Scholarly article on topic 'Temporal changes (1997–2012) of perfluoroalkyl acids and selected precursors (including isomers) in Swedish human serum'

Temporal changes (1997–2012) of perfluoroalkyl acids and selected precursors (including isomers) in Swedish human serum Academic research paper on "Environmental engineering"

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{"Per- and polyfluoroalkyl substances" / Precursors / "Human serum" / "Temporal trend" / PAPs / Isomers}

Abstract of research paper on Environmental engineering, author of scientific article — Wouter A. Gebbink, Anders Glynn, Urs Berger

Abstract Concentrations (including isomer patterns) and temporal changes (1997–2012) of perfluoroalkyl acids (PFAAs) and selected perfluorooctane sulfonate (PFOS) and perfluoroalkyl carboxylic acid (PFCA) precursors were determined in serum samples from Swedish women. Perfluorooctane sulfonamide (FOSA) and perfluorooctane sulfonamidoacetic acid (FOSAA), as well as its N-methyl and N-ethyl derivatives (MeFOSAA and EtFOSAA) were consistently detected. Highest PFOS precursor concentrations were found for EtFOSAA (before year 2000) or MeFOSAA and FOSAA (after 2000). Disappearance half-lives for all PFOS precursors were shorter compared to PFOS. 4:2/6:2 and 6:2/6:2 polyfluoroalkyl phosphate diesters (diPAPs) were detected in <60% of the samples, whereas 6:2/8:2 and 8:2/8:2 diPAPs were detected in >60% of the samples, but showed no significant change in concentrations over time. Linear and sum-branched isomers were quantified separately for three PFAAs and three precursors. Significant changes between 1997 and 2012 in the % linear isomer were observed for PFOA and FOSA (increase) and PFOS (decrease).

Academic research paper on topic "Temporal changes (1997–2012) of perfluoroalkyl acids and selected precursors (including isomers) in Swedish human serum"

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Temporal changes (1997-2012) of perfluoroalkyl acids and selected precursors (including isomers) in Swedish human serum


Wouter A. Gebbink a' *, Anders Glynn b, Urs Berger

a Department of Environmental Science and Analytical Chemistry (ACES), Stockholm University, SE 10691, Stockholm, Sweden b Department of Risk and Benefit Assessment, National Food Agency, SE 75126, Uppsala, Sweden


Article history: Received 26 October 2014 Received in revised form 19 January 2015 Accepted 21 January 2015 Available online


Per- and polyfluoroalkyl substances


Human serum

Temporal trend



Concentrations (including isomer patterns) and temporal changes (1997-2012) of perfluoroalkyl acids (PFAAs) and selected perfluorooctane sulfonate (PFOS) and perfluoroalkyl carboxylic acid (PFCA) precursors were determined in serum samples from Swedish women. Perfluorooctane sulfonamide (FOSA) and perfluorooctane sulfonamidoacetic acid (FOSAA), as well as its N-methyl and N-ethyl derivatives (MeFOSAA and EtFOSAA) were consistently detected. Highest PFOS precursor concentrations were found for EtFOSAA (before year 2000) or MeFOSAA and FOSAA (after 2000). Disappearance half-lives for all PFOS precursors were shorter compared to PFOS. 4:2/6:2 and 6:2/6:2 polyfluoroalkyl phosphate diesters (diPAPs) were detected in <60% of the samples, whereas 6:2/8:2 and 8:2/8:2 diPAPs were detected in >60% of the samples, but showed no significant change in concentrations over time. Linear and sum-branched isomers were quantified separately for three PFAAs and three precursors. Significant changes between 1997 and 2012 in the % linear isomer were observed for PFOA and FOSA (increase) and PFOS (decrease).

© 2015 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND

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1. Introduction

Per- and polyfluoroalkyl substances (PFASs) are chemicals that have been used in industrial applications and in consumer products since the 1950s, and have been identified as environmental pollutants with perfluoroalkane sulfonic acids (PFSAs) and per-fluoroalkyl carboxylic acids (PFCAs) as major compound classes of concern (Buck et al., 2011). PFSAs and PFCAs are persistent in the environment and their long-chain homologues (>C6 PFSAs and >C8 PFCAs) have bioaccumulative properties. There are, however, also PFASs that are less persistent and that can undergo transformation. Methyl and ethyl perfluorooctane sulfonamidoethanols (Me- and EtFOSE) and perfluorooctane sulfonamides (Me- and EtFOSA) are PFASs that have been identified as precursors to perfluorooctane sulfonic acid (PFOS) (Benskin et al., 2009; Peng et al., 2014; Xie et al., 2009). The manufacturing of PFOS and its precursors resulted in a mixture of linear and branched isomers (Martin et al., 2010), and production was phased out by the main manufacturer

* Corresponding author.

E-mail address: (W.A. Gebbink). 1 Present address: Department of Analytical Chemistry, Helmholtz Centre for Environmental Research — UFZ, DE 04318, Leipzig, Germany.

between 2000 and 2002 in North America and Europe. Fluorotelomer-based chemicals such as polyfluoroalkyl phosphate esters (PAPs) have been identified as precursors to PFCAs (D'Eon and Mabury, 2011a). The production of fluorotelomer-based chemicals (producing only linear isomers) is still ongoing, however, industry has committed to reduce production of PFCAs with a chain length > C8 and other fluorinated chemicals that can degrade to these PFCAs, and to eliminate emissions of these chemicals by 2015 (US EPA, 2006).

External human exposure to PFOS isomers, PFCAs and precursors has recently been estimated by Gebbink et al. (2015a). Exposure pathways such as ingestion of food, dust, and drinking water and inhalation of air were all identified as sources of PFOS, PFCAs and/or their precursors to humans. Cumulative exposure intakes of PFOS precursors were estimated to be comparable to the daily intakes of PFOS itself, while daily intakes of 6:2, 8:2, and 10:2 fluorotelomer based chemicals were higher compared to per-fluorohexanoic acid (PFHxA), perfluorooctanoic acid (PFOA) and perfluorododecanoic acid (PFDA). In Sweden, diet and drinking water have been identified as exposure pathways for the general population for PFOS and its precursors (FOSA, perfluorooctane sulfonamidoacetic acid (FOSAA) and its methyl and ethyl derivatives (MeFOSAA and EtFOSAA)), while the diet was a source of

0269-7491/© 2015 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND license (

diPAP exposure (Filipovic and Berger, 2015; Gebbink et al., 2013, 2015b; Ullah et al., 2014). In food basket samples and herring, the relative importance of precursors for PFAA exposure has recently decreased over time (Gebbink et al., 2015b; Ullah et al., 2014), however, such temporal trend studies for other exposure pathways are lacking.

With respect to human biomonitoring, PFSAs and PFCAs have been studied extensively in different populations (Glynn et al., 2012; Haug et al., 2009; Kato et al., 2011; Yeung et al., 2013a, 2013b). For most of these populations, declining temporal trends in serum were seen for PFSAs and in particular PFOS. This was linked to the phase out of PFOS and related chemicals by industry in 2002. For longer chain length PFCAs (>C8), increasing trends were often observed in the studied populations. Recently, human biomonitoring studies have reported a variety of precursor compounds. FOSAA, MeFOSAA, and EtFOSAA, as well as diPAPs were detected in human serum from the US, Germany, and Hong Kong (D'Eon et al., 2009; Lee and Mabury, 2011; Loi et al., 2013; Olsen et al., 2004; Yeung et al., 2013a, 2013b). Declining temporal trends were seen in whole blood or serum for PFOS precursors, while diPAP concentrations and detection frequency were generally low. In Swedish serum samples only the PFOS precursor FOSA was analyzed so far with rapidly declining concentrations between 1996 and 2010 (Glynn et al., 2012). On the other hand, Liu et al. (2015) recently suggested that the relative importance of PFOS precursors for total internal exposure to PFOS has increased in the Swedish population (based on isomer and enantiomer pattern biomarkers). However, it is unclear which PFOS precursors (other than FOSA) and PFCA precursors the Swedish population has been and still is exposed to.

The first aim of this study was to investigate the presence, concentrations, and temporal changes of PFAAs and selected precursors in serum samples donated by nursing primiparous women from Uppsala, Sweden, between 1997 and 2012. Using a more sensitive analytical method compared to Glynn et al. (2012), quantitative data for a larger range of PFAA homologues were obtained. The second aim was to investigate whether temporal changes in the relative contribution of branched and linear PFAA and precursor isomers in serum could shed light on the importance of precursors for total human exposure to PFAAs and therewith confirm the findings for PFOS by Liu et al. (2015). The serum samples were analyzed for the following PFASs: C4,6,8,10 PFSAs, C4-14 PFCAs, FOSA and FOSAA and their N-methyl and N-ethyl derivatives, as well as 4 monoPAPs and 11 diPAPs.

2. Material and methods

2.1. Chemicals and reagents

Native and labeled standards of PFSAs, PFCAs, FOSAs, FOSAAs, and mono- and diPAPs used in this study are listed in Table S1 in the Supporting Information. All solvents and reagents were of the highest commercial purity and employed as received.

2.2. Serum samples

Between 1996 and 2012 primiparous women living in Uppsala County, Sweden, donated blood samples within the fourth week after delivery for a study on temporal trends of persistent haloge-nated organic compounds in pregnant and nursing women, the POPUP study (Persistent Organic Pollutants in Uppsala Primiparas) (Glynn et al., 2012; Lignell et al., 2009). In the present study, per year, 30 individual serum samples were pooled into 3 pools (9 or 10 individual samples per pool). Serum pools from the following years were included in this study: 1997 (each of the three pools

contained two individual serum samples collected in 1996), 1998, 2000, 2002, 2004, 2006, 2008, 2010, and 2012 (Table S2). For each year, two aliquots of all three serum pools were analyzed. The study was approved by the regional ethical vetting board in Uppsala, Sweden, and the participating women gave informed consent before donating the blood samples.

2.3. Sample preparation

The extraction and clean-up of the samples is based on published methods (Gebbink et al., 2013, 2015b). Briefly, serum samples (1 g) were spiked with labeled internal standards (500 pg each, see Table S1 for all internal standards used) and 3 mL acetonitrile were added. The samples were mixed by vortex and placed in a sonication bath for 15 min, after which the samples were centri-fuged for 5 min at 3000 rpm. The organic phase was transferred to a separate tube and the extraction procedure was repeated. The combined extracts were concentrated to ~1 mL under a stream of nitrogen. SPE WAX cartridges (150 mg, 6 mL, Waters) were conditioned with 6 mL methanol and 6 mL water. The sample extracts were loaded onto the WAX columns and the columns were washed with 1 mL 2% aqueous formic acid and then with 2 mL water. The columns were dried by applying a vacuum and by centrifugation before neutral compounds were eluted with 3 mL methanol (fraction 1). The ionic compounds were subsequently eluted with 4 mL of a solution of 1% ammonium hydroxide in methanol (fraction 2). Both fractions were dried under a stream of nitrogen and the residuals were redissolved in 150 mL of methanol. The extracts were filtered using centrifugal filters (modified nylon 0.2 mm, 500 mL, VWR International) and 13C8-PFOA and 13C8-PFOS (500 pg each) were added as recovery internal standards prior to ultraperformance liquid chromatography-tandem mass spectrometry (UPLC/MS/MS) analysis.

2.4. Instrumental analysis and quantification

Fraction 1 was analyzed for FOSAs, while fraction 2 was analyzed twice, first for PFSAs, PFCAs, and FASAAs and then for mono- and diPAPs. For all instrumental analyses, separation was carried out on an Acquity UPLC system (Waters) equipped with a BEH C18 (50 x 2.1 mm, 1.7 mm particle size, Waters) analytical column. Mobile phases were (A) 95% water and 5% methanol and (B) 75% methanol, 20% acetonitrile, and 5% water. Both mobile phases contained 2 mM ammonium acetate and 5 mM 1-methyl piperidine. The column temperature was set at 40 ° C, and the injection volume was 5 mL. Mobile phases, the gradient programs and flow rates for the different UPLC methods can be found in Tables S3 and S4. Connected to the UPLC system was a Xevo TQ-S triple quadrupole mass spectrometer (Waters) operated in negative ion electrospray ionization (ESI-) mode. The capillary voltage was set at 3.0 kV, and the source and desolvation temperatures were 150 °C and 350 °C, respectively. The desolvation and cone gas flows (nitrogen) were set at 650 L/h and 150 L/h, respectively. Compound-specifically optimized cone voltages and collision energies are listed in Table S1.

Quantification was performed using an internal standard approach. Analytes lacking an analogous labeled standard were quantified using the internal standard with the closest retention time (Table S1). Quantification was performed using the precursor — product ion multiple reaction monitoring (MRM) transitions reported in Table S1. For all precursor compounds an additional product ion was monitored for confirmation. For diPAPs for which no authentic standards were available, a technical mixture was used to optimize MRM channels and for confirmation of retention times (Gebbink et al., 2013). Quantification of these diPAPs was

based on calibration curves of authentic diPAP standards with similar chain length (Table SI). Results for these compounds should be considered semi-quantitative. For each compound a nine point calibration curve was made ranging from 0.03 to 300 pg/mL for PFAAs and 0.003—30 pg/mL for precursors. Calibration curves were linear over the whole concentration ranges with r values greater than 0.99 for all compounds.

The linear isomer and the sum of branched isomers of PFHxS, PFOS, FOSA, MeFOSAA, EtFOSAA, and PFOA were chromatographi-cally separated and quantified individually. Branched isomers were quantified using the linear isomer calibration curve. For quantification of the sum of branched PFOS isomers an average of the response obtained with the product ions m/z 80 and 99 was used (Berger et al., 2011).

2.5. Analytical quality control

In each batch of samples (n = 10) three method blanks were included to monitor for background contamination. For compounds where blank contamination was observed the method quantification limits (MQLs) were determined as the mean plus three times the standard deviation of the quantified procedural blank signals. A blank correction was performed by subtracting the average quantified concentration in the blanks from PFAS concentrations in the samples. For other compounds the MQL was determined as the concentration in a serum sample giving a peak with a signal-to-noise ratio of 10. Table S5 lists all compound-specific MQLs. Recoveries of the labeled internal standards in the serum samples are listed in Table S6, and ranged between 40% and 92% with the exception of 25% for d3-MeFOSA and 35% for 13C4-8:2/8:2 diPAP.

Table 1

Temporal trends of PFAAs and precursors in Swedish serum samples collected between 1997 and 2012.

Compound Disappearance half-life (yr) Doubling time (yr) P

PFBS 0.61

br-PFHxS 10.2 0.0004

lin-PFHxS 11.8 0.0001

tot-PFHxS 11.7 0.0001

br-PFOS 9.2 <0.0001

lin-PFOS 8.5 <0.0001

tot-PFOS 8.7 <0.0001

PFDS 4.2 0.0007

br-FOSA 2.6 <0.0001

lin-FOSA 3.5 <0.0001

tot-FOSA 3.3 <0.0001


FOSAA 3.5 <0.0001

br-MeFOSAA 3.4 <0.0001

lin-MeFOSAA 3.5 <0.0001

tot-MeFOSAA 3.5 <0.0001


lin-EtFOSAA 1.7 <0.0001

tot-EtFOSAA 1.8 <0.0001

PFHpA 0.26

br-PFOA 7.0 <0.0001

lin-PFOA 20.8 <0.0001

tot-PFOA 20.2 <0.0001

PFNA 15.9 <0.0001

PFDA 15.7 <0.0001

PFUnDA 16.3 <0.0001

PFDoDA 21.1 0.0003

PFTrDA 20.2 0.004

4:2/6:2 diPAPa

6:2/6:2 diPAPa

6:2/8:2 diPAP 0.13

8:2/8:2 diPAP 0.21

a Detection frequency <60%, therefore no trends were estimated.

Together with the serum samples a NIST standard reference material (SRM 1957) was analyzed in each batch. Concentrations obtained for several PFAAs in SRM 1957 were lower compared to the certified values, however, they were in agreement with reported concentrations in other studies (Table S7). Relative standard deviations (RSDs) for quantified concentrations in SRM 1957 (n = 9) were <20% for PFAAs (with the exception of PFTrDA), <25% for PFOS precursors, and between 15 and 57% for diPAPs. Duplicate analyses of the pooled serum samples showed <10% deviation for PFHxS, PFOS, and C8—C12 PFCAs, <25% deviation for PFBS, PFDS, PFHpA, PFTrDA, FOSA and FOSAAs, and <35% deviation for diPAPs. Larger variations were occasionally seen for compounds detected close to their respective MQLs.

2.6. Data treatment and statistical analysis

Doubling times (DT) or disappearance half-lives (T/ of individual compounds over the study period were only calculated when the detection frequency was >60% in the serum pools and when a significant trend was observed (p < 0.05). Concentrations below the MQLs were replaced with MQL/-—2 for statistical data treatment. Doubling times or disappearance half-lives were estimated with T1/2 = ln(2)/m, where m represents the slope of ln transformed concentration versus time. Correlations between concentrations of individual PFAAs and precursors in the serum samples were analyzed by Pearson's correlation test. Temporal changes in PFAA isomer patterns were only investigated when the detection frequency of both linear and branched isomers was >60% in the serum pools. Concentrations below the MQL were replaced with MQL/—2.

3. Results and discussion

3.1. Temporal trends of PFAAs in serum samples

PFAAs have previously been analyzed in pooled serum samples from the young mothers in the POPUP-cohort (Glynn et al., 2012; Liu et al., 2015). Overall the present PFAA concentrations (Tables S8 and S9) and temporal trends were in agreement with the results reported in the earlier study (Table S10). However, some differences were observed. For example, no significant concentration change was seen in the serum samples for PFBS over the presently studied time period (1997—2012). Glynn et al. (2012) reported significantly increasing concentrations between 1996 and 2010, which later was shown to be the result of a drinking water contamination situation in the city of Uppsala (Gyllenhammar et al., 2013). This contamination was also driving the observed increase of the PFHxS concentration over the investigated time period (Gyllenhammar et al., 2013). PFUnDA, PFDoDA, and PFTrDA were presently quantified in the serum samples with significantly increasing concentrations between 1997 and 2012 (Table 1), while these compounds were close to or below detection limit in all samples in the earlier study (MDLs reported by Glynn et al. (2012) were at least one order of magnitude higher compared to the MQLs in the present study). FOSA concentrations reported by Glynn and coworkers were systematically higher compared to the present data; however, the calculated half-life of FOSA was comparable. These differences could be attributed to one or several of the following factors: (i) for the present study, serum pools were freshly made and contained fewer individual serum samples in pools from the early part of the study period, (ii) different sample preparation and analytical techniques (including different types of MS detectors) were used, (iii) there were differences in the quantification procedure for, e.g., FOSA, and (iv) the study design was different; Glynn et al. (2012) investigated a

Fig. 1. Temporal trends of PFOS and its precursors (ng/g ww, sum linear + branched isomers) in pooled serum samples between 1997 and 2012. Disappearance half-lives (T/) are provided for each compound.

1997 1998 2000 2002 2004 2006 2008 2010 2012

Fig. 2. Relative abundance (%) of PFOS and its precursors (on a molar basis) in serum samples from primiparous women collected between 1997 and 2012. Note that only the range 80%-100% of the relative abundances is shown.

slightly different time period (1996-2010) and included samples collected in all years (except 2003 and 2005) resulting in higher statistical power compared to the present study with samples from alternating years collected between 1997 and 2012.

3.2. PFOS and its precursors in Swedish serum samples

Of the monitored PFOS precursors, FOSAA, MeFOSAA, EtFOSAA, and FOSA were quantified in the serum samples from all years included in this study (1997-2012) with declining concentrations over time (p < 0.0001) (Tables 1, S11, and S12, and Fig. 1). The concentrations in the individual samples ranged from 0.031 to 0.50 ng/g for FOSAA, 0.015-0.48 ng/g for MeFOSAA, <0.002-1.54 ng/g for EtFOSAA, and 0.002-0.75 ng/g for FOSA (Tables S11 and S12). EtFOSA was only detected in serum samples collected in 2000, 2002, and 2004 with concentrations ranging from 0.005 to 0.034 ng/g. PFOS concentrations in the serum samples were generally 1 -2 orders of magnitude higher compared to its precursors, and ranged between 5.6 and 20 ng/g. Significant positive correlations (p < 0.01) were observed between

concentrations of PFOS and its precursors, as well as among all precursors (Table S13). The concentrations of FOSAAs in the present Swedish serum samples were comparable to other European populations, however, lower compared to US serum samples (Lee and Mabury, 2011; Yeung et al., 2013b).

Exposure to the PFOS precursors detected in the present serum samples (FOSA, FOSAA, MeFOSAA, and EtFOSAA) could have occurred through different exposure pathways. Dust and food ingestion have been identified as direct exposure pathways for Me-and EtFOSAA (Beesoon et al., 2011; Gebbink et al., 2015b; Ullah et al., 2014), but also intake of Me- and EtFOSE (e.g., through inhalation of air and ingestion of dust (Shoeib et al., 2011)) and subsequent transformation is a potential exposure source to Me-and EtFOSAA (Peng et al., 2014). FOSAA has been reported in chicken eggs obtained in Sweden (Gebbink et al., 2015b), however, degradation of FOSEs could also be an indirect exposure pathway for this fluorinated chemical (Xu et al., 2004). FOSA detected in human serum could originate from direct exposure through one or more exposure pathways (diet, dust, drinking water, and air) (Gebbink et al., 2015a) or from intakes of fluorinated chemicals such as Me- and EtFOSE, Me- and EtFOSAA, and/or Me- and EtFOSA followed by biotransformation to FOSA (Peng et al., 2014).

Of the detected FOSAAs and FOSA in the serum samples, EtFO-SAA was the dominant precursor in 1997, 1998, and 2000, after this year MeFOSAA and FOSAA were the dominant precursors, with comparable concentrations to each other (Fig. 2). At all years FOSA had the smallest relative abundance of all PFOS precursors. The PFOS + precursors pattern in serum pools was dominated by PFOS, ranging from 91% (1997) to >99% (2010) (Fig. 2). Similar PFOS and precursor patterns were also seen in a German population (Yeung et al., 2013b). It should be stressed, though, that the relative abundances shown in Fig. 2 refer to concentrations in serum only and do thus not reflect total internal exposure, as the different PFASs distribute differently in the human body. For example, unlike ionic PFASs, FOSA has been reported to be more associated with the blood cell fraction than with the serum (Hanssen et al., 2013); this could possibly explain the low relative abundance of FOSA compared to the other precursors in the serum samples.

Fig. 3. Concentrations of 4:2/6:2, 6:2/6:2, 6:2/8:2, and 8:2/8:2 diPAPs (pg/g ww) in pooled serum samples between 1997 and 2012. The detection frequency of each diPAP in the serum pools and the significance of the temporal trend (if detection frequency was >60%) are provided. The dashed line indicates a non-significant temporal change in concentration.

Disappearance half-lives of the precursors based on the present serum samples ranged between 1.8 and 3.5 years, while for PFOS the disappearance half-life was 8.7 years (Table 1 and Fig. 1). These half-lives were also in accordance with PFOS half-lives reported for two German populations (Yeung et al., 2013b). The observed decreasing trends are likely the result of the production phase out of PFOS and its precursors in 2002.

When comparing disappearance half-lives (Fig. 1) and inspecting the PFOS + precursor pattern in serum (Fig. 2), an apparent decreasing trend in the relative importance of precursors for total PFOS exposure is seen. Although this trend would be consistent with the change in relative importance of (known) precursors in Swedish food (Gebbink et al., 2015b; Ullah et al., 2014), Liu et al. (2015) recently suggested that the relative importance of precursor exposure in relation to PFOS exposure for the Swedish population has actually increased since the late 1990s, based on isomer and enantiomer pattern biomarker analyses in serum. These observations do not necessarily conflict with each other, since the observed disappearance half-lives in serum are not only a function of temporal changes of external exposure, but also of individual PFAS pharmacokinetics (elimination half-lives including metabolism for precursors). Furthermore, Gebbink et al. (2015a) estimated that dust and air are dominant exposure media for precursors over diet and the Swedish women may have been exposed to PFOS precursors that were not analyzed in the present or previous studies (through any of the mentioned exposure pathways). In order to be able to unravel the role of PFOS precursors for total human exposure to PFOS over time, knowledge of phar-macokinetic parameters of all relevant PFOS precursors is essential, but currently not available.

3.3. DiPAPs in Swedish serum samples

In the serum pools, 4:2/6:2, 6:2/6:2, 6:2/8:2, and 8:2/8:2 diPAPs were detected, although detection of 4:2/6:2 and 6:2/6:2 diPAPs was infrequent (<60%) (Fig. 3 and S1, Table S14). Four monoPAPs (4:2, 6:2, 8:2, and 10:2) and seven other diPAP (4:2/4:2, 6:2/10:2, 8:2/10:2, 6:2/12:2, 10:2/10:2, 8:2/12:2, and 6:2/14:2) were below their respective MQLs in the present samples. 6:2/6:2 and 8:2/8:2 diPAPs were generally quantified with higher concentrations in the serum pools compared to 4:2/6:2 and 6:2/8:2 diPAPs. For 6:2/6:2 diPAP the concentrations above MQL ranged between 6.6 and 55 pg/g, and for 8:2/8:2 diPAP between 2.1 and 18 pg/g. Quantified

concentrations for 4:2/6:2 diPAP and 6:2/8:2 diPAP were below 1.3 and 2.7 pg/g, respectively. No significant relationships were observed between concentrations among the four detected diPAPs (p > 0.05, Table S15). DiPAPs detected in the Swedish serum samples have also been found in German, Hong Kong, and US serum samples (Lee and Mabury, 2011; Loi et al., 2013; Yeung et al., 2013a). Over the studied time period, temporal changes were only investigated for 6:2/8:2 and 8:2/8:2 diPAPs (detection frequency >60%), however, no significant temporal changes were observed (Table 1, Fig. 3). The lack of temporal trends of diPAPs in the Swedish serum pools is consistent with the absence of trends in Germans between 1982 and 2009 (Yeung et al., 2013a).

The Swedish population was recently exposed to six diPAPs (4:2/6:2, 6:2/6:2, 6:2/8:2, 8:2/8:2, 6:2/10:2, and 10:2/10:2 diPAPs) through their general diet, however, other diPAPs have also been reported to occur in specific food samples (fish, popcorn, and potato chips) (Gebbink et al., 2013, 2015b). Furthermore, dust and drinking water have been identified as human exposure media containing diPAPs (De Silva et al., 2012; Ding et al., 2012), however, not specifically in Sweden. From Swedish food basket samples (collected in 1999, 2005, and 2010) the 6:2/6:2 diPAP dietary intake was highest in 1999 and decreased at each following time point, whereas the dietary intakes for 4:2/6:2, 6:2/8:2, 8:2/8:2, 6:2/10:2, and 10:2/10:2 diPAPs peaked in 2005 (Gebbink et al., 2015b). Despite these indications of changes over time in dietary exposure to diPAPs, no temporal trends were observed in the serum samples. The detection frequencies of 4:2/6:2 and 6:2/6:2 diPAPs were too low in order to determine any temporal change in concentrations, whereas for 6:2/8:2 and 8:2/8:2 diPAPs indications of increasing concentrations were seen over time, although this increase was not significant (Fig. 3). Although food contributes to human exposure to diPAPs, dust was recently suggested to be the major exposure medium for diPAPs (Gebbink et al., 2015a). This could explain why the temporal variations in dietary intakes were not directly reflected in the serum samples.

DiPAPs have been identified as PFCA precursors (D'Eon and Mabury, 2011b). Therefore, relationships between diPAP and PFCA concentrations were investigated. Potential degradation products of the diPAPs detected in the present serum samples include C4-C9 PFCAs; however, of these PFCAs only PFHpA, PFOA, and PFNA were detected whereas PFBA, PFPeA, and PFHxA were below their respective MQLs. No significant correlation was found between any diPAP and their potential degradation products (Table S15).

Fig. 4. Temporal changes in the relative contribution of the linear isomer to sum linear + branched isomers for PFHxS, PFOS, FOSA MeFOSAA and PFOA in human serum samples between 1997 and 2012. A solid line indicates a significant temporal change in the isomer pattern, whereas a dashed line indicates a non-significant temporal change.

Exposure intakes for individual diPAPs were recently estimated to be higher compared to the corresponding PFCAs as degradation products (Gebbink et al., 2015a); however, diPAP concentrations in the present serum samples were lower compared to PFCA concentrations. Information regarding pharmacokinetic parameters such as metabolism of diPAPs, elimination half-lives and/or volume of distribution for individual PFCAs and diPAPs is either extremely limited or non-existing. These knowledge gaps hinder the assessment whether there were temporal changes in the relative importance of diPAPs for total human exposure to PFCAs based on the present dataset.

3.4. PFAA and precursor isomers in Swedish serum samples

In the present study the linear and the sum of all branched isomers of several PFAAs and precursors were chromatographically separated and quantified. Linear and branched isomers of PFHxS, PFOS, and PFOA were detected in all serum samples. For PFHxS, the isomer pattern contained between 91.4% and 93.7% linear PFHxS, linear PFOS contributed between 61.9% and 65.8% to total PFOS, while the isomer pattern of PFOA contained between 97.8% and 99.2% linear PFOA (Fig. 4, Tables S8 and S9). Linear and branched FOSA isomers were detected in all serum samples, while for MeFOSAA and EtFOSAA the branched isomers were detected in 89% and 33% of the serum pools, respectively (linear isomers were detected in all samples) (Fig. 4, Tables S11 and S12). For FOSA the linear isomer contributed between 70.3% and 94.9% to total FOSA, while the percentage of the linear MeFOSAA and EtFOSAA isomer ranged between 90.5%-97.9% and 97.8%—99.6%, respectively.

Over the studied time period, decreases were seen in the relative abundance of the linear PFHxS and PFOS isomer (although only significant for PFOS), while the PFOA isomer pattern was significantly enriched with the linear isomer between 1997 and 2012 (Fig. 4).

The relative abundance of linear FOSA and MeFOSAA increased between 1997 and 2012; however, the increase was not significant for MeFOSAA (Fig. 4). Branched EtFOSAA isomers were only detected in serum pools from 1997 to 2002; therefore, no temporal changes were determined.

Reports on PFAA and/or precursor isomer patterns in human

serum are limited. PFOS and/or PFOA isomer patterns have been reported in human serum from Sweden, Norway, China, and the US (Berger et al., 2011; Glynn et al., 2012; Gutzkow et al., 2012; Haug et al., 2009; Liu et al., 2015; Zhang et al., 2013b). The PFOS and PFOA isomer patterns reported in these studies, as well as temporal changes in the PFOS isomer pattern are comparable to the present study. To our knowledge, there are presently no literature data available on temporal changes in PFHxS, FOSA, MeFOSAA, and EtFOSAA isomer patterns in human serum samples.

Although PFOA isomer profiles have been reported in human serum samples, no temporal changes in PFOA isomer pattern have been investigated so far. PFOA has historically been produced by both electrochemical fluorination (ECF), resulting in ~78% linear and ~22% sum branched PFOA (Zhang et al., 2013b), and by telo-merization, exclusively producing the linear isomer. Information on PFOA isomer patterns in human exposure media is extremely limited. In Swedish drinking water the isomer pattern contained >90% linear PFOA (Filipovic and Berger, 2015), however, drinking water was estimated to only contribute 15% to the total PFOA intakes (Gebbink et al., 2015a). The observed PFOA isomer pattern and its temporal trend in the serum samples (Fig. 4) could be explained by several factors. (i) The production of ECF PFOA was largely phased-out by 2002, whereas the production of telomer-based PFOA and precursor chemicals is still ongoing; and (ii) branched PFOA isomers are excreted faster than linear isomers (Zhang et al., 2013a).

The major part of the cumulative historic production of PFOS and its precursors occurred by ECF, producing a mixture of ~70% linear and ~30% sum branched isomers (Martin et al., 2010). Information on the isomer pattern of precursors in human exposure media is extremely limited. FOSA isomer patterns have been reported in Swedish food (with increasing importance of linear FOSA in the isomer pattern between 1991 and 2011; Gebbink et al., 2015b; Ullah et al., 2014), and in drinking water (Filipovic and Berger, 2015), while for FOSAAs no information on isomer patterns in exposure media is available. The enrichment of linear FOSA in the isomer pattern over time seen in human serum (Fig. 4), in the situation of generally decreasing levels (Fig. 1), is in line with the trends observed in food. This trend could also (partly) be explained by shorter disappearance half-lives of branched isomers compared

to the linear isomer (Table 1), which could be the result of faster biotransformation of branched isomers to branched PFOS relative to the linear isomer (Benskin et al., 2009; Ross et al., 2012). Faster urinary elimination of linear precursors relative to branched precursors, as was observed in humans (Zhang et al., 2013a), would, on the other hand, result in increasing relative importance of branched precursor isomers over time. Urinary elimination appears thus to be a minor mechanism influencing the FOSA isomer patterns in the present samples. Isomer-specific transformation of other fluorinated chemicals present in the serum samples (MeFOSAA and EtFOSAA) could also have influenced the FOSA isomer pattern.

Isomer patterns of PFOS have been reported in human exposure media (dust, diet, drinking water, and/or air), as summarized by Gebbink et al. (2015a). This study showed that the PFOS isomer pattern of total PFOS intake largely reflects the isomer pattern in food, since dietary intake is the dominant exposure pathway. In Swedish food an increase in the relative abundance of the linear isomer in the PFOS isomer pattern has been reported, from 81% linear PFOS in 1999 to 91% in 2005 and 2010 (Gebbink et al., 2015b). In all serum samples of the present study the PFOS isomer pattern was enriched with branched isomers (33-40%) relative to the isomer pattern reported in Swedish food. Furthermore, the serum samples showed a decreasing abundance of linear PFOS in the isomer pattern, which was opposite to the change in isomer pattern in food. Faster biotransformation of branched precursors to branched PFOS relative to linear precursors could have contributed to increasing abundance of branched PFOS over time; however, Gebbink et al. (2015a) recently estimated that precursors only contributed 16% to the total PFOS intake. Even when including (known) precursor contribution, the isomer pattern of total PFOS intake cannot explain the PFOS isomer patterns and trends found in human serum (Gebbink et al., 2015a). This lack of agreement between the temporal changes in isomer pattern of intake and serum illustrates the incomplete understanding of isomer-specific phar-macokinetic processes of PFAAs and their precursors.


We are grateful to the participating women who donated blood samples to the study. The anonymous reviewers are thanked for excellent and insightful comments on the manuscript. This study was financially supported by the Swedish Research Council Formas, grant number 219-2012-643 (to W.A.G.). The Swedish EPA gave financial support for the recruitment and sampling of the study participant.

Appendix A. Supplementary data

Supplementary data related to this article can be found at http://


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