Scholarly article on topic 'Perfluoroalkyl acids and their precursors in Swedish food: The relative importance of direct and indirect dietary exposure'

Perfluoroalkyl acids and their precursors in Swedish food: The relative importance of direct and indirect dietary exposure Academic research paper on "Environmental engineering"

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{"Per- and polyfluoroalkyl substances (PFASs)" / "Perfluoroalkyl acids (PFAAs)" / Precursors / "Dietary exposure" / "Temporal changes" / PAPs}

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

Abstract We analyzed food market basket samples obtained in Sweden from 1999, 2005, and 2010 for perfluoroalkyl acids (PFAAs) and a range of precursor compounds. Perfluorooctane sulfonic acid (PFOS) precursors were detected in all food year pools with the highest concentrations in 1999. Six polyfluoroalkyl phosphate diesters (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) were detected in the year pools with the highest ∑diPAP concentrations in 1999 and 2005. All precursors were predominantly found in meat, fish, and/or eggs based on analysis of individual food groups from 1999. Based on year pools, PFOS precursors contributed between 4 and 1% as an indirect source to total dietary PFOS intakes between 1999 and 2010. Perfluorohexanoic acid (PFHxA) exposure originated entirely from diPAPs, whereas for perfluorooctanoic acid (PFOA) and perfluorodecanoic acid (PFDA), diPAPs contributed between 1 and 19% to total exposure. The lowest precursor contributions were generally seen in food samples from 2010.

Academic research paper on topic "Perfluoroalkyl acids and their precursors in Swedish food: The relative importance of direct and indirect dietary exposure"

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Perfluoroalkyl acids and their precursors in Swedish food: The relative importance of direct and indirect dietary exposure


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

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


Article history:

Received 10 October 2014

Received in revised form

15 December 2014

Accepted 16 December 2014

Available online 12 January 2015


Per- and polyfluoroalkyl substances (PFASs)

Perfluoroalkyl acids (PFAAs)


Dietary exposure

Temporal changes


We analyzed food market basket samples obtained in Sweden from 1999, 2005, and 2010 for perfluoroalkyl acids (PFAAs) and a range of precursor compounds. Perfluorooctane sulfonic acid (PFOS) precursors were detected in all food year pools with the highest concentrations in 1999. Six polyfluoroalkyl phosphate diesters (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) were detected in the year pools with the highest J^diPAP concentrations in 1999 and 2005. All precursors were predominantly found in meat, fish, and/or eggs based on analysis of individual food groups from 1999. Based on year pools, PFOS precursors contributed between 4 and 1% as an indirect source to total dietary PFOS intakes between 1999 and 2010. Perfluorohexanoic acid (PFHxA) exposure originated entirely from diPAPs, whereas for perfluorooctanoic acid (PFOA) and perfluorodecanoic acid (PFDA), diPAPs contributed between 1 and 19% to total exposure. The lowest precursor contributions were generally seen in food samples from 2010.

© 2014 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) have been identified as environmental pollutants with perfluoroalkane sulfonic acids (PFSAs) and perfluoroalkyl carboxylic acids (PFCAs) as major compound classes of concern due to their persistence and bio-accumulative properties (Buck et al., 2011). There are also other PFASs that are less persistent and that can undergo degradation to form perfluoroalkyl acids (PFAAs, i.e., PFSAs or PFCAs). These PFASs are called precursors. Examples of precursors are perfluorooctane sulfonamidoacetic acids (FOSAAs) and perfluorooctane sulfonamides (FOSAs) as they can be biotransformed to perfluorooctane sulfonic acid (PFOS) (Benskin et al., 2009; D'Eon and Mabury, 2011a; Peng et al., 2014; Xu et al., 2004). PFOS precursors have been used as building blocks for commercial products (per-fluorooctane sulfonamidoethanols (FOSEs) or FOSAs) or are transformation products of these building blocks (FOSAAs). Fluorinated compounds that can be degraded to PFCAs include polyfluoroalkyl phosphate esters (PAPs) (D'Eon and Mabury, 2011a). These

* 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.

chemicals are commercial products that are applied to, e.g., food contact paper. Humans can be directly exposed to PFAAs or to precursors with subsequent metabolism to PFAAs. The latter pathway is considered an indirect exposure to PFAAs.

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). Only recently, monitoring studies have started including precursor compounds. Perfluorooctane sulfonamide (FOSA), per-fluorooctane sulfonamidoacetic acid (FOSAA), as well as its N-methyl and N-ethyl derivatives (MeFOSAA and EtFOSAA) and pol-yfluoroalkyl phosphate diesters (diPAPs) were detected in human serum from the US, Germany, Sweden, and/or Hong Kong (Glynn et al., 2012; Lee and Mabury, 2011; Loi et al., 2013; Olsen et al., 2004; Yeung et al., 2013a, 2013b).

Diet, drinking water, air, and dust have all been identified as human exposure pathways for PFAAs, although food has been suggested as the major direct exposure pathway for most PFAAs in the general population from Sweden and Norway (Haug et al., 2011a; Vestergren et al., 2012). These studies also showed that fish, meat, egg, and dairy products are dominant food groups with respect to human exposure to PFAAs; however precursors were not included in these studies. Human exposure pathways for precursors are not as well characterized as for PFAAs. PAPs have been

0269-7491/© 2014 The Authors. Published by Elsevier Ltd. This is an open access article under the CC BY-NC-ND license (

detected in drinking water (Ding et al., 2012), while PAPs, fluo-rotelomer alcohols (FTOHs), FOSAs, and FOSEs have been detected in dust and air (indoor and/or outdoor air) (De Silva et al., 2012; Haug et al., 2011b; Shoeib et al., 2011). With respect to dietary exposure to precursors, up to ten diPAPs were found in target food samples packed in material containing PAPs obtained from the Swedish market (Gebbink et al., 2013), and FOSA, MeFOSAA and EtFOSAA were detected in herring from the Swedish west coast (Ullah et al., 2014). Both these studies provided information on precursors in specific food items; however, information on precursor dietary intakes based on general diet, and the importance of precursor dietary intakes in relation to PFAA dietary intakes is lacking.

The first aim of this study was to identify and quantify PFAAs and selected precursors in Swedish market basket samples. Year pools from 1999, 2005, and 2010 representing the general diet in Sweden as well as 12 individual food groups from 1999 were analyzed. The second aim was to estimate dietary intakes for the precursors and PFAAs, and assess the relative contribution of each of these intake pathways to total dietary human exposure to PFAAs.

2. Materials and methods

2A. Chemicals and reagents

Native and labeled standards of all PFASs used in this study are listed in Table S1 in the Supporting Information. The terminology and acronyms are according to Buck et al. (2011). All solvents and reagents were of the highest commercial purity and employed as received.

2.2. Food basket samples

Food items were purchased at two major grocery store chains in four major Swedish cities in 1999 and 2005. In 2010 sampling was limited to Uppsala city, a metropolitan area in proximity to Stockholm, since no systematic geographical differences in food contamination with persistent organic pollutants was observed in the two earlier market basket studies (Darnerud et al., 2006). The food items (ca 130) were selected based on Swedish food and production statistics. By this method, ca 90% of the direct Swedish food consumption is covered when expressed on fresh weight basis. The food items, for example fish, meat, eggs, and pasta, were not cooked before analysis. For each sampling year, the food samples were initially stored in their original packaging as instructed on the package. Before the expiry date, the food items were divided into 12 groups (see Table S2) and homogenates for each food group were prepared by mixing food items proportionally according to food consumption statistics. These food group homogenates were stored at -20 °C in pre-clean glass vials. The vials and lids were washed using standard laboratory washing procedures, after which they were thoroughly rinsed with laboratory grade acetone. Acetone-rinsed aluminum foil was used as lining of the inner side of the lids. For 1999, all the 12 food groups were analyzed individually. Additionally, for all three sampling years, a homogenate was prepared by mixing proportional amounts of each food group according to consumption data for the respective year (Table S2), and each year pool was analyzed in duplicate.

2.3. Sample preparation

The extraction and clean-up of the samples was based on published methods (Gebbink et al., 2009, 2013). Briefly, homogenized food basket samples (2.5—5 g) were spiked with labeled internal standards (500 pg each) (see Table S1 for all internal standards

used) and 6 mL acetonitrile were added. The samples were mixed and sonicated for 15 min, after which the samples were centrifuged 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 methanol and 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 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. Prior to ultra-performance liquid chromatography-tandem mass spectrometry (UPLC/MS/MS) analysis, 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.

2.4. Instrumental analysis and quantification

Fraction 1 was analyzed for perfluoroalkane sulfonamides, while fraction 2 was analyzed twice, first for PFSAs, PFCAs, and perfluoroalkane sulfonamidoacetic acids and then separately 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. 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 quadru-pole mass spectrometer (Waters) operated in negative ion elec-trospray ionization (ESI-) mode. The capillary voltage was set at 3.0 kV, and the source and desolvation temperatures were 150 and 350 °C, respectively. The desolvation and cone gas flows (nitrogen) were set at 650 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 lacking an authentic standard quantification was performed using calibration curves of diPAP standards with similar chain lengths (Table S1). Previous studies have shown matrix effects (chain length specific signal enhancement) when analyzing for diPAPs in various matrices (De Silva et al., 2012; Gebbink et al., 2013). As there are only labeled internal standards for 6:2/6:2 and 8:2/8:2 diPAPs available, using these internal standards for the quantification of other diPAPs does not fully correct for the matrix effects during UPLC/MS/MS analysis. Results for these compounds should therefore be considered semiquantitative. Calibration curves (nine points) were linear over the whole concentration range with r2 values greater than 0.99 for all compounds.

The linear and the sum of branched isomers of PFOS and FOSA were separated by the chromatographic method 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). For the food year pools, and the individual fish, meat, egg, and fat food groups, an interference coeluted with branched PFOS isomers using m/z 80 as

product ion. However, using a different chromatographic method (using an Ace 3 C18 reversed phase column, 150 x 2.1 mm, 3 mm particles, Advanced Chromatography Technologies) (Verreault et al., 2007) the matrix peak was separated from both branched and linear PFOS.

2.5. Analytical quality control

For each batch of samples (n = 10) three method blanks were included to monitor for background contamination. Trace levels of PFOA, PFNA, and PFDA were detected in the blanks with concentrations between 0.3 and 1 pg/g (calculated for a 5 g sample). A batch-specific blank correction was performed for these compounds by subtracting the average quantified concentration in the three blanks from PFAS concentrations in the samples. Calculated absolute recoveries (including underlying potential matrix effects) of the labeled internal standards ranged between 6 and 119% for PFAAs and between 7 and 262% for PAPs (Table S5). Generally lower recoveries were obtained for FOSAs (1—48%), which was likely due to volatilization of these compounds during evaporation of the organic solvent in fraction 1. DiPAP responses have been shown to be influenced by matrix effects (De Silva et al., 2012; Gebbink et al.,

2013), which could explain the recoveries apparently exceeding 100% in some food groups. Due to the variability of recoveries and matrix effects, reported concentrations for diPAPs and other PFASs that lack an authentic internal standard or where internal standard recoveries are <30% should be considered as semi-quantitative. The semi-quantitative data is included in this study as internal standard quantification was performed and generally good agreement between sum individual food groups and year pool dietary intakes was observed (Section 3.4.) despite large differences in recoveries. Compound-specific method limits of quantification (MLOQs) were determined for the food matrices by calculating the PFAS concentration producing a peak with a signal-to-noise ratio of 10 in the sample chromatograms. Compound-specific MLOQs can be found in Tables S6—S8. Method limits of detections (MLODs; concentrations producing a peak with a signal-to-noise ratio of 3) are generally a factor of 3 lower compared to the MLOQs. Analyses of the food year pools showed a 9% deviation between duplicate samples for PFOS concentrations, <7% for PFOS precursors, and generally <30% for PFCAs and diPAPs. Larger variations between duplicate samples were occasionally seen for compounds quantified close to the MLOQ.

Table 1

Body weight normalized dietary exposure (pg/kg/d) to PFAAs and precursors calculated from analysis of 12 individual food groups from 1999 and year pools from 1999,2005, and 2010 in the lower bound scenario.

Compound 1999 1999 2005 2010

Dairy Meat Fats Pastries Fish Egg Cereal Vegetables Fruit Potatoes Sugar/ Soft Total Year Year Year

products products products products sweets drinks poola poola poola

PFHxS 4.5 12 3.4 12 0.45 0.90 33 55 26 20

br-PFOS 34 9.0 43 289 70 59

lin-PFOS 37 329 0.8 1.1 383 360 1.1 0.24 0.64 1.1 1110 1290 792 667

tot-PFOS 37 329 0.8 1.1 416 369 1.1 0.24 0.64 1.1 1160 1580 862 725

br-FOSA 15 40 56 78

lin-FOSA 4.9 205 3.9 214 114 70 23

tot-FOSA 20 245 3.9 269 192 70 23

FOSAA 27 27

MeFOSAA 30 30 3.2 63 52

EtFOSAA 53 6.4 32 13 104 108

PFHxA 15 4.8 20

PFHpA 17 0.3 5.4 1.1 24

PFOA 22 2.7 27 12 18 4.9 0.2 16 102 125 43 154

PFNA 3.0 14 2.4 0.6 35 8.0 1.4 0.2 2.2 67 191 440 422

PFDA 0.8 20 1.9 22 43 58 66

PFUnDA 19 0.8 55 14 89 70 129 170

PFDoDA 26 16 3.4 45 32 28 26

PFTrDA 43 5.3 48 22 23 34

PFTeDA 15 5.0 1.0 0.7 22 19 2.8 17

4:2/6:2 4.1 4.1 2.1 6.5

6:2/6:2 6.5 0.2 0.1 212 5.5 13 3.1 0.8 0.4 0.3 9.2 250 294 230 168

6:2/8:2 0.02 0.1 29 0.1 0.7 0.4 0.3 31 32 40 20

8:2/8:2 8.2 1.6 9.0 1.5 0.5 1.1 22 51 114 10

6:2/10:2 7.4 0.5 0.1 8.0 1.8 9.9 7.7

8:2/10:2 2.3 2.3

6:2/12:2 3.4 3.4

10:2/10:2 0.7 28 29 14 26 7.8

8:2/12:2 18 18

6:2/14:2 13 13

Note: If no values are reported, the concentrations were below MLOQ. Compound-specific MLOQs can be found in Tables S6—S8. a Dietary exposure was calculated from the average of duplicate analysis of the year pools.

2.6. Dietary intake estimations

Dietary intakes were estimated by multiplying the PFAS concentrations (in pg/kg) with the weight (kg) of annual intake of each food group (for individual food groups) or total weight (kg) of annually consumed food (for year pools). The annual intake per food group or year pool was divided by 365 (days per year) and 73.7 (mean body weight (kg) of men and women) to estimate a body weight normalized mean (per capita) dietary intake of the PFAAs and precursors in pg/kg/d. All dietary intakes were calculated using two scenarios. In the lower bound (LB) scenario, the concentrations <MLOQ were set as zero. In the upper bound (UB) scenario the concentrations <MLOQ were set as MLOQ.

3. Results and discussion

3.1. PFAAs in food samples

PFAAs have previously been analyzed in Swedish food basket samples from the same years by Vestergren et al. (2012). Overall the present PFAA data (Tables S6 and S7) were in agreement with the concentrations and calculated dietary exposure intakes from the Vestergren study (2012) (Figure S1). However, the dietary exposure estimates for PFOA for all three years were lower in the present study compared to Vestergren et al. (2012), whereas PFNA dietary intakes were higher (Figure S1). These differences could be the result of one or several of the following factors: (i) for the present study, pools for each food group were freshly made from banked food items and were thus different from the samples analyzed by Vestergren et al. (2012) (ii) Different analytical methodologies and quantification approaches were used. For example, Vestergren et al. (2012) did not correct analytical results for blank levels (Dr. R. Vestergren, Stockholm University, personal communication), which may have led to overestimation of especially PFOA concentrations and exposure estimates. (iii) The study design was different. Vestergren et al. (2012) analyzed all food groups individually and combined the concentrations to estimate dietary exposure intakes, whereas in the present study the year pool data was used. Comparison of these two calculation methods with the 1999 data in the present study shows that specifically for PFNA a lower dietary intake was calculated using the Vestergren approach (Table 1; 67 vs. 191 pg/kg/d).

Meat Fish Egg Pastries 1999 2005 2010

Individual food group 1999 Year pool

Fig. 1. Relative abundance (%) of PFOS and its precursors on a molar basis in individual food groups (meat, fish, egg, and pastries) collected in 1999 and year food pools from 1999, 2005, and 2010.

3.2. PFOS precursors in food samples

Of the monitored PFOS precursors, FOSA (linear and/or branched isomers), FOSAA, MeFOSAA and EtFOSAA were detected in the meat product, pastries, fish product, and/or egg food groups from 1999 (Table S6). FOSA concentrations were 9.7 and 12 pg/g in meat and egg food groups, respectively, while the concentration in fish was 495 pg/g. Concentrations of FOSAA, MeFOSAA, and EtFO-SAA in the individual food groups ranged between 2 and 79 pg/g. The highest concentrations of the precursors were generally found in the fish products and eggs, which was also the case for PFHxS and PFOS. The PFOS + precursor patterns in meat, fish, and egg food groups from 1999 were all dominated by PFOS (58-89% of SPFOS + precursors on a molar basis), while in pastries the pattern was dominated by EtFOSAA (Fig. 1). In the fish products, FOSA was the dominant precursor comprising 34%, while in meat and eggs EtFOSAA and FOSAA, respectively, were the major precursors. Sources of EtFOSAA in the food samples could include migration of EtFOSE-based chemicals (EtFOSE was used as a building block for surfactants used in food contact material (D'Eon and Mabury, 2011b)) from packaging into food and subsequent transformation, or environmental accumulation and transformation of EtFOSE. EtFOSE-based phosphate diester (SAmPAP diester) has been shown to migrate from food contact material (Begley et al., 2008), and degradation of this chemical (Peng et al., 2014) or EtFOSE (Plumlee et al., 2009) has produced EtFOSAA. As there were no food packaging related applications for MeFOSE (D'Eon and Mabury, 2011b), the MeFOSAA found in animal-related food groups originated most likely from animal exposure to MeFOSE and subsequent transformation to MeFOSAA. Further biotransformation of FOSEs and/or Me- and EtFOSAA is likely the source of FOSAA and FOSA as they were only detected in animal products, although migration from packaging material and subsequent transformation of EtFOSE-based chemicals cannot be ruled out as a source.

In the year pools, a declining trend in the FOSA concentration was observed, with a concentration of 7.7 pg/g in 1999, 2.6 pg/g in 2005, and 0.8 pg/g in 2010 (Table S6). MeFOSAA and EtFOSAA were detected in the 1999 year pool at 2.2 and 4.4 pg/g, respectively, and were below the detection limit in 2005 and 2010. The decline of PFOS precursors, and also PFOS itself, is probably a direct result of the phase out of these chemicals by the 3M Company in 2002. Since the PFOS concentration declined less rapidly over time compared to the precursors, the relative contribution of PFOS to SPFOS + precursors increased in the year pools between 1999 (82% on a molar basis) and 2010 (97%) (Fig. 1). In all the year pools FOSA was the dominant precursor (and the only detected precursor in 2005 and 2010), however, with declining relative abundance in the pattern from 1999 (10%) to 2010 (3%). Comparable trends with more rapidly declining levels of precursors relative to PFOS were also seen in Swedish herring between 1991 and 2011 (Ullah et al., 2014).

The production of PFOS and its precursors by 3M electrochemical fluorination produced a mixture of isomers, generally 70% linear and 30% branched (Martin et al., 2010). Both linear and branched isomers of PFOS and FOSA were quantified in meat products, fish products, and/or eggs as well as in the year pools (Table S6). Lin/br FOSA concentration ratios in meat and fish products were 25/75 and 84/16, respectively. This difference could be the result of several factors such as differences in exposure of farm animals and (wild) fish to linear and/or branched isomers or species specific differences in uptake, distribution, metabolism and excretion of linear and branched isomers. Exposure to FOSA precursors and ability to metabolize these isomer-specifically to FOSA could also have influenced the observed FOSA isomer patterns. In the 1999 year pool the lin/br concentration ratio for FOSA was 60/

Fig. 2. Relative abundance (%) of diPAP concentrations on a pg/g basis in individual food groups collected in 1999 and year food pools from 1999,2005, and 2010. DiPAPs were below detection limit in the dairy products.

40, whereas in the 2005 and 2010 year pools only the linear isomer was detected. The lin/br isomer ratio for PFOS was 92/8 in fish and 98/2 in eggs, while in meat only the linear isomer was detected. In the year pools the ratio was 81/19 in 1999 and 91/9 in both 2005 and 2010. The observed increases in the lin/br isomer ratios for both FOSA and PFOS over time in the year pools were also seen in herring from the Baltic Sea between 1991 and 2011 (Ullah et al., 2014). This is, however, contradictory to temporal changes in the PFOS isomer pattern commonly found in human serum. For example, in Swedish human sera the relative contribution of linear PFOS to total PFOS decreased from 69% to 61% between 1996 and 2010 (Glynn et al.,

2012). As was also stated by Ullah et al. (2014) for fish consumption, dietary exposure to PFOS based on the present food basket samples alone cannot explain the PFOS isomer pattern and trends observed in humans.

3.3. DiPAPs in food samples

Up to 10 diPAPs were detected in the individual food groups with 6:2/6:2 diPAP being the dominant diPAP in 8 out of 12 food groups (Table S8). With the exception of the dairy products, 6:2/6:2 diPAP was detected in all food groups with concentrations ranging between 0.2 and 428 pg/g. 6:2/8:2 and 8:2/8:2 diPAPs were also frequently detected, in 8 and 6 food groups, respectively, with concentrations ranging between 0.04 and 59 pg/g. The fish products contained not only the highest detection frequency amongst all food groups, but also the highest diPAP concentration, i.e., 659 pg/g (concentrations of diPAPs for which no authentic standards or internal standards were available should be considered semi-quantitative, see Materials and Methods section). Concentrations in the present samples, with the exception of the fish products, are comparable to concentrations in food samples from the Swedish market packed in PAP-coated material (Gebbink et al.,

2013). The diPAP patterns in the individual food market basket groups were dominated by 6:2/6:2 diPAP in 8 food groups, while 8:2/8:2 diPAP dominated the pattern in the three remaining food groups (no diPAPs were detected in dairy products) (Fig. 2). In the individual food groups ]TPFCA concentrations were 1.3—13 times higher compared to diPAP concentrations, however, in cereal and

fish products the ]TdiPAP concentrations were 19 and 1.5 times higher than ]TPFCA concentrations, respectively (Tables S7 and S8).

In the three year pools, 6 diPAPs were consistently detected (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; Fig. 2), with the highest concentrations of 6:2/6:2 diPAP (6—12 pg/ g). In 1999 and 2005 the ]TdiPAP concentrations were comparable (~16 pg/g), whereas in 2010 it was 8 pg/g (Table S8). 6:2/6:2 diPAP concentrations showed a decline in the year pools from 1999 to 2010, whereas for 4:2/6:2, 6:2/8:2, 8:2/8:2, 6:2/10:2, and 10:2/10:2 diPAPs the highest levels were quantified in the 2005 year pool. At all years, the diPAP pattern was dominated by 6:2/6:2 diPAP (54—79%), followed by 8:2/8:2 (5—27%) and 6:2/8:2 diPAP (8—9%) (Fig. 2, Table S8). For all year pools, PFCA concentrations were higher (1.3—4.1 times) compared to ]TdiPAP concentrations.

PAPs are approved for application to food packaging material, have been shown to migrate from packaging material into food (Begley et al., 2008), and have been detected in food samples obtained from the Swedish market that were packed in PAP-coated material (Gebbink et al., 2013). In technical mixtures the 6:2/6:2, 6:2/8:2, and 8:2/8:2 diPAPs have been identified as major compounds (Gebbink et al., 2013; Trier et al., 2011). These diPAPs also dominated the patterns in the majority of the individual food groups as well as in the three year pools. Migration of diPAPs from packaging material into the food samples could be a contamination source of these chemicals to the presently analyzed food samples, which were all bought from supermarkets. After purchasing there was only a short storage time of the food samples in their original packaging material before they were transferred to glass vials. Potential migration of PAPs from packaging material after the samples were purchased is therefore comparable for the samples from each time point. For the relatively high levels of diPAPs in fish, food web accumulation could have been an additional (or even the major) source of diPAPs besides contact with packaging material. DiPAPs have been detected in different environmental matrices (water and sediment) and have been found to accumulate in fish (Guo et al., 2012; Loi et al., 2013). As the aim of this study was to determine human dietary exposure to PFASs and not elucidate the sources of food contamination, potential sources of diPAP contamination to the food samples were not further investigated.

3.4. Estimated dietary intakes of precursors

Body weight normalized dietary intake estimations of PFAAs and precursors using the LB scenario for individual food groups from 1999 and for the three year pools are reported in Table 1 and for the UB scenario in Table S9. When comparing the cumulative intakes from the food groups in 1999 in the LB and the UB scenarios, there were relatively small differences (<25% deviation) for PFHxS, PFOS and its precursors (with the exception of FOSAA, which had 82% deviation), all PFCAs (with the exception of PFHxA, which had 61% deviation), and all diPAPs (with the exception of 8:2/10:2, 6:2/ 14:2, and 8:2/12:2 diPAPs, which had 34—40% deviation) (Figure S2). The good agreement between the dietary intakes estimated by the two scenarios demonstrates that the uncertainty in the intakes as a consequence of concentrations below the MLOQs in the food samples is low for most PFAAs and precursors. Therefore, further discussion of the dietary intakes will be based only on the LB scenario (Table 1). The estimated cumulative dietary intakes for the individual PFASs from the 1999 food groups were generally in agreement with the dietary intakes based on the 1999 year pool (Table 1). There were however, PFASs that were detected in some individual food groups but not in the 1999 year pool (i.e., FOSAA, PFHxA, PFHpA, and several diPAPs). The relative contributions of the food groups containing these PFASs to the year pool composition were low; therefore, PFAS concentrations were diluted in the year pools and fell below their respective MLOQs.

In 1999, the estimated dietary intakes of PFOS precursors from the year pool were 192 pg/kg/d for FOSA, 52 pg/kg/d for MeFOSAA, and 108 pg/kg/d for EtFOSAA, while FOSA dietary intakes were 70 and 23 pg/kg/d in 2005 and 2010, respectively (Table 1). Even though all FOSAAs were below their MLOQs in the 2005 and 2010 food pools, human biomonitoring data from Germany and the USA showed that the general population was still exposed to these precursors in recent years (Lee and Mabury, 2011; Yeung et al., 2013b). Dietary intake and subsequent biodegradation of FOSEs or SAmPAP diester could be exposure routes for FOSAAs in humans. Alternatively, also exposure pathways other than diet, such as

ingestion of dust or inhalation of air (e.g., via FOSEs) could be exposure sources of these PFOS precursors (Huber et al., 2011; Shoeib et al., 2011).

At all three time points, the dietary intake of 6:2/6:2 diPAP was the highest among the diPAPs, with 294, 230, and 168 pg/kg/d for 1999, 2005, and 2010, respectively (Table 1). For the remaining detected diPAPs in the three year pools (4:2/6:2, 6:2/8:2, 8:2/8:2, 6:2/10:2, and 10:2/10:2 diPAPs), dietary intakes between 2 and 114 pg/kg/d were calculated. For each of these diPAPs the highest intake was estimated for 2005. The ]TdiPAP dietary intakes in 1999 and 2005 calculated from the year pools were comparable (395 and 426 pg/kg/d, respectively), however, in 2010 the intake had decreased to about halfcompared to the earlier years (214 pg/kg/d). Changes in production and emission volumes and/or changes in the application of diPAPs to food packaging material could have led to the observed decrease. Analyses of food packaging materials obtained in 2012 showed several orders of magnitude lower diPAP concentrations compared to packaging material obtained in 2009 (Gebbink et al., 2013; Trier, 2011). Also, in 2006 many of the leading fluorinated chemical manufacturers joined a global PFOA stewardship with the goal to eliminate the emissions of PFOA and longer chain PFCAs as well as their precursors by 2015 (US EPA, 2006).

3.5. Precursor dietary intakes in relation to PFAAs

FOSA and FOSAAs have all been identified as precursors to PFOS in biotransformation studies (Benskin et al., 2009; Peng et al., 2014), however, it remains unclear to what extent humans can metabolize these precursors. When using a biotransformation factor determined in a study with rats (20% biotransformation of precursors to PFOS (Martin et al., 2010)) the total dietary intakes of PFOS (precursor intakes converted to PFOS on a molar basis) were 1640, 880, and 730 pg/kg/d for 1999, 2005, and 2010, respectively (Fig. 3A). Of these total intakes, the indirect PFOS intakes contributed 4, 2, and 1% in 1999, 2005, and 2010, respectively. As it is unclear whether the biotransformation factor obtained from a rat study can be applied to humans, precursor contribution was also

Fig. 3. Potential contribution of precursors (using biotransformation factors from rats) to dietary exposure (pg/kg/d) to PFOS (A), PFHxA (B), PFOA (C), and PFDA (D) from food pools from 1999, 2005, and 2010 in the lower bound scenario. Concentrations of precursors were converted to PFAA concentrations on a molar basis.

estimated assuming 100% biotransformation as an upper bound scenario. In this scenario, the indirect PFOS intakes contributed 17, 8, and 3% in 1999, 2005, and 2010, respectively, to the total dietary intakes (Figure S3). Regardless of these two scenarios the relative importance of the analyzed PFOS precursors in food as an indirect exposure source of PFOS was low and has decreased over the studied time period. These results are in line with findings from a recent study on PFOS and its precursors in herring filet (Ullah et al., 2014) and also in agreement with the conclusion that direct dietary exposure is currently a major human exposure pathway for PFOS in Sweden (Vestergren et al., 2008, 2012).

Dietary intake and subsequent metabolism of diPAPs could be an indirect human exposure pathway for PFCAs. In order to estimate the potential contribution of diPAPs (assuming 1% biotransformation of <6:2/6:2 diPAPs and 10% biotransformation of >6:2/ 6:2 diPAPs (D'Eon and Mabury, 2011a)) to the total dietary exposure to PFCAs, diPAP intakes were converted to their major PFCA metabolites on a molar basis (Fig. 3B—D). Cumulative dietary intake of various diPAPs resulted in an estimated indirect dietary exposure to PFHxA of 3.5, 3.5, and 2.3 pg/kg/d in 1999, 2005, and 2010, respectively (Fig. 3B). PFHxA itself was below MLOQ in all three food pools. For PFOA, the contribution of diPAPs was variable among the studied years (Fig. 3C). In 2005, 19% of the total PFOA dietary intake resulted from indirect exposure (diPAPs), whereas in 1999 and 2010 indirect exposure was estimated to account for <4%. The total dietary intakes of PFOA (direct + indirect) ranged between 53 and 155 pg/kg/d among the years. For PFDA, the majority of the total dietary intakes came from direct exposure to this PFCA (96—99% depending on the year; Fig. 3D). The total PFDA dietary intake based on the year pools ranged between 45 and 67 pg/kg/d. In the upper bound scenario assuming complete biotransformation of the diPAPs to PFCAs, PFHxA dietary intakes estimated from exposure to diPAPs were more than 50 times higher compared to the scenario using biotransformation factors from rat studies (Figure S3). For PFOA, up to 70% of the total dietary intakes resulted from indirect exposure (in 2005), whereas direct dietary intake of PFDA remained the major exposure source for this PFCA under the assumption of 100% biotransformation (Figure S3).

The relative importance of dietary diPAP exposure as an indirect source of PFCAs is thus chain length dependent. For PFHxA (which was not detected itself in the year pools), diPAPs constituted the only dietary exposure source, but resulting in low PFHxA exposures due to the low biotransformation factors, especially for 6:2/6:2 diPAP (1%). On the other hand, for PFOA and PFDA direct dietary exposure was dominant over indirect dietary exposure. However, high consumption of specific food items (fish or food packed in precursor containing materials (Gebbink et al., 2013)) could lead to elevated dietary intakes of precursors. A major uncertainty in the estimation of the relative contribution of precursors to human body burdens of PFAAs is the biotransformation efficiency. This parameter is poorly understood, yet very sensitive, as was demonstrated above with the two biotransformation factor scenarios for diPAPs. Future research should focus on the determination of compound specific human biotransformation factors for precursors. Regardless of the uncertainty in precursor biotransformation, the total dietary intakes (direct + indirect) of PFOS and PFOA determined in the present study were below the tolerable daily intakes reported by the European Food Safety Authority (EFSA, 2008).


This study was financially supported by the Swedish Research Council Formas (to W.A.G.; Grant number 219-2012-643).

Appendix A. Supplementary data

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


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