on human studies.

Despite the fact that they exclude the impacts of extraneous factors, laboratory-based animal studies can mimic a pure exposure scenario and provide useful dose-response data, although interspecies uncertainties will remain. Due to the pharmacokinetic


differences seen with PFAS in different species, administered doses are not directly comparable. Administered doses are normalised to a human equivalent dose (HED) using the species-specific half-lives and volume of distribution as discussed in appendix A, section 5.1.6.

US EPA and IPCS have highlighted two types of uncertainty when deriving human TDI from animal studies – determination of the appropriate most sensitive critical effect for the point of departure (PoD), and uncertainty introduced in extrapolation from the PoD to a human TDI incorporating differences between species, routes of exposure, doses and durations (Dong et al. 2015).

A summary of TDIs (including draft and provisional TDIs) from a range of international organisations is provided in table 1.

Table 1 Comparison of TDIs adopted internationally Organisation PFOS PFOA PoD

(µg/kg/d) UF TDI (µg/kg/d)

PoD (µg/kg/d) UF TDI

(µg/kg/d)

UK COT (2006a, b) 30 (NOAEL) 100a 0.3 300

(BMDL) 100a 3

EFSA (2008) 30 (NOAEL) 200b 0.15 300

(BMDL) 200b 1.5

US EPA (2009) 30 (NOAEL) 390 0.08 460

(BMDL) 2,430 0.2

BfR Germany (Priestly 2015) NA NA 0.1 NA NA 0.1

US EPA (2016a, b) 0.51 (HED) 30 0.02 5.3

(HED) 300 0.02

Danish EPA (2015) 33 (BMDL10) 1230 0.03 3

(HED) 30 0.1

TDI – tolerable daily intake, PoD – point of departure, UF – uncertainty factor, NOEL – no observable effects level, BMDL – benchmark dose modelling, HED – human equivalent dose, NA – none available, a. inter- and intraspecies uncertainties, b. UF of 100 for inter- and intraspecies uncertainties and UF of 2 for compensate for uncertainties in connection to the relatively short duration of the key study and the internal dose kinetics

In 2006, the UK committee on toxicity (COT) established an RfD of 0.3 µg/kg/day for PFOS (table 1) (UK COT 2006a), based on serum T3 levels being decreased in a 26-week monkey study (Seacat et al. 2002). Two types of uncertainty factors (UFs) were further considered – inter-species UF of 10 and intra-species UF of 10. Subsequently, the European Food Safety Authority (EFSA) issued an RfD of 0.15 µg/kg/day with an additional UF of 2 to account for the monkey study not representing life-time exposure (EFSA 2008). Similarly, considering hepatic effects in male rats (Palazzolo 1993, Perkins et al. 2004), the UK COT (2006b) and the EFSA (2008) proposed the PFOA RfDs of 3 µg/kg/day and 1.5 µg/kg/day (table 1) respectively.

For its 2009 Provisional Health Advisory, the US EPA used a chemical-specific adjustment factor method to account for the interspecies uncertainties and therefore estimate a TDI. These adjustment factors account for the differences in volume of distribution and half-lives of the compounds in different species. Specifically, a factor of 13 was used for PFOS, for extrapolation from monkey to human, and a factor of 81 was employed for PFOA, for extrapolation from rat to human. Together with the uncertainty factor of 10 to represent the intraspecies uncertainties and uncertainty factor of 3 to denote the toxicodynamic uncertainties, the total uncertainty factors in US


EPA’s 2009 version are estimated to be 390 for PFOS and 2430 for PFOA (US EPA 2009).

In the 2016 health effects documents for PFOS and PFOA the pharmacokinetic differences between species were accounted for by deriving a HED from modelled animal serum values (US EPA 2016a, b).

2.6 Interim guidance published by enHealth

In June 2016 enHealth issued a Statement interim national guidance on human health reference values for per- and poly-fluoroalkyl substances for use in site investigations in Australia.

The enHealth statement considered the approach taken by various international bodies, including the European Food Safety Authority (EFSA), USEPA, ATSDR, DME, German Ministry of Health Drinking Water Commission and Federal Environment Agency, Swedish Environmental Protection Agency, United Kingdom Committee on Toxicity of Chemicals in Food and the Minnesota Department of Health, and also the approach being taken by CRC CARE in the preparation of this report.

enHealth concluded that the 2008 EFSA approach is consistent with relevant Australian science policy, and recommended TDI values for PFOS (+ PFHxS) and PFOA of 0.15 µg/kg/day and 1.5 µg/kg/day, respectively.

In its statement, enHealth advised:

• enHealth considers that the 2008 European Food Safety Authority’s (EFSA) derivation of Tolerable Daily Intake (TDI) values for perfluorooctane sulfonate (PFOS) and perfluorooctanoic acid (PFOA) is appropriate as interim national guidance for use in site investigations in Australia.

• enHealth considers the use of the 2008 EFSA approach is consistent with relevant Australian science policy, as per enHealth’s 2012 publication, Environmental Health Risk Assessment, Guidelines for Assessing Human Health Risks from Environmental Hazards. enHealth also considers the 2008 EFSA approach is relevant to the exposure pathways of concern in Australia, which include contaminated food. EFSA TDI values for PFOS and PFOA are used for the assessment and management of these hazards in the food chain in Europe. This approach was validated by EFSA’s 2012 comprehensive assessment of dietary exposure to the PFAS group of chemicals.

• enHealth notes that the EFSA, USEPA and US ATSDR approaches considered the same collection of toxicological studies. EFSA’s method to derive a TDI uses a similar approach to that generally used in Australia.

• enHealth notes that the US EPA has recently released the US EPA 2016 Health Effects Support Document for Perfluorooctane Sulfonate (PFOS) and US EPA 2016 Health Effects Support Document for Perfluorooctanoic Acid (PFOA). enHealth considers that the final 2016 versions of the US EPA Health Effects Support Documents do not change its recommendation to use the EFSA TDI. The US EPA 2016 versions contain the same studies and methods as the 2014 draft documents considered at the April 2016 enHealth workshop and therefore do not invalidate enHealth’s reasoning in recommending the EFSA TDI values.


• enHealth recommends that the EFSA TDI values are used by jurisdictions in site investigations in Australia to determine interim drinking water and recreational water guideline values using the methodology described in chapter 6.3.3 of the National Health and Medical Research Council (NHMRC) Australian Drinking Water Guidelines. enHealth considers that calculating an interim drinking water guideline value in this way is more appropriate for the Australian context than using the 2016 US EPA Drinking Water Health Advisories.

• The interim drinking water guideline values are not intended to be a guide for drinking water utility providers across Australia, but rather for use to confirm the quality of drinking water supplies potentially affected by specific instances of site contamination.

• enHealth has also considered the available information on perfluorohexane sulfonate (PFHxS) and agreed that the EFSA TDI for PFOS also be applied to PFHxS exposures. In practice this means PFOS and PFHxS exposures should be summed and the total compared to the TDI for PFOS.

• enHealth recommends that Food Standards Australia New Zealand (FSANZ) undertakes an assessment of the available toxicity data on PFOS, PFOA and PFHxS and publishes relevant reference values in the Australia New Zealand Food Standards Code. Values published by FSANZ will immediately replace interim toxicity reference values recommended by enHealth.

The recommended enHealth interim values are provided in table 2.

Table 2 enHealth interim toxicity reference values

Toxicity reference value PFOS/PFHxS PFOA Tolerable daily intake (µg/kg/d) 0.15 1.5 Drinking water quality guideline (µg/L) 0.5 5 Recreational water quality guideline (µg/L) 5 50

2.7 Procedural review of the health reference values established by enHealth

In June 2016, enHealth, the peak environmental health body in Australia, selected the European Food Standards Authority (EFSA) health reference values (HRVs) for PFOS (and PFHxS combined) and PFOA for application as an interim measure in Australia.

In September 2016, the Procedural review of health reference values established by enHealth for PFAS (Review) noted that adoption of EFSA (2008) HRVs for PFAS in drinking and recreational water was appropriate and is protective of public health.1 The review considered more recent (and significantly lower) TDIs, such as those published by the US EPA, and went on to state that 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 (Bartholomaeus 2016, section 6.3).’

1 Prof (Adj) Andrew Bartholomaeus, 30 August 2016, available at www.health.gov.au/internet/main/publishing.nsf/Content/44CB8059934695D6CA25802800245F06/$File/Procedural-review-health-reference-values-PFAS.pdf


The review stated that the:

‘adoption of the EFSA HRVs and their use to derive Australian drinking water guideline values as an interim measure pending FSANZ review can be concluded to be appropriate, 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 (Bartholomaeus 2016, section 7).’

It noted that differences between the EFSA and US EPA assessments and derivations of HRVs are not due to new data (section 5.1).2 The most substantial difference 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 (section 5.2). The review states that the US EPA’s ‘modelling approach has potentially replaced one set of uncertainties with another’. Further details of the sources of variation between the US EPA and EFSA approaches are considered in the Review. Additional review is being undertaken by FSANZ.

For drinking water, the review noted that the long half-life of PFAS in humans means that the ‘systemic (i.e. internal) exposure to PFAS is determined by oral (or other routes of) exposure over long periods of time (section 6.3).’ Given this, the review states that:

‘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 (Bartholomaeus 2016).’

It cannot be ascertained from the review whether lowering PFOS/PFOA guidance levels for soils (e.g. HILs/HSLs) would lead to significantly different public health outcomes for exposure to PFOS/PFOA contaminated soil.

In Australia, it is customary to adopt TDI values which have been recommended by national bodies such as enHealth. enHealth have selected the EFSA HRVs on an interim basis, subject to further review by FSANZ. It is expected that practitioners and risk assessors will take into account any future recommendations from enHealth (as well as any recommendations from state and territory authorities), in addition to other pertinent information as it becomes available.

2 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 (section 5).


3. Risk characterisation

3.1 Overview

The risk characterisation component is the link between the toxicity assessment (the potential for health effects) and the exposure assessment (the predicted exposure) that determines whether the chemicals of potential concern pose an unacceptable health risk and what measures need to be taken to reduce the risk.

In a forward risk assessment, a known contamination source quantity is utilised to estimate exposure and, combined with the toxicology criteria, provides an estimate of risk.

When deriving screening levels (i.e. HSLs) the equations are reversed, such that a target acceptable risk and the relevant TDIs are used to estimate the maximum allowable exposure, and hence the allowable contamination levels (the HSL).

Refer to section 1.3 for further details on risk assessment methodology.

3.2 Threshold toxicology

As discussed in section 2.5, the toxicological criteria for PFOS and PFOA may be represented by a threshold toxicity reference value (TRV). The TRV for PFOS and PFOA considers long-term (chronic exposure) levels and biokinetics (the rates of contaminant intake, accumulation and loss), such that blood serum levels remain below an accepted level.

Threshold health effects may be assessed by direct comparison of the chronic daily intake (CDI) with the tolerable daily intake (TDI). In assessing threshold toxicity effects, total exposure to all sources must be considered, and the background intake is required to be included in the assessment of risk or the development of screening level.

For assessing risk the following equation may be used:

𝑅𝑅𝑅𝑅 =𝐶𝐶𝐶𝐶𝐶𝐶𝑐𝑐 + 𝐶𝐶𝐶𝐶𝐶𝐶𝑏𝑏𝑏𝑏𝑐𝑐𝑏𝑏

𝑇𝑇𝐶𝐶𝐶𝐶

Where:

• RQ = risk quotient (dimensionless) • CDIc = chronic daily intake from contamination (mg/kg/d) • CDIback = chronic daily intake contribution for background exposure (mg/kg/d) • TDI = tolerable daily intake (mg/kg/d) (refer to section 2.5)


For deriving a HSL the following equations may be used:

𝐶𝐶𝐶𝐶𝐶𝐶𝑐𝑐 = �𝑅𝑅𝑅𝑅𝑡𝑡𝑏𝑏𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡 × 𝑇𝑇𝐶𝐶𝐶𝐶� − 𝐶𝐶𝐶𝐶𝐶𝐶𝑏𝑏𝑏𝑏𝑐𝑐𝑏𝑏

or

𝐶𝐶𝐶𝐶𝐶𝐶𝑐𝑐 = 𝑅𝑅𝑅𝑅𝑡𝑡𝑏𝑏𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡 × 𝑇𝑇𝐶𝐶𝐶𝐶 × 𝐶𝐶𝐶𝐶

Where:

• RQtarget = the maximum risk quotient • CC = contamination contribution to exposure (dimensionless)

The HSLs for PFOS and PFOA may be calculated from CDIc using considerations of exposure (refer to section 5).

An RQ less than 1 indicates that adverse health effects resulting from that particular chemical are not likely to occur. Health investigation levels (HILs) in the NEPM are based on a RQ of 1 (i.e. RQtarget = 1) (NEPC 2013).

The contribution to contamination is a fraction of the portion of the TDI allowed from contamination sources. In assessing site contamination, the NEPM has generally taken the approach that the proportion allowed is related to the level of background exposure. Chemicals for which there is a low background exposure are generally allowed a greater proportion of the TDI to arise from contamination, whereas chemicals for which there is considerable background exposure are usually allowed a lower contribution. The contribution to exposure by PFOS and PFOA soil contamination should not exceed 1.0 (the target risk quotient is 1.0), as the background exposure contribution is less than 1% of the adopted TDI (refer to section 4.6 for more details)


4. Exposure assessment and development of HSLs

4.1 Overview

The exposure assessment documents the key exposure routes with regard to chemical migration properties and human behaviour characteristics. Quantification of exposure and dose is undertaken and combined with toxicity information, to derive soil and water health screening levels for a number of particular exposure scenarios.

4.2 Fate and transport

Certain physical properties of chemicals can have a significant influence on fate and transport mechanisms. These in turn determine the key exposure media and exposure routes by which humans and the environment may be exposed to contamination.

4.2.1 Chemical stability

PFOS and PFOA are manufactured chemicals that are chemically and biologically stable and hence are persistent in the environment; being resistant to biodegradation, atmospheric photooxidation, direct photolysis, and hydrolysis (US EPA 2014). PFOS and PFOA can also be formed from related substances or precursor compounds by microbial degradation or larger organism metabolism (e.g. rainbow trout transform perfluorinated acids to PFOS) (CRC CARE 2014b).

4.2.2 Groundwater migration potential and characteristics

PFOS and PFOA are moderately soluble, and are stable in the environment as evidenced by their long half-lives in water (PFOA half-life reported as 92 years; PFOS half-life >41 years on basis that over study length did not identify change in concentration) (DEPA 2015). Although persistent in groundwater and surface waters, they have been found to partition from the groundwater column to organic matter rich sediments and soil particles due to their propensity to sorb to natural organic matter and soil surfaces.

4.2.3 Biomagnification in food chain

PFAS are extremely thermally, biologically and chemically stable, and exhibit hydrophobic characteristics (Konwick et al. 2008). The carbon-fluorine bond in PFAS is the strongest halogen-carbon bond, shielding the carbon chain bonds from reactive species resulting in PFAS being highly persistent in the environment (Boudreau et al. 2003a, 2003b, Hazelton et al. 2012). PFOS and PFOA have been shown to be metabolically and chemically inert, resisting both biotic and chemical degradation (Boudreau et al. 2003a, 2003b). Furthermore, PFOS and PFOA have been found to biomagnify in food webs, as evidenced by PFOS detection in humans and wildlife (Konwick et al. 2008).

PFOS and PFOA exhibit tendencies to bind to blood serum proteins, alter fatty acid metabolism, and adversely affect cellular membranes and intercellular communication (Konwick et al. 2008). PFOS accumulates in hepatic tissues, and tissue-specific accumulation in freshwater and marine fish has been observed. As a result, PFOS has the potential for biomagnification in aquatic food webs (Sharpe et al. 2010,


Hazelton et al. 2012). In the marine food chain the potential for bioaccumulation of PFOS and PFOA differs, with PFOS having the potential to biomagnify along the food chain whereas PFOA does not have this potential (Yang et al. 2012). There is a potential for PFOA to bioaccumulate in benthic organisms, such as found in inter-tidal flats, through direct ingestion of particles (Yang et al. 2012).

As a result of their persistence and potential for bioaccumulation, PFOS and PFOA have been detected in higher trophic level animals such as fish, dolphins, seals, polar bears and birds (Butt et al. 2008, Fair et al. 2013, Greaves et al. 2013, Leat et al. 2013, Lucia et al. 2015). PFOS has a higher tendency for bioaccumulation compared to PFOA due to its longer perfluoroalkyl chain length (eight fluorinated carbon chain compared to seven and sulfonate group (Thompson et al. 2011)). PFOS has also been shown to bioaccumulate and biomagnify in fish and piscivorous birds in higher concentrations than PFOA (Lucia et al. 2015). The biomagnification factor ranges from 1.4 to 17 in predatory birds and mammals (US EPA 2014). PFOS is the only PFAS that has been shown to accumulate to levels of concern in fish tissue, with an estimated kinetic bioconcentration factor in fish ranging from 1,000 to 4,000 – the time to reach 50% clearance in fish has been estimated to be around 100 days (US EPA 2014).

Research also indicates that PFOS and PFOA are readily taken up by garden produce. Blaine (2014) found that edible plants grown in soil contaminated with PFAS accumulate short-chain PFAS to a greater extent than longer-chain PFAS, and fruit crops accumulate less PFOA than shoot or root crops. The amount of PFOA accumulated by plants shows a negative correlation to the organic carbon content of the soil and carbon tail length. Blaine (2014) also showed that irrigation of food crops using reclaimed water also results in the accumulation of PFOA; the amount accumulated is dependent on the concentration in the irrigation water. Using irrigation water containing 0.02−4 µg/L of PFAS, 0.2 µg/L of PFOS in lettuce and 0.05 µg/L in strawberries was detected (Blaine 2014). Irrigation with contaminated groundwater containing 10 to 40 µg/L PFAS resulted in accumulation of 1 µg/L in lettuce. The research conducted by Blaine (2014) supports that of Lechner and Knapp (2011) and Stahl et al. (2008), which found that PFOS and PFOA were detected in the vegetative portions of carrots, cucumbers, maize, wheat, potato and oats. All the studies discussed here reported that concentrations of PFOA were higher than PFOS in the plants.

A more detailed discussion of biomagnification and bioaccumulation is presented in part 3 of the guidance.

4.3 Soil receptor populations

HSLs have been developed for four receptor groups, consistent with the NEPM. The HSLs are human health-based, and do not consider ecological endpoints (which are derived separately, and discussed in part 3 of the guidance).

The generic land use scenarios considered in the development of the HSLs are as per HILs in the NEPM:

• HSL-A – Residential scenario with garden/accessible soil (home-grown produce <10% fruit and vegetable intake), (no poultry), includes childcare centres, preschools, primary schools.


• HSL-B – Residential with minimal opportunities for soil access includes dwellings with fully and permanently paved yard space such as high-rise buildings and flats.

• HSL-C – Public open space such as parks, playgrounds, playing fields (e.g. ovals), secondary schools and footpaths. It does not include undeveloped public open space (such as bushland and reserves) which should be subject to site-specific assessment, where appropriate.

• HSL-D – Commercial/industrial such as shops, offices, factories and industrial sites.

As for HILs in the NEPM, screening levels developed for metals and organic substances apply generally to a depth of 3 m below the surface for residential use. Site-specific conditions should determine the depth to which HSLs apply for other land uses.

The various receptor groups are as follows:

4.3.1 Residential land (HSL-A and HSL-B)

Residential land can involve development varying from separate lot low-density residential dwellings to high-density apartment blocks. The NEPM classifies HIL-A as low-density residential land with direct access to soils, and HIL-B as high-density residential land where there is limited opportunity for direct access to soil. Medium-density residential falls between these two categories and is dependent on the property design; if there is exposed (uncovered) land it may be classified as low density, and if there is limited access to soils through permanent paving or structures then it may be considered to be similar to high-density residential.

Residents have been divided into the following age categories as suggested by enHealth (2012a):

• young children (birth to 5 years old inclusive) • older children and adults (6 years old to adult)

It should be noted that the HSL-A does not truly reflect the assumptions in the NEPM for HIL-A as the health screening level does not include the 10% consumption of food from home-grown produce. This is due to the uncertainty of modelling plant uptake of PFAS based on the current limited available information. Further discussion on the uncertainty is discussed in section section 4.3.4 (consumers of home grown produce and animal products) and 4.4.3 (irrigation water).

4.3.2 Public open space (HSL-C)

Public open space includes parklands and ovals that the general public can access and potentially spend significant time. Receptors of interest are the same age groups as those defined for residents.

4.3.3 Commercial/industrial (HSL-D)

Commercial and industrial land can vary from small office blocks, large multi-storey buildings and industrial warehouses, to open space storage grounds and training grounds.

An adult worker has been selected as the primary receptor for this land use. While the public (including children) may visit the site for short periods, it is assumed that the


exposures and risk that would be involved for the public will be less than those for workers present at the site on a full-time working basis, and will not be limiting.

It is important to note that commercial/industrial HSLs cannot be applied where there may be prolonged exposures to sensitive receptors via direct contact with contaminated soil. While the HSLs are protective of adult workers, the HSLs may not be adequately protective of children in schools or childcare centres, or inpatients and residents in health care facilities, nursing homes and hospitals or other similar land use that in some jurisdictions may be permitted under a commercial or industrial zoning. Occupational exposures may be best addressed separately.

4.3.4 Consumers of home grown produce and animal products

As indicated in section 4.2.3, PFOS and PFOA can accumulate and magnify in the food chain. As well as uptake by produce that might be ultimately used for human or stock consumption, PFOS and PFOA are known to bind to proteins and hence biomagnify through milk, egg and meat products from chickens, ducks, milk and cattle/goat, and any other stock animals.

HSLs that provide protection for human consumption of produce and animal products (excluding fish) have not been derived due to complexities associated with the rate of uptake and biomagnification. These vary from site to site, species to species, and as such it is instead recommended that site-specific sampling of produce and animal products be undertaken and estimates made of the likely rate of consumption of these products so that a determination of the intake of PFOS and PFOA that might result can be made. This can be compared to the TDIs.

4.3.5 Surface water receptors – consumption of seafood

Consumption of seafood is significant with regards to PFAS. Bioaccumulation factors over 1000 for PFOS in fish result in water criteria significantly lower than drinking water criteria. RIVM (2010) derived a maximum permissible concentration (MPC) of 0.65 ng/L for freshwater for protection of human consumers of fish. Because of the potential for such bioaccumulation, this could be the limiting exposure scenario in some situations and require consideration. Further details on the derivation of water HSLs protective of seafood consumption are provided in section 4.7.5 .

4.4 Groundwater receptor populations

Beneficial uses of groundwater have been considered in the development of groundwater HSLs. The following sections detail the considered scenarios. Note screening levels for ecological protection (ESLs) have been regarded separately and are discussed in part 3 of the guidance.

4.4.1 Drinking water

The enHealth statement (July 2016) comprised an interim drinking water quality guideline, derived using the methodology presented by NHMRC (refer to sections 2.6 and 4.9.4).

4.4.2 Primary contact recreation


Primary contact recreation is an extractive beneficial use of groundwater that in some situations requires assessment (e.g. extraction of groundwater for swimming pools). The NHMRC Guidelines for Managing Risks in Recreational Waters (NHMRC 2008) specify that as a general principal, a factor of 10 may be applied to the drinking water standard this takes into account that ingestion of 200 mL of water during swimming activities (as opposed to the 2 L/d used to derive the drinking water guideline) is likely to be an overestimate.

In June 2016 enHealth issued a statement interim national guidance on human health reference values for per- and poly-fluoroalkyl substances to be used in site investigations in Australia. In this statement enHealth published an interim recreational waters quality guideline, derived using the methodology presented by NHMRC (refer to sections 2.6 and 4.9.4).

4.4.3 Irrigation/agriculture

Use of groundwater for irrigation may lead to direct exposure during watering activities, and indirect exposure through uptake of contaminants in plants that are then consumed. Agricultural soil may be contaminated by PFAS, for example, if PFAS affected biosolids have been applied.

With respect to direct exposure during watering activities, adults may be exposed minimally whereas children may receive greater exposure during play under sprinklers. The exposure by children playing under sprinklers is likely to be less than exposure experienced during swimming (primary contact recreation). As discussed in section 4.5.2, as an indication of risk during watering activities, the factor of 10 applied to drinking water criteria may be considered protective of this exposure.

Indirect exposure through accumulation of PFOS and PFOA in soil and uptake by plants that are to be consumed is possible, and requires consideration. As a result of the variability of uptake in plants, a HSL for protection of irrigation and agriculture has not been derived and should be assessed on a site-specific basis. This may include estimates of the accumulation of PFAS that will occur in soil through the use of contaminated water, and direct sampling of plants that are irrigated with contaminated water and are growing in contaminated soil, to determine the concentrations of PFOS and PFOA that result.

4.4.4 Stock water

As discussed in section 4.2.3, PFOS and PFOA can biomagnify in the food chain. Exposure of stock to contaminated water can lead to increased concentrations of PFOS and PFOA in products derived from stock, such as in meat, milk, and eggs.

The ANZECC (2000) water quality guidelines advise that for pesticides and other organic contaminants the drinking water guidelines are recommended to be used for stock water in the absence of adequate information derived specifically for livestock under Australia and New Zealand conditions. Generally, this would be protective of stock health. Despite there being an increase of up to five times the consumption rate of water to body weight ratio, compared to average human intake, there is enough conservatism built into the other parameters, including toxicology, for the drinking water screening levels to be considered appropriate for stock health.

However, the majority of chemicals considered by ANZECC (2000) are not bioaccumulative in the way PFAS is. Whereas pesticides like DDT tend to be lipophilic


and bind strongly to fatty tissue, PFAS strongly binds to proteins. There have also been numerous studies on the bioaccumulation potential of pesticides in birds and stock animals, and their products (i.e. milk, meat and eggs). In this regard, ANZECC has identified that for pesticides the drinking water criteria is protective of stock health, but also human consumption of animal products. The same cannot be said for PFAS. There is limited information on the relationship between intake rates by animals and measurable PFAS in animal products and as such, there is currently significant uncertainty whether the drinking water screening levels for PFAS would adequately protect human health from consumption of animal products (a HSL for stock water has not been derived).

This beneficial use requires assessment on a site-specific basis and, where relevant, consideration should be given to direct sampling of PFOS and PFOA in products from animals that have been drinking contaminated water.

4.4.5 Consumption of seafood

The assessment of groundwater impacts on surface water bodies with regard to bioaccumulation and biomagnification should be undertaken on a site-specific basis. Biomagnification of PFOS and PFOA in fish is discussed in detail in section 4.2.3 and 4.2.7 for surface water. Where HSLs for surface waters are applied to groundwater, additional considerations such as groundwater-surface water interactions, benthic-biota bioaccumulation factors (see section 4.7.5), discharge flux from groundwater to surface water, and the flow rate of surface water, may assist site-specific assessments.

4.4.6 Aquaculture

Aquaculture refers to the breeding, rearing, and harvesting of plants and animals in all types of water environments including ponds, rivers, lakes, and the ocean. Aquaculture includes the production of seafood from hatchery fish and shellfish grown to market size in ponds, tanks, cages, or raceways. Stock restoration is a form of aquaculture in which hatchery fish and shellfish are released into the wild to rebuild wild populations or coastal habitats such as oyster reefs. Aquaculture also includes the production of ornamental fish for the aquarium trade, and growing plant species used in a range of food, pharmaceutical, nutritional, and biotechnology products.

In some regions, aquaculture is a regulated beneficial use of groundwater. Both freshwater and slightly brackish groundwater is suitable for aquaculture depending on the plant or animal species.

For this beneficial use of groundwater, the surface water HSL for consumption of seafood should apply.

4.5 Exposure pathway assessment

The key exposure pathways by which humans may be exposed to PFOS and PFOA are discussed in the following sections.

4.5.1 Oral exposure

Ingestion is the primary route of exposure considered for drinking water and contaminated soil/dust. Oral ingestion may also be significant with respect to


consumption of plants and animals (including seafood) impacted through the food chain.

Currently the majority of soil screening level for PFOS and PFOA adopted by international jurisdictions only consider the ingestion pathway. Exposure to surface soils occurs primarily through incidental hand-to-mouth activities, such as touching food followed by ingestion. Young children also have a tendency to have increased exposure from pica (eating soil), greater hand-to-mouth activity (including crawling on treated carpets that may contain PFOS) and reduced hygiene (i.e. washing of hands).

With respect to bioavailability, there is little in the way of information on bioavailability of PFOS and PFOA with respect to human exposure. However, Gibson and Johnson (1979) reported that 95% of the radioactively labelled PFOS dose (4.3 mg/kg body weight) and 93% of the labelled PFOA dose (11 mg/kg body weight) were resorbed by male rats within 24 hours, based on measuring the radioactivity in faeces and in the digestive tract. This suggests that nearly all of the applied dose of PFOS/PFOA was absorbed.

With respect to bioavailability in soil and the effect of ageing, limited information has been found on the effects of ageing on the bioavailability of PFOS and PFOA to soil-dwelling organisms, and there were insufficient data to derive an application factor for ageing. For the purposes of an initial assessment, it has been assumed that ageing does not reduce the bioavailability of PFOS and PFOA in soil. This assumption is supported by the observation of long-range global distribution of PFAS, which would occur over significant time periods and hence result in ageing of contaminants. which still results in accumulation in exposed organisms remote from the original source (Houde et al. 2006, Butt et al. 2008, Keller et al. 2012, Fair et al. 2013, Greaves et al. 2013, Leat et al. 2013, Lucia et al. 2015).

4.5.2 Dermal absorption

Franko et al. (2012) reported that dermal absorption of PFOA in testing on mice occurred with very high dermal application in an acetone vehicle, as evidenced by increased blood levels of PFOA. However, it is the unionised form of PFOA present under acidic conditions that is likely to absorb (Franko et al. 2012), with a skin permeability rate three orders of magnitude higher than ionised PFOA (ATSDR 2015). It can be concluded that under normal pH conditions the potential for absorption via dermal exposure is limited.

Most internationally published soil and water screening levels for PFOS and PFOA have excluded exposure through the dermal pathway; this reflects the limited information that is available on dermal absorption. At neutral water conditions dermal absorption will not contribute significantly to risk at lower concentrations, and therefore dermal exposure has not been included in the HSL derivation calculation.

4.5.3 Inhalation of dust

In an inhalation study, Kennedy et al. (1986) measured 108 mg/L of ammonium perfluorooctanoate (APFO) in the blood of male rats exposed ten times to 84 mg/m3 APFO, indicating that absorption through the lungs readily occurs for PFAS.

While dust inhalation exposure in the workplace of manufacturing plants can be very high, exposure to dust via soil/dust fugitive emissions are orders of magnitude lower than through the oral pathway. This is shown in the development of HILs presented in


the NEPM. For this reason, the dust inhalation component has not been included in the derivation of the PFOS and PFOA HSLs.

4.5.4 Intakes from other sources

A national survey in Australia showed PFOS and PFOA detected in drinking water to range from LOR–15.6 ng/L and LOR–9.6 ng/L with corresponding uptake rates estimated to be LOR–11 ng PFOS per day and LOR–13 ng PFOA per day (Thompson et al. 2011).

In 2006–2007, 2420 human blood serum samples were collected in southeast Queensland, and were pooled according to age. Across all pools (1 pool represents 100 samples), PFOS and PFOA were detected with mean concentrations of 15.2 ng/mL and 6.4 ng/mL (Toms et al. 2009). More recently, pooled human sera from 2008–2009 (n=24 pools) and 2010–2011 (n=24 pools) in Australia were obtained from de-identified surplus pathology samples (Toms et al. 2014).

Serum PFOS and PFOA concentrations appear to have decreased from 2002 to 2011 in the Australian population; this is consistent with the declining global use of PFOS and PFOA since around 2002. Biomonitoring data (2010–2011) for PFOS and PFOA report mean concentrations for PFOS and PFOA of 10.2 and 4.5 ng/mL, respectively. Assuming first-order toxicokinetics with half-lives of 5.5 years and 3.2 years, these concentrations correspond to estimated total daily intake rates for PFOS and PFOA of 0.89 ng/kg per day and 0.50 ng/kg per day, and hazard quotients of 0.006 and 0.0003 for PFOS and PFOA based on EFSA (2008) TDIs.

These background exposure rates are less than 1% of the adopted TDIs. Inclusion of these background values will not result in a change to the derived HSLs to two significant figures, and are therefore considered low and insignificant. On this basis, for deriving HSLs from contamination it has been assumed that the background intake contributes a negligible amount of the hazard quotient (<0.01), and the hazard quotient arising from the contribution from soil or water contamination should not exceed 1.0 for PFOS (+PFHxS) and PFOA.

A detailed assessment of background exposure in Australia is presented in appendix A.

4.6 Exposure estimation – soil

4.6.1 Exposure parameters

When calculating the risk of exposure to contaminated soil through direct contact, a chronic daily intake (CDI) is estimated by using exposure parameters that reflect the potential exposure pathways. The CDI may be derived using the equations published in appendix B, Schedule B7 of the NEPM (NEPC 2013). The equation for oral ingestion used to derive the soil HSL is as follows:

BW AT 365CFED EF IGRs F A Cs CDI o

×××××××

=

Where:

• CDI = chronic daily intake (mg/kg/d) • Cs = chemical concentration in soil (mg/kg) • EF = exposure frequency (d/yr)


• 365 = number of days in a year • ED = exposure duration (yr) • IGRs = soil incidental ingestion rate (mg/d) • AoF = oral absorption factor (-) • CF = conversion factor (10-6 kg/mg oral) • BW = body weight (kg) • AT = averaging time (yr), average life expectancy for non-threshold carcinogens,

ED for threshold chemicals (non-carcinogens)

The exposure parameter values assumed and their basis are summarised in table 3. Note:

• The exposure parameters adopted are consistent with those adopted in the derivation of NEPM HILs.

• Oral bioavailability is assumed to be 100%. • The contribution of inhalation and dermal absorption is considered negligible in

comparison to that occurring through the oral exposure pathway. • Adult body weight is assumed to be 70 kg, and for children 0-6 years 15 kg as per

the NEPM. • The soil ingestion rate of 100 mg/d by young children, recommended by enHealth,

is based on typical hand-to-mouth activities (enHealth 2012a). These do not consider pica behaviour.

• The contribution to exposure by PFOS and PFOA soil contamination should not exceed 1.0 (the target risk quotient is 1.0), as the background exposure contribution is less than 1% of the adopted TDI (refer to section 4.6.4)

For deriving HSLs for soil ingestion the above equation (CDI) is combined with the equation in risk characterisation (risk quotient, RQ) (section 3.2) and is as follows:

𝐻𝐻𝐻𝐻𝐻𝐻 =𝑅𝑅𝑅𝑅𝑡𝑡𝑏𝑏𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡 × 𝑇𝑇𝐶𝐶𝐶𝐶 × 𝐶𝐶𝐶𝐶 × 365 × 𝐴𝐴𝑇𝑇 × 𝐵𝐵𝐵𝐵

𝐴𝐴𝑜𝑜𝐹𝐹 × 𝐶𝐶𝐼𝐼𝑅𝑅𝐼𝐼 × 𝐸𝐸𝐹𝐹 × 𝐸𝐸𝐶𝐶 × 𝐶𝐶𝐹𝐹

4.6.2 Summary of derived soil HSLs

The calculations for deriving soil HSLs using the EFSA (2008) TDIs recommended by enHealth are presented in Appendix B. The resulting HSLs, rounded in accordance with the NEPM, are summarised in table 3. For HSL-A (low density residential), HSL-B (high density residential) and HSL-C (open space, recreation) the HSLs are based on the child receptor being the limiting criterion. Soil ingestion rates for the HSL-C scenario are based on half of the HSL-A soil/dust ingestion rate . If this assumption is considered overly conservative for a site, a site-specific assessment could incorporate the current and potential future risks ensuring that jurisdictional requirements are met.

When assessing both PFOS and PFOA the combined contribution of risk needs to be assessed. This is discussed in section 4.10.5. In addition, exposure levels should be considered in site-specific investigations where multiple exposure routes occur.

It should be noted that the HSL-A does not include an allowance for consumption of home grown garden produce that might take up contaminants through growing in contaminated soil, whereas NEPM HIL-A values generally include an allowance for 10% consumption of home-grown produce. In the case of PFAS there is insufficient information to provide a reliable estimate of plant uptake. As PFAS molecules are large, transfer through root membranes is likely to be limited and this may not represent a significant addition to the overall exposure, although soil adhering to plant material may provide a route for exposure.


If this exposure route is considered to be significant (e.g. in rural and agricultural settings) further consideration may be required, with direct sampling of produce if practical. Accumulation and distribution of PFASs depend on plant species. Edible plants such as vegetative portions of carrots, cucumbers, maize, wheat, potato and oats have also been found to accumulate PFOS and PFOA (Blaine 2014, Lechner & Knapp 2011, Stahl et al. 2008). In addition, long-chain PFASs are mainly distributed inplant roots, while short-chain PFASs are mainly found in leaves and fruits (Felizeter et al. 2014).

The HSL A does not allow for uptake via consumption of home grown eggs, poultry or other stock animals. If this exposure route is considered significant, it will require further consideration.

The NEPM advises that the screening levels are conservative and derived to be protective of the majority of the general population, but not necessarily infer protection of the most sensitive members of the community. With regards to PFAS exposure, where a person is exposed currently or historically to a separate significant source of PFAS (for example firefighters), then it is possible that the HSLs may not be protective for this individual. Such situations should be addressed for sites on a case by case basis.


Table 3 Exposure parameters – soil exposure

Parameter Symbol Unit Residential low density (HSL-A)

Residential high density (HSL-B)

Recreation/open space low density (HSL-C)

Commercial/ industrial (HSL-D)

Body weight BW kg 15 70 15 70 15 70 70

Exposure duration ED yrs 6 29 6 29 6 29 30

Exposure frequency EF days/yr 365 365 365 365 365 365 240

Soil/dust ingestion rate (1) IGRs mg/d 100 (2) 50 (2) 25 (3) 12.5 (3) 50 (4) 25 (4) 25 (5)

Averaging time AT yrs 6 29 6 29 6 29 30

Contribution to contamination CC - 1.0 (refer to section 4.6.4)

Notes: 1. Soil ingestion rates for children are based on a child aged 2–3 years where normal hand-to-mouth activity is assumed and does not account for pica behaviour; 2. Soil ingestion rates for the HSL A scenario include the ingestion of both outdoor soil, including soil adhering to home-grown produce, and indoor dust (derived from outdoor soil tracked indoors); 3. Soil ingestion rates for the HSL-B scenario are based on the assumption that a quarter of the HSL-A soil/dust ingestion occurs; 4. Soil ingestion rates for the HSL-C scenario are based on the assumption that half of the HSL-A soil/dust ingestion occurs, i.e. ingestion of outdoor soil only (no indoor dust).

Table 4 Summary of soil HSLs for PFOS (+PFHxS) and PFOA (mg/kg) Land use PFOS+PFHxS PFOA

Low density residential (HSL-A) 22 220 High-density residential (HSL-B) 90 900 Open space, recreation (HSL-C) 45 450 Commercial/industrial (HSL-D) 640 6400


The low density residential HSL-A values derived are compared with international levels in table 5. It can be seen that:

• The HSL-A values are similar to the US EPA 2009 levels; however, these are based on older toxicity information. The US EPA 2016 water guideline is based on a more stringent toxicological endpoint.

• The HSL-A values are less stringent (order of magnitude) than the soil guidelines recently published in Denmark (DEPA 2015). The primary difference between the derivations is that the Danish EPA assumed that contamination exposure should not exceed 10% of the TDI, while the derivation of the HSLs has assumed that the contribution from other (background) sources is minor and does not require an allowance, following the practice in the NEPM. The Danish guidelines are also formulated in terms of the sum of PFOS and PFOA and derivatives. Refer to section 4.9.5 for discussion on assessment of cumulative risks using the HSLs.

Table 5 Example international soil (residential) guidelines for PFOS and PFOA (mg/kg)

Location PFOS PFOA Reference USA 6 16 US EPA 2009b

Denmark 0.4 (sum of PFOS and derivatives)* 1.3* DEPA 2015 Derived HSL-A 22 (includes PFHxS) 220 -

* Denmark require assessment of accumulative risks of PFOS, PFOA and derivatives

4.7 Exposure estimation – water (fish consumption)

As indicated in Section 4.4, consumption of seafood can be a significant exposure pathway for PFOS and PFOA. Bioaccumulation factors over 1000 for PFOS in fish can result in a water criterion that is significantly lower than the drinking water criterion, and RIVM (2010), for example, has published a criterion of 0.65 ng/L for water to protect humans exposed to PFOS via consumption of fish.

In order to develop screening levels for water and sediment that will protect human consumption of seafood, consideration has been given to:

• The maximum permissible concentration (MPC) of PFAS in seafood, based on protection of human health. This has been estimated based on the amount of fish human receptors consume, and the toxicity endpoint.

• The HSLs for water and sediment, based on bioaccumulation factors for water and sediment to fish that result in PFAS concentrations in fish equal to the calculated MPC.

4.7.1 Fish consumption rates

The Australian exposure factors guide (enHealth 2012b) publishes a number of Australian and international statistics on consumption rates of different food groups for males and females. In the seafood category, values are published for fin fish (excluding canned), which is the key consideration in developing HSLs protective of fishing waters. Based on information reported by the Australian Bureau of Statistics (ABS 1999) the highest average consumption rate was for the age group 45–64, and has been used to represent the adult age group. Data for the child age group 2–3 has also been presented. The results are summarised in table 6.


Table 6 Average seafood consumption rates (g/day)

Age group 2–3 years (children) 45–64 (adults) Male

Total seafood 6.9 32.8 Fin fish only 0.3 8.5

Female Total seafood 6.5 20.0 Fin fish only 1.0 3.9

Source: table 4.4.1a and 4.4.1b from Australian exposure factors guide (EnHealth 2012b)

Since the publication of the enHealth exposure handbook, the Australian Bureau of Statistics released further average seafood consumption rates (g/day) based on the Australian Health Survey: 2011–12 (ABS 2014). For people, the average consumption of finfish only was 3.6 g/day in children (2–3 years old) and 10.3 g/day in in adults (51–70 years old). Total seafood consumption (including fish and seafood products and dishes) was 8.5 g/day in children (2–3 years old) and 33.7 g/day in in adults (51–70 years old) (ABS 2014).

The enHealth guide also publishes data on recreational fishing. The data are sourced from the Exposure factors handbook with mean and 95th percentile marine recreational fish consumption data are provided (US EPA 2009a). The data are based on per capita values for recreational fishing population in the USA. These values are presented in table 7.

Table 7 US estimates for recreational marine fish intake (g/day)

Age group Mean 95th percentile 3–6 years 2.5 8.2

>18 5.6 18 Source: table 4.4.6 from Australian exposure factors guide (EnHealth 2012b)

The 95th percentile for recreational fishing (table 7) is more conservative than average fin fish consumption rate (e.g. table 6). For the purpose of deriving HSLs protective of fish consumption, the 95th percentile for recreational fishing listed in table 7 has been used.

4.7.2 Maximum permissible concentration in fish

The maximum permissible concentration in fish (MPCfish) is the concentration of PFAS in fish that corresponds to consumption of the maximum allowable human intake levels of PFAS.

The derivation of MPCfish is similar to the derivation of soil HSLs, and is calculated as follows:

𝑀𝑀𝑀𝑀𝐶𝐶𝑓𝑓𝑓𝑓𝑓𝑓ℎ =𝑅𝑅𝑅𝑅𝑡𝑡𝑏𝑏𝑡𝑡𝑡𝑡𝑡𝑡𝑡𝑡 × 𝑇𝑇𝐶𝐶𝐶𝐶 × 𝐶𝐶𝐶𝐶 × 365 × 𝐴𝐴𝑇𝑇 × 𝐵𝐵𝐵𝐵

𝐴𝐴𝑜𝑜𝐹𝐹 × 𝐶𝐶𝐼𝐼𝑅𝑅𝐼𝐼 × 𝐸𝐸𝐹𝐹 × 𝐸𝐸𝐶𝐶 × 𝐶𝐶𝐹𝐹

Where IGRf is the fish ingestion rate (refer to section 4.8.1). All other parameters are the same as presented for soil HSL-A (low density residential) (refer to section 4.7).

As discussed in section 4.6.4, background exposure rates for PFAS in Australia are less than 1% of the adopted TDIs. Inclusion of these background values will not result in a change in the derived HSLs to two significant figures. On this basis, for the


purpose of deriving HSLs from contamination, background intake contributes a negligible amount of the hazard quotient (<0.01), and the requirement is that the hazard quotient arising from the contribution from soil or water contamination should not exceed 1.0 for PFOS (+PFHxS) and PFOA. This differs from the more conservative approach adopted in the drinking water guidelines (NHMRC 2011), in which an allowance of 90% for other (non-drinking water) exposure is made. In the case of the drinking water guidelines, large populations are potentially involved whereas in the case of fish consumption, exposure to fish contaminated by PFAS is likely to be a localised issue affecting only a small number of people, for which the NEPM methodology can be considered more appropriate.

For PFOS (+PFHxS) and PFOA, using the EFSA (2008) TDI values of 0.15 µg/kg/d and 1.5 µg/kg/d respectively, the maximum permissible concentration in fish (MPCfish) for PFOS (+PFHxS) is 270 µg/kg and for PFOA is 2700 µg/kg. These values are based on the child receptor; the adult receptor MPCs are about double these values.

The cumulative risk of exposure to both PFOS and PFOA requires assessment, as well as when exposure to multiple media occurs. To assess cumulative risk refer to section 4.10.5.

4.7.3 RIVM approach to assessing bioaccumulation

RIVM (2010) has calculated a biomagnification factor (BMF) to derive risks limits for secondary poisoning and human fish consumption using laboratory and field studies. The BMF has been introduced to account for accumulation in the food chain – from lower organisms (such as benthic organisms) to higher organisms (such as fish and birds). A biomagnification factor (BMF) takes into account accumulation of substances from the aqueous phase.

Assuming that the bioaccumulation factor (BAF) provides a measure of the bioaccumulation (which includes bioconcentration and biomagnification) in lower organisms relative to the exposure concentration in water, the resulting concentration in higher organisms such as fish, for example, is BCF x BMF. Hence, the BAF takes into account exposure via water phase and food intake. The HSL for water can be estimated using the following formula which takes into account the maximum permissible concentration (MPC) in fish:

HSL hh food, water (mg/L) = MPC hh food/(BCF*BMF) (1)

In terms of the maximum permissible concentration in fish, RIVM (2010) assumed that the quantity of contaminated fish consumed is 115 g/day, the intake from fish must not exceed 10% of the TDI (as defined in section 2.5), and the RIVM TDI is 0.15 µg/kgbw/d. Applying these assumptions, RIVM (2010) calculated a MPChhfood of 9.1 µg/kg.

Using equation (1) to derive a value for HSLhh food, water (mg/L), RIVM (2010) adopted a BCF value of 2,800 L/kg, and a BMF value of 5. The BCF value was derived from a laboratory study using the whole body of bluegill sunfish exposed to PFOS at 86 µg/L. As shown in RIVM (2010), the water concentration has a large effect on the BCF; carp exposed to 20 µg/L showed a BCF of 818 L/kg, and a BCF of 2,180 L/kg when exposed to PFOS at 2 µg/L. Assuming that we are most interested in low water concentrations that could be limiting in terms of bioaccumulation in fish, the application of a BCF value in the order of the latter value as RIVM have done would be appropriate.


Following the RIVM (2010) approach, applying a BCF value of 2800 L/kg and a BMF factor of 5 in equation (1) results in an HSLhh food, water value of 0.65 ng/L.

This value is based on laboratory studies, and assumes that the laboratory-measured bioaccumulation applies to organisms lower in the food chain, and subsequent biomagnification at higher trophic levels in the food chain. Field-derived BAF values for fish could provide a more direct indication of the concentrations of PFOS that result from the concentrations of PFOS measured in water or sediment, although RIVM (2010) suggest that there is considerable uncertainty in field derived BAF values. The revised formula in terms of BAF would be:

HSLhh food, water (mg/L) = MPChh food/BAF (2)

The BAFs for freshwater fish vary from 2,500 L/kg to 95,000 L/kg and the BAF for marine fish range from 1,580 to 9,780 L/kg (RIVM 2010). For marine species, the geometric mean of BAF values calculated from field data using fish exposed to environmental concentrations as shown in table A1.2 in RIVM (2010) was 4,500 L/kg. The geometric mean of the BAF values for freshwater species from the RIVM (2010) data was 12,900 L/kg.

Applying these BAF values in equation (2) results in the following values for MPChh food, water:

• Marine = 2.0 ng/L, and • Freshwater = 0.71 ng/L.

These estimates support the RIVM estimate of 0.65 ng/L for freshwater, but suggest that a somewhat higher value might be adopted for marine water.

4.7.4 Application of the RIVM approach to assessing bioaccumulation in Australia

The concept of determining the concentration of a contaminant that will result in edible fish species is relevant to Australia, although generally the approach of predicting such a concentration is not adopted, and instead the practice is to rely on direct measurement of potentially affected fish species. The question arises as to whether the approach and value that derives is appropriate for use in Australia.

With respect to international regulatory standing, the RIVM approach to accounting for bioaccumulation in biota and the value of 0.65 ng/L has been adopted in the European Union as an environmental quality standard for PFOS in freshwater, based on the potential for secondary poisoning in humans due to fish consumption. The date set for EU-wide compliance with the standard is 2027, with member states required to submit to the Commission a supplementary monitoring programme and a preliminary programme of measures to achieve compliance by December 2018. The value has also been referenced for application in the USA (e.g. US EPA 2012). As such, the approach and value has some regulatory standing.

In Europe, there has been considerable concern by many persons in the EU that the value set is very low and may not be able to be complied with (CONCAWE 2016). Concerns noted include:

• Background PFOS concentrations in many European surface water bodies are higher than the standard, which presents major challenges for compliance.


• The analytical methods currently used by commercial laboratories yield quantification limits above or close to the standard.

• The standard is nearly 3 orders of magnitude lower than the (provisional) drinking water standards (generally between 0.1 and 0.5 µg/l for PFOS.

It was noted that the standard is dependent on the assumptions that have been adopted in its preparation, and alternative assumptions can result in a much higher value (CONCAWE 2016). For example, if the following alternative assumptions are made: TDI (0.3 versus 0.15 µg/kg day), contribution of fish to dietary uptake (73% versus 10%), average fish consumption rate (0.028 kg/day versus 0.115 kg/day), bioconcentration and biomagnification factors (1,124 versus 2,800 and 2 versus 5), the standard could have been derived 375 times higher (0.24 µg/L). These factors have been considered in this report, and assumptions have been made that are considered to be appropriate for Australia, resulting in a value that is somewhat higher than the EU standard.

A further area of uncertainty is the bioaccumulation factors that should be adopted. There is insufficient data to support the setting of bioaccumulation factors relevant to Australian biota, and in the absence of such data, it has been considered appropriate to adopt the RIVM factors.

In summary, because of these various factors it is considered that there is uncertainty in the HSLs to protect human consumers of fish, and it is recommended that, where possible, direct sampling of potentially affected biota be undertaken where there is concern, to determine the level of contamination that occurs.

In addition, it is recommended that there be a watching brief to determine whether alternative approaches are adopted internationally, and their relevance to Australia.

4.7.5 Water and sediment HSLs for seafood consumption

A detailed analysis and discussion of the relationships between PFAS concentrations in water, sediment and aquatic organisms is presented in part 3 of the guidance. These relationships allow estimates to be made of the concentration of PFAS that will result in fish and, based on the maximum permissible concentration in fish (MPCfish – refer to section 4.8.2), HSLs for waters and sediment protective of seafood consumption may be derived as discussed in section 4.8.3 using the RIVM method.

In the case of sediments, a sediment-biota bioaccumulation factor (BSAF) is similarly defined. BSAF incorporates the relationship between the concentrations in sediment and organisms i.e. how sediment samples reflect the actual organism’s recent exposure (Burkhard et al. 2003)

Part 3 provides summaries of various studies which provide information on these factors for both fresh and marine waters. For PFOS the BAFs for freshwater fish vary from 2,500 L/kg to 95,000 L/kg and the BAF for marine fish range from 1,580 to 9,780 L/kg (RIVM 2010). The geometric means of BAF values calculated from field data using fish exposed to environmental concentrations as shown in table A1.2 in RIVM (2010) were:

• BAFmarine species = 4,500 L/kg • BAFfreshwater species = 12,900 L/kg


A safe concentration for sediment based on a safe concentration in fish for human consumption has not been calculated by RIVM. The geometric mean of the BSAF values in Table 5 in part 3 is 12.7.

Note that a site-specific BSAF would be ideal as this would incorporate all processes and conditions influencing bioaccumulation at that site. A site-specific BSAF is not always available, especially at the initial stages of risk assessment, and therefore BSAFs from different ecosystems may be used assuming that there are similarities in the underlying conditions and parameters affecting bioaccumulation (Burkhard et al. 2003). Notably, Burkhard et al. (2003) outlines the conditions and parameters affecting BSAF i.e. distribution of the chemical between sediment and water column; relationship of food web to water and sediment; length of food web (or trophic level); bioavailability of the chemical due to amounts and types of organic carbon and metabolic transformation rates of the chemical within the food web. The latter varies little across ecosystems.

HSLs for water and sediment protective of human consumption of fish is calculated using the following equations:

HSLwater, fish consumption (mg/L) = MPCfish/BAF, or

HSL sediment, fish consumption (mg/kg) = MPCfish/BSAF Due to the uncertainties associated with bioaccumulation, where possible it is preferred that biota sampling of fish be undertaken and compared to the MPCfish.

Limited information is available on bioaccumulation of PFOA in seafood. One report from the USA published a bioaccumulation factor of 2,680 L/kg in oligochaete (worms) (Laiser et al. 2011). Compared to the BAFfreshwater species of 12,900 L/kg (RIVM 2010), this suggests that PFOA is bioaccumulative in seafood, but to a lesser degree than PFOS. This is consistent with observations of bioaccumulation in humans based on the estimated half-lives (refer section 2.3). Assessment of the PFAS sediment and pore water concentrations, together with body burdens in benthic biota will provide important information in assessing the impacts on aquatic biota and the potential for human health impacts. Some further options for assessing sediment ecosystems are provided in part 3, section 5.4. Such considerations will also be relevant in human health risk assessments for edible organisms.

4.7.6 Summary of derived surface water HSLs for human consumption of seafood

The calculations of HSLs for water and sediment protective of human consumption of seafood are presented in appendix B. A summary of the derived HSLs are presented in table 8; these have been rounded in accordance with the NEPM conventions. Note the HSLs are based on consumption of fish by children, the limiting human receptor. Due to the limited understanding of PFAS bioaccumulation in environments, where possible it is preferred that biota sampling of fish be undertaken and compared to the MPCfish.


Table 8 Summary of derived water and sediment HSLs protection of human health consuming seafood

Derived health screening levels PFOS (+PFHxS) PFOA Maximum permissible concentration in fish (MPCfish) 270 µg/kg 2700 µg/kg

HSLfresh water, fish consumption 21 ng/L 210 ng/L HSLmarine waters, fish consumption 61 ng/L 610 ng/L HSLsediments, fish consumption 22 µg/kg 220 µg/kg

The derived HSL for PFOS (+PFHxS) for consumption of fish from freshwater (21 ng/L) is higher than the RIVM (2010) water guideline of 0.65 ng/L. The primary differences between the derivations are that RIVM used 10% of the TDI to allocate to contamination exposure while the HSL is based on 100% of the EFSA (2008) TDIs. The cumulative risk of exposure to both PFOS and PFOA is required to be assessed, as well as when exposure to multiple media occurs. To assess cumulative risk refer to section 4.8.5.

4.8 Limitations, uncertainty and application of HSLs

The following section outlines the key considerations to the application of the PFOS and PFOA HSLs, and limitations regarding their use. For a more detailed discussion refer to part 4 of the guidance.

4.8.1 Conceptual site model

Prior to applying the HSLs for PFOS and PFOA it is important to develop an appropriate CSM in accordance with Schedule B2 (Site Characterisation) of the NEPM. It is particularly important to understand the transport mechanisms (leaching, runoff, groundwater migration and surface water discharge, bioaccumulation in food chain), and to follow the source-pathway-receptor linkages.

With regard to PFOS and PFOA, because of the potential for persistence and bioaccumulation, the CSM will need to consider scenarios and pathways that account for these issues, and which may not be normally considered for other contaminants. Note that while the HSLs derived in this document provide criteria for evaluation of some of these issues, the HSLs do not extend to all pathways and some will require site-specific assessment, such as biota sampling. Pathways that may need to be examined include the following:

• Leaching from soil to groundwater • Groundwater discharge into surface water bodies • Adsorption to sediments • Bioaccumulation in seafood • Discharge to the sewerage system, with potential for accumulation in biosolids and

discharge in the treated effluent • Irrigation with uptake by plants, and possible accumulation in soil • Ingestion of contaminated plant material and soil by livestock during grazing • Uptake by garden produce and ingestion by persons or animals • Consumption by persons of meat and animal products (e.g. milk and eggs), and • Consumption of breast milk by infants.


Further information regarding development of a CSM for PFOS and PFOA contaminated sites is provided in part 1 and part 5 of the guidance.

4.8.2 Soil HSLs for PFOS and PFOA

The HSLs for PFOS (+ PFHxS) and PFOA for soils have been derived in a similar manner to the HILs presented in the NEPM, and should be similarly applied. The HSLs are primarily based on direct contact exposure to contaminated soil. Application of a statistical approach to estimate the exposure that will occur can be appropriate.

As specified by enHealth (2016) the derived HSLs for PFOS apply to the sum of PFOS and PFHxS.

Note that the soil HSLs for PFOS and PFOA do not consider the following:

Exposure through consumption of home-grown garden produce. The uptake of PFAS by garden produce may be significant, but is difficult to generalise for the purpose of deriving screening levels. Therefore, this pathway has not been included in the derivation of the low density residential (HSL-A) scenario and differs from the default assumptions for this scenario presented in the NEPM. If uptake by garden produce (residential and agricultural scenarios) is required to be assessed, then a site-specific assessment should be undertaken which considers this exposure route. Direct sampling of plant species may be necessary to determine the uptake that occurs.

Exposure through consumption of home-grown poultry, eggs or other animal products. The uptake of PFAS by poultry and animals kept to supply meat, eggs or milk for human consumption may be significant, but it is difficult to generalise for the purpose of deriving screening levels. Therefore, this pathway has not been included in the derivation of the low density residential (HSL-A) scenario. If uptake of poultry, eggs, meat, milk or other animals products (residential and agricultural scenarios) is required to be assessed, then a site-specific assessment should be undertaken which considers this exposure route. Direct sampling may be required.

Protection of groundwater/surface water quality. Due to the persistence and solubility of PFOS and PFOA, leaching into groundwater systems is often observed in areas where PFAS contamination of soil has occurred. Where groundwater contamination is identified, the significance of contamination can be assessed by comparison with the relevant water screening level (human health and/or ecological). If it is proposed that PFOS and PFOA contaminated soil will remain onsite and assessment of the potential for leaching is required, this may be best indicated by direct measurement of the concentrations observed in groundwater (rather than attempting to predict the groundwater concentrations that will result from soil contamination). Additionally, leaching potential of PFAS from soil samples may be determined using the Australian Standard Leaching Procedure (ASLP) (AS 1997). This test measures how much of a chemical can move from soil into water using conditions similar to rain events. Note that such leaching tests generally have a very limited application as they only provide information about the leaching potential of solid materials under the specific chemical conditions (WA DER 2015). Other types of leaching tests are available which are designed to be used to evaluate leaching in landfills, under more extreme conditions, and are less relevant for the initial screening of these sites. The trigger points in leachate as measured in an ASLP test should consider the dilution that will occur as the leachate moves from the soil into the underlying groundwater; this will depend on the relative rates of rainfall and


groundwater movement, absorption in the soil, and the depth of groundwater into which the leachate mixes. Note that US EPA recommends a default factor of 20.

Stormwater runoff quality also needs to be considered (based on jurisdictional requirements).

Ecological receptors. For the assessment of the potential impact on terrestrial ecological receptors, refer to the derived ESLs presented in part 3 of the guidance.

4.8.3 Surface water HSLs for PFOS and PFOA for consumption of seafood

HSLs have been derived for surface waters (fresh and marine waters) and sediment for the protection of human consumption of seafood. Because of the potential for bioaccumulation and biomagnification, the values that have been derived for this situation are significantly more stringent than the values for direct toxicological effects on the ecosystems.

As specified by enHealth (2016) the derived HSLs for PFOS apply to the sum of PFOS and PFHxS.

The following considerations are required when applying these HSLs:

• The HSLs for consumption of edible fish and crustaceans should only be considered for water bodies where fishing or the harvesting of edible crustaceans, molluscs, etc. is known to occur. The HSLs need not apply to small drains, ponds and creeks where such activities do not take place, unless the water is known to discharge into an area downstream where such activities take place.

• The HSLs for protection of edible fish and crustaceans should not be applied directly to groundwater. Instead, the HSLs should be applied to the surface water into which the groundwater discharges. The HSLs for sediments are applied to the sediments where fish and crustaceans are in contact with the sediments or benthic organisms that are in contact with the sediments.

• In assessing surface waters, the HSLs should be compared with the directly measured concentrations of PFAS in the surface water body, or predictions of concentration that could arise in the surface water body through consideration of mass loading (i.e. flux) of the contaminant discharge and water body flow rate, and other factors that can influence dilution and dispersion, such as tidal exchange. Because concentrations can vary markedly over short time periods (such as with a tidal cycle), average concentrations may be more relevant than individual short term peak concentrations.

In terms of uncertainty, the main areas of uncertainty relating to the HSLs for fresh and marine water for protection consumption of edible fish and their application relate to the level of bioaccumulation that will result in edible fish and crustaceans, and the dependence of this level on the extent to which these species will be exposed to the contamination. This level of uncertainty may be in the order of several orders of magnitude.

Because of the high level of uncertainty, the water HSLs relating to consumption of fish and other species derived in this report should be used for screening purposes as an indication of the possibility that the fish and other species may be contaminated, and whether further investigation should be undertaken. Where it is considered that further investigation is required, consideration should be given to direct sampling of PFAS


concentration in fish and other species in question and compared to the MPCfish presented in table 7.

4.8.4 Interim drinking water quality and recreational water quality guidelines

In June 2016 enHealth issued a statement interim national guidance on human health reference values for per- and poly-fluoroalkyl substances for use in site investigations in Australia. In this statement enHealth publish an interim drinking water quality guideline, derived using the methodology presented by NHMRC. The statement contains the following:

• enHealth recommends that the EFSA TDI values are used by jurisdictions in site investigations in Australia to determine interim drinking water and recreational water guideline values using the methodology described in Chapter 6.3.3 of the National Health and Medical Research Council (NHMRC) Australian Drinking Water Guidelines. enHealth considers that calculating an interim drinking water guideline value in this way is more appropriate for the Australian context than using the 2016 US EPA Drinking Water Health Advisories.

• The interim drinking water guideline values are not intended to be a guide for drinking water utility providers across Australia, but rather for use to confirm the quality of drinking water supplies potentially affected by specific instances of site contamination.

• enHealth has also considered the available information on perfluorohexane sulfonate (PFHxS) and agreed that the EFSA TDI for PFOS also be applied to PFHxS exposures. In practice this means PFOS and PFHxS exposures should be summed and the total compared to the TDI for PFOS.

In summary:

• The interim drinking water quality guideline values for PFOS (+PFHxS) and PFOA are 0.5 µg/L and 5 µg/L respectively.

• These drinking water guideline values are based 10% of the EFSA TDI, allowing 90% of the TDI for protection from exposure to other sources.

• The interim Recreational Water Quality Guideline for PFOS (+PFHxS) and PFOA is 5 µg/L and 50 µg/L respectively, based on the 10-fold increase in drinking water quality guideline as specified by NHMRC 2008 Recreational Waters Quality Guidelines.

4.8.5 Multiple pathway exposure and cumulative risk

Schedule B(4) of the NEPM (2013) specifies that when assessing the risk of combined exposure, all relevant exposure routes must be included. This is important in the context of PFAS investigations as exposure to PFAS contamination can occur through soil, groundwater and the food chain, i.e. through more than one mechanism, due to its persistence, mobility and bioaccumulation properties.

In addition, the recent toxicological information suggests an additive effect between PFAS and their derivatives, and therefore assessment should be made for cumulative risks of the chemicals. A specific example of this is enHealth’s 2016 statement that the EFSA (2008) TDI for PFOS should apply to PFOS and PFHxS, and the derived Drinking Water Quality Guideline should apply to the sum of PFOS and PFHxS.


The HSLs presented in preceding sections were derived independent of other exposure routes. For example, the estimate of exposure occurring through soil ingestion (direct contact with soil) does not consider whether there might also be exposure through groundwater used for domestic (including potable) use or to water stock.

HSLs for some exposures are calculated using a 100% relative source contribution. If more than one of these sources is relevant, use of the screening values would underestimate risks, and in this case, cumulative assessment of all relevant exposures must be made. To evaluate the overall risk the cumulative fraction should be calculated as follows:

𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶𝐶 𝐹𝐹𝐹𝐹𝐶𝐶𝐹𝐹𝐶𝐶𝐶𝐶𝐹𝐹𝐹𝐹 = �𝐶𝐶𝑚𝑚𝑡𝑡𝑚𝑚𝑓𝑓𝑏𝑏 𝑥𝑥

𝐻𝐻𝐻𝐻𝐻𝐻𝑚𝑚𝑡𝑡𝑚𝑚𝑓𝑓𝑏𝑏 𝑥𝑥

That is, the PFAS concentration in soil, groundwater and surface water is divided by the relevant HSL for the relevant media and use. A value above 1 for any one of the media indicates an exceedance of the HSL. The sum of these fractions is the cumulative fraction and a result above 1 indicates such an exceedance when all routes are considered. Consequently, further investigation may be required. This calculation can also help with the investigation and management of risk as the pathways which pose the more significant risk can be identified.


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APPENDIX A. Human health toxicity review

CRC for Contamination Assessment and Remediation of the Environment

Expert health review: health criteria for PFOS/PFOA

Zhaomin Dong, Mezbaul Bahar & Ravi Naidu

March 2016, revised August 2016

CRC CARE Technical Report no. 38b 46

Appendix A

Executive summary

Perfluorooctanesulfonate (PFOS) and perfluorooctanoic acid (PFOA) have been

identified as contaminants of emerging concern in Australia. PFOS and PFOA both

comprise a perfluorinated carbon chain (or tail) that is both lipid and water repellent.

The generic term perfluoroalkyl substances (PFAS) is used to describe the mix of

straight- and branched-chain congeners that normally comprise the human exposure

matrix. PFOS and PFOS-related substances have been used globally in many

applications. These include aqueous film forming foams (AFFF), semiconductors,

hydraulic fluids, photolithography, as a key ingredient in Scotchgard brand fabric

treatment products, grease repellents for packaging, surface treatments for rugs and

carpets, paper and packaging, coatings and coating additives, industrial and household

cleaning products, pesticides and insecticides. Both PFOS and PFOA are known to be

persistent, bioaccumulative and potentially toxic, and proven to exist at a number of

sites, particularly where AFFF fire-fighting foams have been used, in concentrations of

potential concern.

Large differences in the half-lives of PFOS and PFOA have been observed between

animals and humans. For animals, reported half-lives range from 2 hours to 150 days.

The half-lives for humans have been reported to be in the range of 2–9 years. Based

on epidemiological studies, we estimate that the mean half-lives for PFOS and PFOA

are 5.5 years and 3.2 years, respectively.

Numerous adverse health outcomes have been observed in animal studies following

chronic low dose exposures including harm done to the liver, gastrointestinal tract,

thyroid hormone levels, as well as immunological, reproductive and developmental

problems. Animal studies at low levels of exposure (<1 mg/kg per day) suggest

moderate acute oral toxicity of PFOS/PFOA, with effects on the liver and

gastrointestinal tract. No data are available for assessing acute toxicity in humans.

Epidemiological studies conclude that there is a weak positive association between

serum PFOS/PFOA and increased serum cholesterol and uric acid levels. There are

also a few but inconsistent positive correlations with some health outcomes such as

increased serum liver enzymes, or immune system effects. Statistically significant

associations have been observed between PFOA exposure and kidney and testicular

cancers. Most results are from cross-sectional analysis and therefore the data are

insufficient to draw unambiguous conclusions about the effects of PFOS/PFOA in the

progress of any particular disease. For some effects, reverse causation cannot be ruled

out, where the increased incidence of the effect under consideration is associated with

a potentially greater accumulation of the PFAS.

Although using animal data to estimate effects on human health means introducing

interspecies uncertainties, animal models exclude human variability factors (such as

diet, drugs, infections, radiation, endogenous processes), they can provide a dose-

response study to derive a tolerable daily intake (TDI). This report was originally

designed to recommend TDI values for PFOS and PFOA. In June 2016, enHealth, the

peak environmental health body in Australia, selected the European Food Standards

Authority (EFSA) health reference values (HRVs) for PFOS (and PFHxS combined)

and PFOA for application as an interim measure in Australia (enHealth 2016). In

September 2016, the procedural review of health reference values established by


Appendix A

enHealth for PFAS noted that adoption of EFSA (2008) HRVs for PFAS in drinking and

recreational water was appropriate and is protective of public health

(Bartholomaeus 2016).

This report adopts the enHealth recommended approach for TDIs. The aim of this

report is to mainly provide background information on PFOS and PFOA required to

determine background intakes in Australia.

A recent study shows that the mean background for Australian serum concentrations

for PFOS and PFOA are 10.2 ng/mL and 4.5 ng/mL, respectively. Several reports

based on national surveys indicate a decrease in serum PFOS and PFOA

concentrations from 2002 to 2011 in the Australian population. The total daily intakes

for PFOS and PFOA are estimated to be 0.89 ng/kg per day and 0.50 ng/kg per day,

respectively. The background exposure level is far below the TDI. In particular, the

ratios of intake to TDI for PFOS and PFOA are estimated to be 0.006 and 0.0003,

respectively, indicating that Australians are generally at a low risk of danger to their

health from typical PFOS/PFOA exposures.


Appendix A

Acknowledgements

The authors would like to acknowledge Prof Brian G. Priestly, Director of the Australian

Centre for Human Health Risk Assessment, Monash University, Australia for his peer

review of this report. The authors also wish to acknowledge Ms Pamela Wadman MS,

toxicologist with the Maine Department of Health and Human Services, USA, and Naji

Akladiss PE with the Maine Department of Environmental Protection, who carried out

the peer review of the earlier version of this report. The authors would also like to

acknowledge the support of Dr Joytishna Jit, Project Manager with CRC CARE.


Appendix A

Table of contents

Executive summary 46

1. Introduction 51

1.1 Project background 51

1.2 Approach 51

2. Properties of PFOS/PFOA 52

3. Sources and uses of PFOS/PFOA 53

4. Background exposure in Australia 54

5. Toxicity review 56

5.1 Toxicokinetics of PFOS/PFOA 56

5.1.1. Uptake 56

5.1.2. Distribution 56

5.1.3. Bioaccumulation 56

5.1.4. Metabolism 57

5.1.5. Elimination 57

5.1.6. Total daily intakes for PFOS and PFOA 59

5.2 Toxicology of PFOS/PFOA 60

5.2.1. Acute toxicity 60

5.2.1 Health effects of chronic exposure: animal data 61

5.2.2 Human data: Epidemiological studies 68

6. Tolerable daily intake and the hazard quotient 70

6.1 Tolerable daily intake 70

6.2 Hazard Quotient 72

7. Conclusions 73


Appendix A

Tables

Table 1. Haematologic biomarkers for PFC exposure from Australian

epidemiological studies

54

Table 2. BCF and BAF values for PFOS and PFOA 58

Table 3. Elimination half-lives of PFOS/PFOA for various species 58

Table 4. Elimination half-lives of PFOS/PFOA for humans 59

Table 5. Acute toxicity criteria for PFOS/PFOA in animals 61

Table 6. Adverse effects observed in chronic exposure: animal studies

(PFOS)

62

Table 7. Adverse effects observed in chronic exposure: animal studies

(PFOA)

63

Table 8. Epidemiological studies related to PFOS exposure 64

Table 9. Epidemiological studies related to PFOA exposure 65

Table 10. Comparison of TDI values from various organisations 71


Appendix A

1. Introduction

1.1 Project background

Perfluorooctanesulfonate (PFOS) and perfluorooctanoic acid (PFOA) have been

identified as contaminants of emerging concern in Australia. These contaminants

belong to the large group of chemicals referred to as polyfluorinated alkyl substances

(PFAS) or fluorosurfactants. The term PFAS describes both fully fluorinated

compounds which are resistant to degradation, as denoted by the ‘per’ prefix, and also

partially fluorinated compounds such as fluorotelomers, including 6:2FtS and 8:2FtS.

Perfluorinated compounds (PFC) are known to be persistent, bioaccumulative and

potentially toxic, and are present in concentrations of potential concern at a number of

sites. This is particularly the case where aqueous film forming foams (AFFF) have been

used for fire-fighting. Industry and public awareness of the presence of PFOS and to a

lesser extent PFOA is growing rapidly, with regular reporting of PFOS contamination in

the news media. Concerns have been raised regarding the risk PFC contamination

may pose to the health of persons who may be exposed.

In general, there is limited and incomplete information surrounding the occurrence, fate

and toxicity of PFCs in Australia, and Australian criteria for protection of human health

have not been established. This makes it difficult to determine the potential risk that

PFC contamination poses.

More information about the toxicity of PFOS and PFOA has become available in recent

years, with many jurisdictions now revising or formulating health-based guidance.

Several reports recently published by international authoritative bodies provide detailed

discussions about human exposures, dose-response relationships, and effects on

human health (Danish EPA 2015, ATSDR 2015, EPA 2016a, 2016b). A comprehensive

literature review on the potential health effects of PFOS and PFOA has been prepared

by Professor Brian Priestly (2015).

A consistent, practicable approach to the risk-based assessment of PFC contamination

is required in Australia. The aim of this project is to facilitate this requirement by

developing human health toxicity criteria for PFOS and PFOA. The tolerable daily

intake has been selected by enHealth (2016) and in adopted in this report. This report

facilitates the development of additional information on background levels (e.g. daily

intakes levels and hazard quotient) which are also required for the calculation of health

toxicity criteria.

1.2 Approach

Although the toxicity database for PFCs is rapidly evolving, for this current effort we

have summarised what is known about the health effects of PFOS and PFOA. This

report provides information on the background exposure levels in Australian

populations.


Appendix A

2. Properties of PFOS/PFOA

PFOS and PFOA are synthetic chemicals that are chemically and biologically stable

and hence are persistent in the environment, resistant to biodegradation, atmospheric

photooxidation, direct photolysis, and hydrolysis (US EPA 2014). PFOS and PFOA can

also be formed as breakdown products from related substances, by microbial

degradation of precursor compounds, or as an end-product of biological metabolism

(e.g. rainbow trout transform precursor perfluorinated acids into PFOS and PFOA)

(CRC CARE 2014). Although the ultimate net contribution to environmental loadings of

PFOS from individual PFOS-related substances cannot be easily predicted, any

molecule containing the PFOS moiety could be a potential precursor to PFOS (UNEP

2006). It is possible that as PFCs are gradually phased out, precursor transformation

may contribute significantly to the PFOS contaminant load (CRC CARE 2014).

Both PFOS and PFOA consist of an eight-carbon chain that is completely substituted

with fluorine atoms, making it both lipid repellent (oleophobic) and water repellent

(hydrophobic), and a charged hydrophilic head (CRC-CARE 2014). PFOS is a

perfluoroalkyl sulfonate, with a sulfonic acid moiety forming the hydrophilic head, while

PFOA has a carboxylic acid moiety and is often called C8 (Seow 2013). The stability of

these PFCs is due to the strength of the carbon-fluorine bonds; each fluorine atom is

shielded by three electron pairs, and the carbon atoms are shielded by the fluorine

atoms (CRC CARE 2014).

PFOS and PFOA are moderately soluble and have long half-lives in water (41 years

and 92 years, respectively) (CRC CARE 2014). They are persistent in groundwater and

surface waters, although they have been found to partition from the groundwater

column into organic matter rich sediments and soil particles due to their propensity to

adsorb to organic carbon (US EPA 2014).

The vapour pressure for PFOS at 20 °C is 2.48 x 10-6 mmHg and for PFOA it is

0.017 mm Hg (US EPA 2014), and so vaporisation appears to be of little concern.

However, PFOS and PFOA can be transported long distances in the air because of

their high atmospheric half-lives (114 days and 90 days, respectively) (US EPA 2014).


Appendix A

3. Sources and uses of PFOS/PFOA

PFOA and PFOS-related substances have been used globally in many applications.

These include aqueous film forming foams (AFFF), chromium plating, semiconductors,

hydraulic fluids, photolithography, as a key ingredient in Scotchgard brand fabric

treatment products, grease repellents for packaging, surface treatments for rugs and

carpets, paper and packaging, coatings and coating additives, industrial and household

cleaning products, pesticides and insecticides (UNEP 2006, Seow 2013,

CRC CARE 2014).

PFOA is used in the production of other fluoropolymers, which are fire resistant, and

oil, stain, grease and water repellent (US EPA 2014). PFOA is contained in trace

amounts in Teflon™ (polytetrafluoroethylene) used on non-stick cookware, waterproof

and breathable membranes for clothing, and in the aerospace, automotive,

building/construction, chemical processing, electronics, semiconductors and textile

industries (Seow 2013, CRC CARE 2014). PFOA may also be produced under certain

conditions from the breakdown of some fluorotelomers such as 8:2 fluorotelomer

sulfonate (8:2FtS) which is also used in stain, grease and water resistant surface

treatment products, paints, coatings, cleaning products, fire-fighting foams or

engineering coatings (Seow 2013).

Disposing of PFCs into sewers may result in the accumulation of PFCs in the biosolids

of sewage treatment plants. Furthermore landfilling of products containing PFCs may

result in contamination of groundwater in the vicinity of landfills.

Although the use of PFOS and PFOA in Australia has been largely phased out, legacy

contamination issues remain due to the persistent nature of the chemicals. It is

expected that most of the sources of PFOS and PFOA contamination noted above

have existed in Australia, with the use of AFFF at fire-fighting training grounds being

probably the most significant, and one that requires particular attention.


Appendix A

4. Background exposure in Australia

Human exposure to PFOS/PFOA can occur via multiple pathways. Potential pathways

which may lead to widespread exposure include ingestion of food and water, use of

commercial products, or inhalation from long-range air transport (ATSDR 2015). A

national survey in Australia showed PFOS and PFOA detected in drinking water at

levels of 0–15.6 ng/L and 0–9.6 ng/L with corresponding uptake rates estimated to be

0–11 ng PFOS per day and 0–13 ng PFOA per day (Thompson, Eaglesham &

Mueller 2011). However, limited information for other pathways is available in Australia.

For example, in Europe, fish and fishery products appear to be one of the primary

sources of human exposure to PFOS (EFSA 2008). However, few data for

PFOS/PFOA content in fish are available for Australia. Taking the lack of exposure

information into consideration, daily intake for Australians is estimated by using

biomonitoring information.

In 2002–2003, 3802 serum samples were collected and analysed for PFOS and PFOA.

The samples were collected from five different age groups, both genders, and from

rural and urban regions in Australia (Kärrman et al. 2006). The mean concentrations for

PFOS and PFOA were 20.8 ng/mL and 7.6 ng/mL, respectively, and an increase in

PFOS concentration with increasing age in both regions and genders was observed

(Kärrman et al. 2006).

In 2006–2007, 2420 human blood serum samples were collected in southeast

Queensland, and were pooled according to age: across all pools, PFOS and PFOA

were detected with mean concentrations of 15.2 ng/mL and 6.4 ng/mL, respectively

(Toms et al. 2009). More recently, serum PFC concentrations were determined in

pooled human sera from 2008/09 (n=24 pools, representing 2400 individual samples)

and 2010/11 (n=24 pools, representing 2400 individual samples) obtained from de-

identified surplus pathology samples (Toms et al. 2014).

Table 1 summarises biomonitoring information from the studies referred to above for

PFOS and PFOA (Toms et al. 2014).

Table 1. Haematologic biomarkers for PFC exposure from Australian epidemiological studies

Data collected Chemical Range (ng/mL) Mean (St.d) (ng/mL) Median (ng/mL)

2002/03

PFOS

19.1–36.1 25.9 (4.7) 25.4

2006/07 5.0–28.5 15.2 (4.9) 14.8

2008/09 5.3–19.2 11.9 (4.6) 11

2010/11 4.4–17.4 10.2 (3.7) 9.4

2002/03

PFOA

7–14.5 10.2 (1.7) 10.6

2006/07 0.8–9.1 6.4 (1.5) 6.4

2008/09 2.8–7.3 5.2 (1) 5.1

2010/11 3.1–6.5 4.5 (0.8) 4.3

Global use of PFCs has been in decline since around 2002 and hence primary

exposure levels are expected to decrease (Toms et al. 2014). Calafat et al. (2007)

compared human data on the serum concentrations of some of the PFAS available

from the US National Health and Nutrition Examination Survey (NHANES) database.

The study compared data available from 2003-2004 with 1999-2000, and found PFOS


Appendix A

and PFOA concentrations had decreased. The decrease was attributed to the

discontinuation in 2002 of industrial production of PFOS and related compounds.

From table 1Table 1, it can be seen (based on available data) that serum PFOS and

PFOA concentrations appear to have decreased from 2002 to 2011 in the Australian

population. For the purpose of estimating daily intake figures, the most recent

biomonitoring data (2010/11) for PFOS and PFOA have been selected. The mean

concentrations for PFOS and PFOA in the Australian population are considered to be

10.2 and 4.5 ng/mL (mean values from table 1Table 1), respectively.


Appendix A

5. Toxicity review

This section provides a brief discussion of the available data and information. While the

information provided is not comprehensive, has been sourced from authoritative

literature reviews and from authoritative government publications. Firstly, the

toxicokinetics of PFOS and PFOA is reviewed to identify time and species dependent

processes related to these toxicants. Secondly, we have presented a brief summary of

some effects of theses toxicants, including acute and chronic toxicity.

5.1 Toxicokinetics of PFOS/PFOA

5.1.1. Uptake

Animal experiments demonstrate that the uptake of PFCs usually occurs via oral,

inhalation or dermal exposure. Oral uptake and inhalation of these compounds results

in quick assimilation (Stahl et al. 2011).

Oral uptake has been confirmed by Gibson and Johnson (1979), who reported that

95% of the radioactively labelled PFOS dose (4.3 mg/kg) and 93% of the labelled

PFOA dose (11 mg/kg) were retained by male rats within 24 hours. These estimates

were based on recovery of 5% and 7% of the total radioactivity in faeces and in the

digestive tract.

Uptake via inhalation was estimated by Kennedy et al. (1986), who measured

108 mg/L of ammonium perfluorooctanoate (APFO) in the blood of male rats which

were exposed ten times to 84 mg/m3 APFO.

Uptake via dermal exposure appears to be somewhat weaker (Kudo &

Kawashima 2003). Kennedy (1985) also showed a dose-dependent increase in blood

concentration of organofluoro compounds in rats after dermal application of APFO.

Subchronic dermal treatment with 2,000 mg/kg APFO resulted in a blood concentration

of 118 µg/mL.

5.1.2. Distribution

Both PFOS and PFOA are water soluble, weakly lipophilic and in the body are

principally found bound to proteins. In the body, PFOS and PFOA bind primarily to

serum albumin (Han et al. 2003), but they also bind to fatty acid binding proteins in the

liver (Luebker et al. 2002). The chain length and the functional group of PFCs have an

impact on the preferential binding sites and binding affinity, and it was found that the

PFCs have the same binding sites and similar affinity to proteins as fatty acids (Chen &

Guo 2009). PFOS and PFOA are primarily extracellular and accumulate in the liver,

blood serum and kidneys (Stahl, Mattern & Brunn 2011).

5.1.3. Bioaccumulation

According to the European Chemicals Regulation (Annex XIII REACH EC No.

1907/2006), a chemical is regarded as bioaccumulative if it has a bioconcentration

factor (BCF) in aquatic species higher than 2,000 (EU 2006). This differs from the US

EPA criterion (BCF 1,000–5,000) and the National Industrial Chemicals Notification

and Assessment Scheme (NICNAS 2013) criterion (BCF >5,000).


Appendix A

Both bioconcentration factor (BCF) and bioaccumulation factor (BAF) are generally

calculated as the ratio, at equilibrium, of internal biota concentration to exposure

concentration (McGeer et al. 2003). Although the calculation of BCF and BAF are

usually the same, the interpretations are slightly different, with accumulation in an

organism arising from multiple exposure conditions for BAF and from water only for

BCF. Therefore, in general, BCF is measured under laboratory conditions and BAF is

derived from measurements in natural environments.

Studies showed that the BAF values of PFOS are considerably greater than the

bioaccumulation criteria (table 2Table 2), which indicates that PFOS is highly

bioaccumulative in some aquatic species. The reported BCFs or BAFs of PFOA in

aquatic organisms are well below the bioaccumulation criteria.

The relatively high water solubility of PFOA could explain its effective excretion by fish

via gill permeation, facilitated by high water throughput (Vierke et al. 2012). Air-

breathing animals do not have this possible excretion pathway. Therefore, the defined

criteria for BCF in fish may not be the most relevant endpoint to consider for humans.

In the revised Annex XIII of the REACH regulation (Commission Regulation (EU) No.

253/2011), the bioaccumulation criteria are expanded to include more recent findings

with respect to biomagnification, bioaccumulation in terrestrial species and

concentrations in human body fluids (Gobas et al. 2009; Vierke et al. 2012). A number

of studies have reported the presence of PFOA in air-breathing animals, for example

piscivorous mammals and in high trophic level avian predators (Kannan et al. 2004), in

herring gull eggs PFOA ranged from 6.5–111 ng/g (ww) (Rüdel et al. 2011), and in

polar bear liver PFOA ranged from 3–13 ng/g (Martin et al. 2003). Furthermore,

Muller et al. (2011) studied the biomagnification behaviour of perfluorinated compounds

in terrestrial food webs consisting of lichen and plants, caribou, and wolves from

remote areas of northern Canada. They have reported that the calculated

biomagnification factor (BMF) and trophic magnification factor (TMF) ranged from 0.3–

11 and 1.1–2.4, respectively. Additionally, the detection of PFOA in human body fluids

strongly indicates bioaccumulation of PFOA. It can therefore be concluded that PFOA

is bioaccumulative.

5.1.4. Metabolism

It is well known from the literature that PFOS and PFOA do not undergo metabolism in

mammals (ATSDR 2015). Therefore, they are not subject to defluorination or phase-II

metabolism of biotransformation (Kudo & Kawashima 2003).

5.1.5. Elimination

Since PFOS and PFOA cannot be metabolised by mammals, these substances are

excreted unchanged (Stahl et al. 2011). PFCs are excreted in the urine and faeces,

and the reported biological half-lives differ among species and between genders in

some species (table 3 and table 4), due to differences in renal clearance rates (Kudo &

Kawashima 2003).


Appendix A

Table 2. BCF and BAF values for PFOS and PFOA

Organism Location BCF

(L/kg) BAF (L/kg) Reference

P F O S

Rainbow trout (Oncorhynchus mykiss)

Experimental, Ontario

3,100 (blood) 2,900 (liver)

NR (Martin et al. 2003)

Lake Trout (Salvelinus namaycush)

Lake Ontario NR 16,000 (Houde et al. 2008)

Alewife (Alosa pseudoharengus)


Rainbow Smelt (Osmerus mordax)


Sculpin (Cottus cognatus) Lake Ontario NR 95,000 (Houde et al. 2008)

Wild turtles (Trachemys scripta

elegans and Chinemys reevesii)

Arai River, Japan

NR 10,964 (Morikawa et al. 2006)

Fish Review 4,000>10,000 (Sinclair et al. 2006,

Fujii et al. 2007)

P F O A

Rainbow trout (Oncorhynchus mykiss)

Experimental, Ontario

25 (blood)

12 (liver)

NR (Martin et al. 2003)

Zebra fish (Danio rerio)

Experimental (14C-PFOA),

Sweden 20–30 NR (Ulhaq et al. 2015)

Wild turtles (Trachemys scripta)

Arai River, Japan

NR 3.2 (Morikawa et al. 2006)

Fish Review <200 (Fujii et al. 2007)

NR – not reported

Table 3. Elimination half-lives of PFOS/PFOA for various species (Lau et al. 2007, Stahl et al. 2011).

Species PFOS PFOA References

Males/females Males Females

Rat 100 days 4–6 h 2–4 h (Gibson & Johnson 1979,

Chang et al. 2008)

Mouse NR 19 days 17 days (Lau et al. 2006)

Rabbit NR 5.5 h 7 h (Hundley, Sarrif & Kennedy Jr 2006)

Dog NR 20–30

days

8–13

days (Hanhijärvi, Ophaug & Singer 1982)

Monkey 150 days 21 days 30 days (Perkins et al. 2004)

NR – not reported


Appendix A

Table 4. Elimination half-lives of PFOS/PFOA for humans

Population Mean

(years) Reference

PFOS

Retired production workers, US 5.4 (3.9–6.3) (Olsen et al. 2007)

Primiparous women, Sweden 8.7 (Gebbink, Glynn & Berger 2015)

Adult, China 6.2 (0.5) (Zhang et al. 2013)

Adult, Germany 4.3–4.8 (Yeung et al. 2013)

Cross-section (age 10 to 69),

US 4.3 (Olsen et al. 2012)

Infants, US 4.4 (Spliethoff et al. 2008)

General population, US 4.7 (Wong et al. 2014)

PFOA

Adult, China 2.1 (0.3) (Zhang et al. 2013)

Retired production workers, US 3.8 (3.1–4.4) (Olsen et al. 2007)

Infants, US 4.1 (Spliethoff et al. 2008)

Arnsberg residents, 2006–2008 2.93 (Russell, Waterland & Wong

2015)

For animals, the reported half-lives can range from 2 hours to 150 days. This is in

marked contrast to the half-life documented in humans. For humans, the mean half-

lives for PFOS and PFOA are 5.5 years and 3.2 years, respectively. This is because

humans have an active reabsorption process using organic anion transport proteins in

the kidney (ATSDR 2015). Due to the complicated toxicokinetic and toxicodynamic

processes involved, comparison of dose response across different species requires

derivation of a human equivalent dose (HED).

5.1.6. Total daily intakes for PFOS and PFOA

Limited information is available to estimate total daily intakes for Australians. The

actual pharmacokinetics of PFOS and PFOA in humans is not likely to be consistent

with a first order model distribution, but the model is able to describe the repeated dose

exposure and serum concentration reasonably well (Thompson et al. 2010).

Olsen et al. (2007) also used a first order model to monitor the serum concentration for

occupational workers, and good agreement was observed between observations

and simulations.

A first order model is applied based on available human biomonitoring data to calculate

total daily intakes for PFOS and PFOA. The relevant equation is:

/ /dc dt k c u f Vd (1)

where

c (ng/mL) is the serum concentration

k (/day) is the elimination rate

u (ng/kg/day) is daily uptake

Vd is the apparent volume of distribution (230 mL/kg bw for PFOS and

170 mL/kg bw for PFOA) (Thompson et al. 2010)

f is gastrointestinal absorption (91% as used in (Thompson et al. 2010), and

The initial condition is c(0)=0.


Appendix A

Thus, the relationship between serum concentration and daily uptake can be described

as:

k tc u, t u f / / Vd 1 e ;k (2)

Considering a stable stage, the equations can be converted into:

c u u f / k/Vd; (3)

where the elimination rate k=Ln(2)/half-life.

Combining all the equations, the final function between half-life and serum

concentration is:

c u half-life u f / / 0.693;Vd (4)

Thus, the daily uptake can be derived:

u c u 0.693 / half-life/f ;Vd (5)

Using the half-lives derived from Table 4 (5.5 years, 2007 days) for PFOS and 3.2

years (1168 days) for PFOA and the toxicokinetic model developed, together with the

background exposure information for Australian populations (10.2 and 4.5 ng/mL for

PFOS and PFOA, respectively), the total daily intakes for PFOS and PFOA are

estimated to be 0.89 ng/kg per day and 0.50 ng/kg per day, respectively.These

estimaties are similar to a recent study conducted in Australia (Thompson et al. 2010).

5.2 Toxicology of PFOS/PFOA

5.2.1. Acute toxicity

The available data indicate that acute toxicity is not observed following high exposure

of PFOS/PFOA by means of inhalation, ingestion, dermal or ocular contact in humans

(ATSDR 2015). However, animal studies demonstrate a moderate acute oral toxicity

resulting in harm to the liver and gastrointestinal tract (PHE 2009). Table 5 summarises

the acute toxicity of PFOS/PFOA in terms of median lethal dose (LD50) and median

lethal concentration (LC50) values. In the studies where PFOA and PFOS can be

directly compared, PFOS is found to be more toxic than PFOA in fresh water

organisms (Ji et al. 2008, Li 2009).


Appendix A

Table 5. Acute toxicity criteria for PFOS/PFOA in animals

Toxicity

criterion PFOS PFOA Animal species References

LD50 NR 430–680 mg/kg,

oral Rats

(Dean, Jessup &

Thompson 1978)

LD50 NR 189 mg/kg,

IP injection Rats

(Olson & Andersen

1983)

LD50 NR 0.98 mg/L APFO,

inhalation Rats (Kennedy et al. 1986)

LD50 NR 540 mg/kg APFO,

inhalation Rats (Griffith & Long 1980)

LD50 NR 4,300 mg/kg

APFO, dermal Rabbits (Kennedy et al. 1986)

LD50 NR 7,000/7,500 mg/kg,

dermal Male/female rats (Kennedy et al. 1986)

LC50 NR 2,000 mg/kg,

dermal Rabbits (Glaza 1995)

LD50 251 mg/kg,

oral NR Rats (US EPA 2000)

48/96 hr

LC50

27–233/

10–178 mg/l

181–732/337–672

mg/L

Four fresh water

species (water flea,

water snail, shrimp,

planarian

(Li 2009)

LC50 18 mg/l 200 mg/L Japanese water flea (Ji et al. 2008)

NR – not reported, LD50 – median lethal dose, LC50 – median lethal concentration, APFO – ammonium

perfluorooctanoate.

5.2.1 Health effects of chronic exposure: animal data

A number of studies over the years have investigated health outcomes in mice, rats

and monkeys resulting from chronic exposure to PFOS/PFOA. The most sensitive

effects are presented in table 6 and table 7. The following impacts were observed from

chronic exposure to PFOS and PFOA at low doses and are considered critical effects

for a Point of Departure (PoD) when deriving a TDI.

PFOS:

Increased liver weight

Liver cell hypertrophy

Histopathological changes to lungs

Decreased hormone level

Decreased reproductive outcome, and

Development delays.

PFOA:

Increased liver weight, and

Reduced Immunoglobulin M (IgM) antibody titres.


Appendix A

Table 6. Adverse effects observed in chronic exposure: animal studies (PFOS) (Wambaugh 2015)

Species Dose

(mg/kg per day) Exposure Endpoints

LOAEL

(mg/kg per day)

NOAEL

(mg/kg per day) Reference

Monkey 0.03, 0.15, 0.75 26 weeks Increased absolute and relative

liver weight 0.75 0.15 (Seacat et al. 2002)

Rat 0.038, 0.15, 0.38, 1.56 14 weeks Increased relative liver weight 1.56 0.38 (Seacat et al. 2003)

Rat 0.035, 0.14, 0.35, 1.4 14 weeks Centrilobular hepatic hypertrophy 0.35 0.14 (Seacat et al. 2003)

Mouse 1.5, 10, 15, 20 GD 1–18 Decreased pup survival 10 5 (Lau et al. 2003)

Rat 1, 2, 3, 5, 10 GD 2–20 Decreased pup survival and

development delays 2 1

(Lau et al. 2003,

Thibodeaux et al. 2003)

Rat 0.1, 0.4, 1.6, 3.2 63–76 days, oral gavage Decreased F1 reproductive

outcome 1.6 0.4 (Luebker et al. 2005a)

Rat 0.4, 0.8, 1.0, 1.2, 1.6, 2.0 63–76 days, oral gavage Decreased F1 reproductive

outcome 1.6 1.2 (Luebker et al. 2005a)

Rat 0.1, 0.4, 1.6, 3.2 63–76 days, oral gavage Development delays 0.4 0.1 (Luebker et al. 2005b)

Rat 0.14, 1.33, 3.21, 6.34 28 days Liver weight, decreased serum

total T4 1.33 0.14 (Curran et al. 2008)

Rat 0.14, 1.33, 3.21, 6.34 28 days Decreased total T4 1.33 0.14 (Curran et al. 2008)

Rat 0.15, 1.43, 3.73, 7.58 28 days Liver weight 0.15 n.r. (Curran et al. 2008)

Rat 0.15, 1.43, 3.73, 7.58 28 days Decreased total T4 1.43 0.15 (Curran et al. 2008)

Rat 15 28 days Thyroid 15 n.r. (Chang et al. 2008)

Mouse 0.00018, 0.0018, 0.0036, 0.019,

0.036, 0.18 28 days

Suppressed SBRC plaque-

forming cell response 0.0036 0.0018 (Peden-Adams et al. 2008)

Mouse 0.0083, 0.083, 0.42, 0.83, 2.08 60 days Increased splenic natural killer cell

activity 0.083 0.008 (Dong et al. 2009)

Rat 0, 0.1, 0.3, 1.0 GD 0–41 Decreased habituation response 1 0.3 (Chang et al. 2009,

Butenhoff et al. 2009)

Rat 0.1,2 GD 1–2 Histopathological changes to

lungs 2 0.1 (Chen et al. 2012)

LOAEL – low observed adverse effects level, NOAEL – no observed adverse effects level, NR – not reported.


Appendix A

Table 7. Adverse effects observed in chronic exposure: animal studies (PFOA) (Wambaugh 2015)

Species Dose

(mg/kg per day)

Exposure

duration Endpoints

LOAEL

(mg/kg per day)

NOAEL

(mg/kg per day) Reference

Monkey 3,10, 30/20a 26 weeks Increased liver weight 3 NR (Butenhoff et al. 2002)

Rat 1,3,10,30b 6 weeks pre

mating

Increased absolute and relative liver

weight 1 NR (Butenhoff et al. 2004)

Rat 0,0.06,0.64,1.94

and 6.5 13 weeks

Higher producing adaptive and reversible

liver changes 0.64 0.06 (Perkins et al. 2004)

Mouse 1,3,5,10,20,40 gestational day:

1–17

Increased liver weight for G1 and

development-accelerated sexual maturity

in males for G2

1 NR (Lau et al. 2006)

Mouse 3, 5 GD: 1–17 Increased liver weight for G1 3 NR (Wolf et al. 2007,

White et al. 2009)

Mouse 3.75,7.5,15,30 15 days Increased liver weight 3.75 NR (DeWitt et al. 2008)

Mouse 3.75,7.5,15,30 15 days Reduced IgM antibody titres 3.75 NR (DeWitt et al. 2008)

Mouse 0.3,1,3 GD1–17 Increased offspring relative liver weights 0.3 NR (Macon et al. 2011)

CD-1 0.3,1,3 GD1–17 Increased offspring relative liver weights 0.3 NR (Macon et al. 2011)

a) 30 mg/kg per day for the first 22 days and then reduced to 20 mg/kg per day, b) 30 mg/kg per day for the first 12 days and then reduced to 20 mg/kg per day, LOAEL – low

observed adverse effects level, NOAEL – no observed adverse effects level, NR – not reported.


Appendix A

Table 8. Epidemiological studies related to PFOS exposure

Disease/health indicator

Reference Study details PFOS level (ng/ml) Observations/health effects

Lipids (Geiger et al. 2014) Cross-sectional study, n= 815 ≤18 years), National Health and Nutrition Examination

Survey 1999–2008

Mean ± SE 17.7 ± 0.7

Exposure to PFOS was positively associated with high serum lipid levels in a representative, multi-

ethnic sample of US adolescents.

(Starling et al. 2014) Cross-sectional study, n= 819 pregnant

women, Norwegian Mother and Child Cohort Study 2003–2004.

Median plasma PFOS 13.03

PFOS concentration was associated with total cholesterol, which increased 4.2 mg/dL per inter-

quartile shift 95%CI = 0.8, 7.7) in adjusted models.

(Château-Degat

et al. 2010)

Cross-sectional, n=723 Inuit adults from Nunavik Quebec who are exposed to

environmental contaminants through their traditional diet. Study period fall 2004

Plasma mean 25.7

Triacylglycerol and ratio of total cholesterol to HDL-C levels were negatively associated with PFOS

plasma levels. HDL-C levels were positively associated, after adjustment for circulating levels of n-3 PUFAs and for the interaction between gender

and PFOS plasma levels.

(Frisbee et al. 2010)

Cross-sectional community-based study at Mid–Ohio River Valley during 2005–2006. 12,476 participants aged 1.0–17.9 years

included in the C8 Health Project.

Mean ± SD 22.7±12.6

PFOS was significantly associated with increased total cholesterol, HDL-C and LDL-C.

(Steenland et al. 2009)

Cross-sectional study C8 Project (2005–2006) among 46,294 community residents ≥18

years), who drank PFOA-contaminated water from a chemical plant in West Virginia

Mean 22.4

Range 0.25–759.2

Higher serum PFOS was associated with higher levels of total cholesterol, LDL and triglycerides.

Uric acid (Steenland et al. 2010)

Cross-sectional study under C8 project among 54,951 adult community residents in Ohio and West Virginia, who lived or worked in six water

districts contaminated with PFOA from a chemical plant. Study period 2005–2006

Mean ± SD 23.4±16.1

PFOS was significantly associated with serum uric acid.

Immune function (Grandjean et al. 2012,

Ashley-Martin et al. 2015)

A trans-Canada cohort study of 2,001 pregnant women and their newborns from 10 Canadian cities during 2008–2011. n=1,258 for statistical

analysis

No association between cord blood levels of PFOS

and immunotoxicity.

(Grandjean et al. 2012)

A longitudinal study of a birth cohort from the National Hospital in the Faroe Islands. A total

of 656 consecutive singleton births were recruited during 1997–2000, and 587

participated in follow-up through 2008. The serum antibody concentrations against tetanus

Maternal geometric mean 27.3

Children aged 5 years 16.7

Strong negative correlation between maternal pregnancy serum PFOS concentration and antibody

concentration in children of age 5 years.


Appendix A

and diphtheria were measured in children aged 5 and 7 years.

Thyroid function (Olsen et al. 2003)

Cross-sectional study in two PFOA manufacturing locations (Belgium and

Alabama), n=255 and 163 workers, longitudinal study, n=174 workers.

Mean 800 (Belgium)

1320 (Alabama)

Cross-sectional: no significant association of PFOS with T3, T4, or TSH.

Cancer (Alexander & Olsen 2007) Current and former workers in PFOS

production (PFOS), 1,400 questionnaires and 144 death certificates

1.30–1.97 Eleven cases of primary bladder cancer were identified from the surveys (n=6) and death

certificates (n 5)

(Alexander et al. 2003) Retrospective cohort mortality study, n=2,083 workers in PFOS production (USA). Minimal

time of employment is one year.

Geometric mean 100–900

High-exposure group: deaths resulting from bladder cancer, 3; standard mortality rate (SMR), 12.8. Two

deaths from liver cancer, SMR 3.08

Reproductive and developmental

outcomes (Fei et al. 2009)

General population, Danish National Birth Cohort, 1996–2002

Median plasma 33.7

Fertility disorders related to elevated plasma PFOS concentrations

(Monroy et al. 2008) General population, Canada (n=101 pregnant

women)

Mean ± SD maternal serum at delivery

16.19 ± 10.43

No association between serum PFCs and birth weight

(Inoue et al. 2004) General population (n=15), Japan <0.5 to 2.3 (maternal)

1.6–5.3 (fetal)

No significant correlations between PFOS concentration in maternal and cord blood samples and age bracket, birth weight, or levels of thyroid-

stimulating hormone or free thyroxine

Table 9. Epidemiological studies related to PFOA exposure

Disease/health indicator

Reference Study details PFOA level (ng/ml) Observations/health effects

Lipids (Geiger et al. 2014) Cross-sectional study, n= 815 (≤18 years), National Health and Nutrition Examination

Survey 1999–2008

Mean ± SE 4.2 ± 0.2

Exposure to PFOA was positively associated with high serum lipid levels in a representative, multi-

ethnic sample of US adolescents.

(Fu et al. 2014)

n = 133 (0–88 yrs old), which were randomly selected from the people coming for health

check-up in Yuanyang Red Cross Hospital of Henan, China in the year 2011

Median 1.43 Range

0.32–39.46

Serum PFOA concentration was positively associated with total cholesterol and LDL.

(Frisbee et al. 2010)

Cross-sectional community-based study at Mid–Ohio River Valley during 2005–2006. 12,476 participants aged 1.0–17.9 years

included in the C8 Health Project.

Mean ± SD 69.2 ± 111.9

PFOA was significantly associated with increased total cholesterol and LDL-C.

(Steenland et al. 2009) Cross-sectional study (C8 Project 2005–2006)

among 46,294 community residents (≥18 Mean 80 Range

Higher serum PFOA was associated with higher levels of total cholesterol, LDL and triglycerides.


Appendix A

years), who drank PFOA-contaminated water from a chemical plant in West Virginia

0.25–17,556.6

Uric acid (Steenland et al. 2010)

Cross-sectional study under C8 project among 54,951 adult residents in Ohio and West Virginia, who lived or worked in six water districts contaminated with PFOA from a chemical plant. Study period 2005–2006

Mean ± SD 86.4 ± 261.3

Higher serum levels of PFOA were associated with a higher prevalence of hyperuricemia.

(Costa, Sartori & Consonni

2009)

Longitudinal study (1978–2007) of medical surveillance of male workers engaged in PFOA

production plant; n=53

Currently exposed 0.20–47.04

Formerly exposed 0.53–18.66

Evidence for a significant association between total cholesterol and uric acid and PFOA serum level.

Cardiovascular disease

(Winquist & Steenland 2014)

A modelled PFOA exposure study included a community cohort (n=40,145; ≥20 years old; C8 projects participants) and a worker cohort (n=6,026; worked at the chemical plant during

1948–2002). Location Mid-Ohio Valley.

Mean ± SD 2005–2006 Community cohort

70.9 ± 151.2 Worker cohort 324.6 ± 920.6

combined cohort 86.6 ± 278.9

Higher PFOA exposure was associated with incident hypercholesterolemia with medication, but not with hypertension or coronary artery disease.

(Sakr et al. 2007, Sakr et al. 2009)

Cross-sectional study: n=1,025 active workers Cohort study: n= 4,747.

Washington Works site, where APFO is used, opened in 1948

Median in 2004 494

(among workers in the PFOA areas)

Cross-sectional study: exposure to ammonium perfluorooctanoate (APFO) has been associated

with increased serum lipid levels. Cohort study: no convincing evidence of increased ischaemic heart disease mortality risk for APFO-

exposed workers.

Immune function (Ashley-Martin et al. 2015)

A trans-Canada cohort study of 2,001 pregnant women and their newborns from 10 Canadian

cities during 2008–2011. n = 1,258 for statistical analysis

No association between cord blood levels of PFOA

and immunotoxicity.

(Emmett et al. 2006a) Cross-sectional community study, n=371,

Ohio, USA Median

354

No association between serum PFOA and with red cell indices, white cell or platelet counts. One small significant association was found between PFOA

and absolute monocyte counts.

Liver function (Costa et al. 2009) Longitudinal study (1978–2007) of medical

surveillance of male workers engaged in PFOA production plant; n=53

Currently exposed 0.20–47.04

Formerly exposed 0.53–18.66

No clinical evidence of any specific trouble or disease has been recorded over the 30 years, and

all the biochemical parameters, including liver, kidney and hormonal functions, were within

reference ranges.

(Emmett et al. 2006b) Cross-sectional community study, n=371,

Ohio, USA Median

354 No significant positive relationships between serum

PFOA and liver or renal function tests.

Thyroid function (Emmett et al. 2006a) Cross-sectional community study, n=371,

Ohio, USA Median

354 No significant positive relationships between serum

PFOA and TSH.


Appendix A

(Olsen et al. 2003)

Cross-sectional study in two PFOA manufacturing locations (Belgium and

Alabama), n=255 and 163 workers; Longitudinal study, n=174 workers

Mean 840 (Belgium)

1780 (Alabama)

Cross-sectional: no significant association of either compound with T3, T4, or TSH.

Longitudinal: No association between PFOA and thyroid hormones.

Cancer (Vieira et al. 2013) Residents in PFOA-exposed water districts and unexposed geographic areas outside of

the C8 Health Project area.

Estimated annual levels 3.7–655

for 10-year residency, assuming 10-year

latency

Probable link between higher PFOA serum levels and testicular, kidney, prostate, and ovarian

cancers and non-Hodgkin lymphoma.

(Barry, Winquist & Steenland 2013)

Adults who resided in contaminated water districts or worked at a local chemical plant in

Mid-Ohio Valley.

Median community 24.2

Median worker 112.7

Estimated cumulative serum PFOA concentrations were positively associated with kidney and

testicular cancer.

(Steenland & Woskie 2012) 5,791 workers exposed to PFOA at a DuPont

chemical plant in West Virginia. Serum data for 1,308 workers from 1979–2004

Average 350

Significant positive exposure-response trends occurred for both malignant and non-malignant

renal disease.

(Lundin et al. 2009) 3,993 workers of 3M plant (USA)

Definitely exposed 2,600–5,200

Probably exposed 300–1,500

No association with liver, pancreatic, and testicular cancer or cirrhosis of the liver. However, elevated

SMR for prostate cancer, cerebrovascular disease, and diabetes were evident.

Reproductive and developmental

outcomes (Fei et al. 2007) General population, Danish National Cohort

Average maternal 5.6

Inverse association between maternal plasma PFOA and birth weight.

(Bach et al. 2015) A review of 14 studies that addressed the

effects of PFCs on birth weight.

Higher PFOA concentration was associated with decreased average birth weight in most most

studies, but only some results were statistically significant.

(Vélez, Arbuckle & Fraser

2015)

A cohort study of 2,001 Canadian women (the Maternal-Infant Research on Environmental Chemicals Study) between 2008 and 2011.

Median 1.7

Increasing concentration of PFOA in maternal plasma was associated with reduced facundability

and infertility.

(Monroy et al. 2008) General population, Canada (n=101 pregnant

women)

Mean ± SD maternal serum at delivery

2.24 ± 1.61

No association between serum PFCs and birth weight.

(Nolan et al. 2009) General population, Ohio, USA, Median

354

No association between elevated PFOA exposure via drinking water and increased risk of lowered

birth weight or gestational age.

(Fei et al. 2009) General population, Danish National Birth

Cohort, 1996–2002 Median plasma PFOA

5.3 ng/mL Fertility disorders related to elevated plasma PFOA

concentrations.


Appendix A

5.2.2 Human data: Epidemiological studies

Biomonitoring of people in the community for evidence of PFCs in various countries

began in 2000, although occupationally exposed populations have been monitored for

much longer periods of time. In general, workers occupationally exposed have serum

levels of both PFOS and PFOA approximately one order of magnitude higher than

those reported in the general population (Lau et al. 2007).

In one study of occupationally exposed workers, the serum to plasma ratio for PFOS

and PFOA was 1:1, independent of the concentrations measured. The serum or

plasma to whole blood ratio was approximately 2:1 (Ehresman et al. 2007). This

implies that the PFOS/PFOA remains in the serum and does not enter the blood cells.

Human studies addressing potential PFOS/PFOA toxicity are growing, as more than

50 epidemiological studies published from 2015–16 onwards. A number of

epidemiological studies have been conducted as cross-sectional or longitudinal

analyses of routine medical surveillance at PFOS/PFOA production facilities. Thus

these studies focused on groups of people who were occupationally exposed to PFCs.

More recently epidemiological studies have also been conducted on general

populations. All the studies have attempted to assess the possible correlations

between PFC concentration in the human body and various health endpoints, and the

main approach taken in these studies has been to compare the incidence of disease

markers and/or other effects between the highest and lowest quartiles/tertiles of blood

PFAS levels. The results from recent epidemiological studies are summarised in

table 8 and table 9. The following observations may be made based on the

epidemiological studies.

Haematologic biomarkers:

there is relatively consistent evidence of modest positive associations of the

PFOS/PFOA in serum with elevated serum cholesterol

higher serum levels of PFOA are associated with increased uric acid in the blood,

but for PFOS, less pronounced trends are found

no significant association was observed with blood cell counts or thyroid

hormones, and

there is some but much less consistent evidence of a modest positive correlation

with increased liver enzymes in the blood.

Carcinogenicity:

no significant correlation between PFOS exposures and increased risks of cancer

has been reported, and

higher PFOA serum levels were found to be associated with kidney and testicular

cancer whereas null or inverse association was observed with lower exposure

groups. According to the recent classification of carcinogenic agents by IARC

Monographs (IARC 2015), PFOA is considered to be ‘possibly carcinogenic to

humans’ which leaves the previous classification unchanged.

Similar to the conclusions drawn in this review, this is on the basis of:

limited evidence in humans that PFOA causes testicular and renal cancer

limited evidence in experimental animals

moderate evidence for PFOA-associated carcinogenesis mechanisms


Appendix A

Reproductive and development toxicity:

positive associations between serum PFOS/PFOA concentrations and adverse

reproductive and developmental outcomes, were not observed in all epidemiology

studies. Fei et al. (2009) suggested that the plasma levels PFOS/PFOA may

reduce fecundity in the general population.

Mutagenic and genotoxic effects:

negative results in a large series of in vitro and/or in vivo short-term tests at gene

and/or chromosome or DNA repair levels suggest genotoxicity does not appear to

be a property of PFOS (EFSA 2008).

From the database on both in vitro and in vivo mutagenicity testing on the

ammonium salt of PFOA, APFO, it was concluded that PFOA should not be

considered a genotoxic substance (EFSA 2008).


Appendix A

6. Tolerable daily intake and the hazard quotient

Hazard quotients (HQ) indicate the overall risk to Australians from contaminants, and

are quotient between the total daily intake divided by tolerable daily intake (TDI). The

total daily intake was derived in section 5.

6.1 Tolerable daily intake

Two types of data may be used to derive a human TDI: human data and animal data.

Epidemiological studies have indicated a link between PFOS/PFOA exposure and

public health. However, most studies only show the odds ratio between control and

exposure groups, or the potential association between exposure and risk. Thus far,

epidemiological studies have concluded there is a weak positive association between

serum PFOS/PFOA and increased serum cholesterol and uric acid levels (table 8 and

table 9). Also, a few but inconsistent positive correlations occur with liver enzymes.

Although recent studies associate exposure to PFCs with adverse health outcomes,

most studies are cross-sectional analyses and therefore the data are insufficient to

draw any unambiguous conclusions about the effects of PFOS/PFOA in the progress of

any particular disease. At this stage, it is therefore difficult to establish a TDI based on

human blood (or serum) data.

As laboratory-based animal studies can mimic a pure exposure scenario and although

they exclude the impacts of extraneous factors, they may provide useful dose-response

data. This is despite the fact that interspecies uncertainties remain. Due to the

pharmacokinetic differences seen with PFCs in different species, administered doses

are not directly comparable. Administered doses are normalised to a human equivalent

dose (HED) using the species-specific half-lives and volume of distribution as

discussed in section 5.1.6.

US EPA and IPCS have highlighted two types of uncertainty when deriving human TDI

from animal studies: determination of the appropriate most sensitive critical effect for

the point of departure (PoD), and uncertainty introduced in extrapolation from the PoD

to a human TDI incorporating differences between species, routes of exposure, doses

and durations (Dong et al. 2015).

Over a 10-year period, various organisations such as UK COT, Europe EFSA, and US

EPA have developed TDI values for PFOS and PFOA (table 10). The early approaches

used included a default uncertainty factor of 100 to account for inter- and intraspecies

uncertainties (UK COT 2006a, UK COT 2006b). Later, EFSA applied another

uncertainty factor of 2 to connect the relatively short duration of the key study and the

internal dose kinetics (EFSA 2008). For its 2009 Provisional Health Advisory, the US

EPA used a chemical-specific adjustment factor (CSAF) method to account for the

interspecies uncertainties to estimate a TDI. These adjustment factors account for the

differences in volume of distribution and half-lives of the compounds in different

species. Specifically, a factor of 13 was used for PFOS, for extrapolation from monkey

to human, and a factor of 81 was employed for PFOA, for extrapolation from rat to

human. Together with the uncertainty factor of 10 to represent the intraspecies

uncertainties and uncertainty factor of 3 to denote the toxicodynamic uncertainties, the

total uncertainty factors in US EPA’s 2009 version are estimated to be 390 for PFOS

and 2430 for PFOA (US EPA 2009). In the 2016 health effects documents for PFOS


Appendix A

and PFOA (EPA 2016a, b) the pharmacokinetic differences between species were

accounted for by deriving a HED from modelled animal serum values.

Table 10. Comparison of TDI values from various organisations

Organisation PFOS (ng/kg per day) PFOA (ng/kg per day)

PoD UF TDI PoD UF TDI

UK COT 2006 a and b 30,000

(NOAEL) 100a 300

300,000

(BMDL) 100a 3,000

EFSA (2008) 30,000

(NOAEL) 200b 150

300,000

(BMDL) 200b 1,500

US EPA 2009 30,000

(NOAEL) 390 80

460,000

(BMDL) 2,430 200

BfR (Germany) (Priestly 2015) NA NA 100 NA NA 100

US EPA 2016a and b 510

(HED) 30 20

5300

(HED) 300 20

Danish EPA 2015 33,000

(BMDL10) 1230 30

3,000

(HED) 30 100

a) inter- and intraspecies uncertainties, b) UF of 100 for inter- and intraspecies uncertainties and UF of 2 to compensate for uncertainties in connection to the relatively short duration of the key study and internal dose kinetics.

In 2006, the UK committee on toxicity (COT) established a RfD of 300 ng/kg/day for

PFOS (table 1) (UK COT 2006a), based on serum T3 levels were decreased in a

26-week monkey study (Seacat et al. 2002). Two types of uncertainty factors (UFs)

were further considered inter-species UF of 10 and intra-species UF of 10.

Subsequently, the European Food Safety Authority (EFSA) issued a RfD of

150 ng/kg/day with an additional UF of 2 to account for this monkey study was not life

time exposure (EFSA 2008). Similarly, considering hepatic effects in male rats

(Palazzolo 1993; Perkins et al. 2004), the UK COT (2006b) and EFSA (2008) proposed

the PFOA RfDs of 3000 ng/kg/day and 1500 ng/kg/day (Table 10), respectively.

For its 2009 Provisional Health Advisory, the US EPA used a chemical-specific

adjustment factor method to account for the interspecies uncertainties to estimate a

TDI. These adjustment factors account for the differences in volume of distribution and

half-lives of the compounds in different species. Specifically, a factor of 13 was used

for PFOS, for extrapolation from monkey to human, and a factor of 81 was employed

for PFOA, for extrapolation from rat to human. Together with the uncertainty factor of

10 to represent the intraspecies uncertainties and uncertainty factor of 3 to denote the

toxicodynamic uncertainties, the total uncertainty factors in US EPA’s 2009 version are

estimated to be 390 for PFOS and 2430 for PFOA (US EPA 2009). In the 2016 Health

Effects Documents for PFOS and PFOA (US EPA 2016a, b) the pharmacokinetic

differences between species were accounted for by deriving a HED from modelled

animal serum values.

In June 2016, enHealth, the peak environmental health body in Australia, selected the

European Food Standards Authority (EFSA) TDIs for PFOS (and PFHxS combined)

and PFOA for application as an interim measure in Australia (enHealth 2016). In

September 2016, the Procedural review of health reference values established by

enHealth for PFAS noted that adoption of EFSA (2008) TDIs for PFAS in drinking and

recreational water was appropriate and is protective of public health

(Bartholomaeus 2016).


Appendix A

6.2 Hazard Quotient

The recommended enHealth interim TDI values are 150 mg/kg/d for PFOS/PFHxS and

1500 mg/kg/d for PFOA (based on EFSA 2008). Consequently the hazard quotients

(HQ) were estimated for PFOS and PFOA in order to indicate the overall risk to

Australians from these compounds. The HQ is the quotient between the total daily

intake (E) divided by TDI:

EHQ

TDI (9)

As the total daily intakes for PFOS and PFOA are estimated to be 0.89 ng/kg per day

and 0.50 ng/kg per day, respectively, the HQ were estimated to be 0.006 and 0.0003,

respectively.


Appendix A

7. Conclusions

1) The elimination rate of PFOS/PFOA in humans is much lower than that in

experimental animals. The half-lives in humans are estimated to be 5.5 and 3.2

years for PFOS and PFOA, respectively.

2) A recent study of Australians’ blood samples shows that the mean concentrations

for PFOS and PFOA are 10.2 ng/mL and 4.5 ng/mL, respectively. Several reports

based on national surveys indicate a decrease in serum PFOS and PFOA

concentrations from 2002 to 2011 in the Australian population.

3) Limited information is available to estimate total exposure for Australians, and thus

it is reasonable to use human biomonitoring data to estimate daily exposure. A first

order model can well describe the repeated dose exposure and stead serum

concentration. Thus, the total daily intake rates for PFOS and PFOA are estimated

to be 0.89 ng/kg per day and 0.50 ng/kg per day, respectively.

4) Animal studies suggest moderate acute oral toxicity of PFOS/PFOA resulting in

effects on the liver and gastrointestinal tract. Numerous health outcomes have

been observed in animals following chronic exposure including changes in the

liver, gastrointestinal tract, thyroid hormone levels, behaviour and reproductive and

developmental activity. However, no data are available to assess the level of acute

toxicity in humans.

5) Epidemiological studies have concluded that there is a weak positive association

between serum PFOS/PFOA and blood cholesterol and uric acid levels. Also, a

few but inconsistent positive correlations occur with liver enzymes. Although recent

studies associate exposure to PFCs with adverse health outcomes, most studies

are cross-sectional analyses and therefore the data are insufficient to draw any

unambiguous conclusions about the effects of PFOS/PFOA in the progress of any

particular disease.

6) Epidemiological studies of exposure to PFOS/PFOA and adverse health outcomes

in humans are not sufficient to quantify dose-response relationships. Based on

animal data, previous studies have used HED, BMDL or NOAEL approaches.

7) The hazard quotients are estimated to be 0.006 and 0.0003 for PFOS and PFOA,

respectively. These findings indicate that Australians are at a low risk of

experiencing adverse health effects from typical PFOS/PFOA exposures.


Appendix A

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APPENDIX B. HSL calculations

Note: See below for details of the input parameters used for the derivation of the HSLs in this guidance. An excel spreadsheet tool for calculating HSLs is provided separately. The 2016 enHealth-endorsed human reference values are used as defaults in the spreadsheet tool.


Table B1 soil HSLs – A Low density residential (HSL-A)

Symbol Units Child Adult Notes Exposure parameters Body weight BW kg 15 70

Exposure duration ED yrs 6 29

Exposure frequency EF d/yr 365 365

Soil/dust ingestion rate IGRs mg/d 100 50

Averaging time AT yrs 6 29

Target risk quotient RQtarget - 1 1

Chemical specific parameters Contribution to contamination CC -

PFOS+PFHxS 1

PFOA 1 1

Tolerable daily intake TDI mg/kg/d

PFOS+PFHxS 1.5E-04 1.5E-04 EnHealth 2016, EFSA 2008 PFOA 1.5E-03 1.5E-03 EnHealth 2016, EFSA 2008 Oral bioavailability AoF -

PFOS+PFHxS 1 1

PFOA 1 1

Health screening level HSL mg/kg PFOS+PFHxS 22.5 210 PFOA 225.0 2100


Table B2 soil HSLs – B High density residential (HSL-B) Symbol Units Child Adult Notes Exposure parameters Body weight BW kg 15 70



Soil/dust ingestion rate IGRs mg/d 25 12.5




PFOS+PFHxS 1 1

PFOA 1 1



PFOS+PFHxS 1 1

PFOA 1 1

Health screening level HSL mg/kg PFOS+PFHxS 90 840 PFOA 900 8400


Table B3 soil HSLs – C Open space/recreation (HSL-C) Symbol Units Child Adult Notes Exposure parameters Body weight BW kg 15 70 Exposure duration ED yrs 6 29 Exposure frequency EF d/yr 365 365 Soil/dust ingestion rate IGRs mg/d 50 25 Averaging time AT yrs 6 29 Target risk quotient RQtarget - 1 1 Chemical specific parameters Contribution to contamination CC - PFOS+PFHxS 1 1 PFOA 1 1 Tolerable daily intake TDI mg/kg/d PFOS+PFHxS 1.5E-04 1.5E-04 EnHealth 2016, EFSA 2008 PFOA 1.5E-03 1.5E-03 EnHealth 2016, EFSA 2008 Oral bioavailability AoF - PFOS+PFHxS 1 1 PFOA 1 1 Health screening level HSL mg/kg PFOS+PFHxS 45 420 PFOA 450 4200


Table B4 soil HSLs – D Commercial/industrial (HSL-D) Symbol Units Child Adult Notes Exposure parameters Body weight BW kg 70 Exposure duration ED yrs 30 Exposure frequency EF d/yr 240 Soil/dust ingestion rate IGRs mg/d 25 Averaging time AT yrs 30 Target risk quotient RQtarget - 1 Chemical specific parameters Contribution to contamination CC - PFOS+PFHxS 1 PFOA 1 Tolerable Daily Intake TDI mg/kg/d PFOS+PFHxS 1.5E-04 EnHealth 2016, EFSA 2008 PFOA 1.5E-03 EnHealth 2016, EFSA 2008 Oral bioavailability AoF - PFOS+PFHxS 1 PFOA 1 Health screening level HSL mg/kg PFOS+PFHxS 639 PFOA 6388


Table B5 HSLs fish consumption Symbol Units Child Adult Notes Exposure parameters Body weight BW kg 15 70



Fish consumption rate IGRs g/d 8.2 18




PFOS+PFHxS 1 1

PFOA 1 1



PFOS+PFHxS 1 1

PFOA 1 1

Maximum permissible concentration – fish MPCfish ug/kg

PFOS+PFHxS 274 583 PFOA 2744 5833


Freshwater Bioaccumulation factor BAF L/kg

PFOS+PFHxS 12900

PFOA 12900

HSL, freshwater HSLfw ng/L

PFOS+PFHxS 21

PFOA 213

Marine Bioaccumulation factor BAF L/kg

PFOS+PFHxS 4500

PFOA 4500 HSL, marine HSLfw ng/L PFOS+PFHxS 61 PFOA 610 Sediment Bioaccumulation factor BSAF L/kg

PFOS+PFHxS 12.7

PFOA 12.7

HSL, freshwater HSLfw ug/kg

PFOS+PFHxS 22

PFOA 216

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