Environment International 63 (2014) 130–136

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Polybrominated diphenyl ethers in UK human milk: Implications forinfant exposure and relationship to external exposure

Mohamed Abou-Elwafa Abdallah a,b,⁎, Stuart Harrad b

a Division of Environmental Health and Risk Management, College of Life and Environmental Sciences, University of Birmingham, Birmingham B15 2TT, United Kingdomb Department of Analytical Chemistry, Faculty of Pharmacy, Assiut University, 71526 Assiut, Egypt

⁎ Corresponding author at: Division of EnvironmentalCollege of Life and Environmental Sciences, University of2TT, United Kingdom. Tel.: +44 121 414 7297; fax: +44

E-mail address: mae_abdallah@yahoo.co.uk (M.A.-E. A

0160-4120/$ – see front matter © 2013 Elsevier Ltd. All rihttp://dx.doi.org/10.1016/j.envint.2013.11.009

a b s t r a c t

a r t i c l e i n f o

Article history:Received 12 August 2013Accepted 6 November 2013Available online xxxx

Keywords:PBDEsHuman milkInfant exposureBDE 209

Fourteen tri-deca polybrominated diphenyl ethers (PBDEs) were investigated in 35 human milk samples fromBirmingham, UK. While none of the hepta–nona BDEs (the main components of the octaBDE technical mixture)was above the limit of quantitation (LOQ); BDE-47 (average concentration = 3.3 ng g−1 lipid weight (lw)) wasquantified in all samples contributing 34–74% to Σtri-hexa BDEs (the principal constituents of the pentaBDEcommercial formulation). BDE-209 (the main congener in the decaBDE formulation) was present above theLOQ in 69% of samples (average concentration = 0.31 ng g−1 lw). Concentrations of Σtri-hexa BDEs rangedfrom 0.2 to 26 ng g−1 lw with concentrations of BDE-47 N BDE-153 N BDE-99. While concentrations ofΣtri-hexa BDEs in this study (average = 5.95 ng g−1 lw) were at the high end of those reported from otherEuropean countries, concentrations of BDE-209 were lower than those reported in human milk from other coun-tries. The average exposure of a UK nursing infant to Σtri-hexa BDEs (35 ng (kg bw)−1 day−1) via breast milkexceeded the upper-bound dietary intakes of both UK adults and toddlers. Using a simple one compartment phar-macokinetic model, PBDE intakes of UK adults via inhalation, diet and dust ingestion were converted to predictedbody burdens. Predictions compared well with those observed for Σtri-hexa BDEs and BDE-209 in breast milk.

© 2013 Elsevier Ltd. All rights reserved.

1. Introduction

Polybrominated diphenyl ethers (PBDEs) have been extensivelyused as flame retardants for a wide range of consumer products includ-ing furniture, carpets, mattresses and casings for electronic equipment(BSEF, 2013). Three technical PBDE formulations were commerciallyavailable: Penta (consisting primarily of BDE-47 and BDE-99 — 38–49% each, alongside smaller amounts of other tri- to hepta-BDEs), octa(a mixture of hexa- to deca-BDEs — the exact congener compositionvarying substantially between the two principal formulationsmarketed) and deca (92–97% decabromodiphenyl ether – BDE 209 –

plus nona- (principally) and octa-BDEs) (La Guardia et al., 2006).DecaBDE has dominated worldwide production with a global marketdemand of 56,100 t in 2001, compared to 7500 and 3790 t for pentaBDEand octaBDE formulations respectively (BSEF, 2013). Despite their util-ity, the persistence and bioaccumulative characters of these compoundshave resulted in increasing concern over their potential adverse effectsto human health (Frederiksen et al., 2009; Harrad et al., 2010). Animalstudies have shown PBDEs to pose potential health risks including:endocrine disruption, neurodevelopmental and behavioral outcomes,hepatic abnormality and possibly cancer (Birnbaum and Staskal, 2004;

Health and Risk Management,Birmingham, Birmingham B15121 414 3078.bdallah).

ghts reserved.

Darnerud, 2008; Hakk, 2010; Wikoff and Birnbaum, 2011). The fewdata available from human epidemiological studies imply effects on:male reproductive hormones (Johnson et al., 2013; Palace et al., 2010),semen quality (Akutsu et al., 2008), thyroid hormone homeostasis(Turyk et al., 2008), cryptorchidism (Main et al., 2007), behavioral fac-tors in pregnant women (Buttke et al., 2013), as well as lower birthweight and length (Chao et al., 2007; Lignell et al., 2013). Such evidencehas contributed to complete EU bans for penta and octaBDE, and restric-tions on the use of decaBDE in addition to other restrictions within sev-eral jurisdictions on the manufacture and new use of the threecommercial PBDE formulations across the world (Harrad et al., 2010).Moreover, PBDEs associated with penta and octaBDE have been listedunder the UNEP Stockholm Convention on POPs, while DecaBDE is cur-rently under consideration for listing under Annexes A, B and/or C to theconvention (Stockholm convention on POPs, 2009). Despite such re-strictions, human exposure to PBDEs is likely to continue for the fore-seeable future, given their persistence and ubiquity of flame-retardedconsumer materials (Harrad and Diamond, 2006).

Several studies have reported different levels of PBDEs in varioushuman tissues including the serum, placenta, liver, adipose tissue andbreast milk from different European, Asian and North American coun-tries in the last few years (Cui et al., 2012; Frederiksen et al., 2009).These biomonitoring data provide a direct measurement of the humanbody burden of BFRs resulting fromvarious external exposure pathways(e.g. inhalation, ingestion of dust, diet and water) and contribute tothe risk assessment of such compounds. However, the only available

131M.A.-E. Abdallah, S. Harrad / Environment International 63 (2014) 130–136

information on BFRs in UK human samples is for tri- to hexa-BDEs(major components of the pentaBDE commercial product) where themedian concentrations for Σtri-to hexa-BDEs in human milk andserum samples collected in 2003were 6.3 and 4.18 ng g−1 lipid weight(lw) respectively (Kalantzi et al., 2004). In addition, BDE-209 wasdetected in 11 out of 153 serum samples at concentrations from 0.015to 0.240 ng g−1 lw (Thomas et al., 2006).

Current understanding is that non-occupational human exposure toPBDEs occurs mainly via a combination of diet, air and indoor dust(either via ingestion or dermal contact) (Frederiksen et al., 2009;Lorber, 2008; Trudel et al., 2011). However, the extent to which theknown contamination of indoor environments with PBDEs influenceshuman body burdens remains unclear. While some studies have man-aged to establish significant positive correlations between the levels ofPBDEs in food or indoor dust and their concentrations in human milkor serum (Dunn et al., 2010; Thomsen et al., 2008; Wu et al., 2007);such correlations could not be established in other studies (Roosenset al., 2009; Wang et al., 2013). An alternative approach involved appli-cation of a simple pharmacokinetic model to predict the body burdensof PBDEs in American adults using intake data from different exposurepathways. The predicted body burdens were then compared to the re-ported levels of PBDEs in humanmatrices and the relationship betweenexternal and internal exposure of American adults to PBDEs wasdiscussed (Lorber, 2008).

To address this paucity of UK human biomonitoring data for PBDEs,this study reports concentrations of Σtri-hexa BDEs and for the first timeBDE-209 in 35humanmilk samples fromBirmingham,UK. These data arethen used to estimate the dietary exposure of UK nursing infants underdifferent exposure scenarios. Finally, a simple, one-compartment phar-macokinetic model is applied to predict the body burdens of the studiedPBDEs in UK adults (using indoor air and dust levels reported elsewhereby our research group for Birmingham, UK) (Abdallah and Harrad,2010; Harrad and Abdallah, 2011; Harrad et al., 2006; Harrad et al.,2008a). The model predictions are then compared to the concentrationsof target compounds measured in the analyzed human milk samples(used as an indicator of adult female body burdens) for further under-standing of the relationship between external and internal human expo-sure to PBDEs in UK adults.

2. Materials and methods

2.1. Sample collection

Breast milk samples (each comprising ~50 mL) were obtained from35adult (aged 22–35), healthy, primiparous volunteers via BirminghamWomen's Hospital Milk Bank after the study protocol was approved byWarwickshire Research Ethics Committee and the R&D Department inBirmingham Women's NHS Foundation Trust. Informed consent wasobtained fromall the participants before sample collection. Samples col-lected in January–February 2010,were kept in clean screw-capped glasscontainers and transferred from theMilk Bank to the laboratory in spe-cial ice boxes then stored at −20 °C until the time of analysis. Due toethical regulations, the samples were collected in a completely anony-mous fashion with all participant information kept strictly confidential.For the purpose of this study, only 1 milk sample was collected fromeach mother during her first 3 months of lactation. Further details onthe sample collection protocol are provided in the Supplementary datasection.

2.2. Sample extraction

Accurately weighted aliquots of the freeze-dried samples (~2 g)were loaded into pre-cleaned 66 mL Accelerated Solvent Extraction(ASE 300, Dionex Inc., UK) cells containing 1.5 g florisil, 3 g alumina,5 g anhydrous Na2SO4 and hydromatrix (Varian Inc., UK) to fill thevoid volume of the cells, spiked with 25 ng of each of the 13C-labeled

BDE-47, BDE-99, BDE-153, BDE-183, and BDE-209 (Wellington Lab.,Guelph, ON, Canada) as internal (surrogate) standards. The ASE cellswere extracted with hexane:dichloromethane (1:9, v/v) at 90 °Cand 1500 psi. The heating time was 5 min, static time 4 min,purge time 90 s, and flush volume 50%, with three static cycles.The lipid weight of the studied samples was determined gravimet-rically on separate aliquots using a standard procedure (TheEuropean Standard EN 1528-2, 1996; see Supplementary data formore details).

2.3. Sample clean-up

The crude extracts were concentrated to 0.5 mL using a ZymarkTurbovap® II (Hopkinton, MA, USA) then washed with 3 mL of 98%sulfuric acid. After phase separation, the hexane layer was trans-ferred onto a florisil column topped with sodium sulfate and elutedwith 25 mL of hexane:dichloromethane (1:1, v/v). The eluate wasevaporated to dryness under a gentle stream of N2 and the driedextract reconstituted in 200 μL of 13C-BDE-100 (25 pg μL−1 inmethanol) used as a recovery determination (or syringe) standardto determine the recoveries of internal standards for QA/QCpurposes.

2.4. LC–APPI–MS/MS analysis

Sample analysis was carried out using an LC–MS/MS system com-posed of a dual pump Shimadzu LC-20AB Prominence liquid chromato-graph equipped with an SIL-20A autosampler, a DGU-20A3 vacuumdegasser coupled to a Sciex API 2000 triple quadrupole mass spectrom-eter. Details of the multi-residue analytical methodology used for sepa-ration and quantification of the studied PBDEs (28, 47, 85, 99, 100, 153,154, 183, 196, 197, 203, 206, 207, 208 and 209) can be found elsewhere(Abdallah et al., 2009) (a brief description is given in the Supplementarydata section).

2.5. Comparison of PBDE intake to human body burdens

We have previously estimated UK adult intake of the targetPBDEs via inhalation, dust ingestion and diet (Harrad andAbdallah, 2011; Harrad et al., 2006, 2008a,b) (a summary of the as-sumptions on which these estimations are based is provided asSupplementary data). To examine the relationship between theseestimated intakes and the body burdens indicated via human milksamples, a simple one-compartment, first order pharmacokinetic (PK)model was used. The studied PBDEs were hypothesized to accumulatein lipids (the single compartment in the model). Therefore, the changein PBDE lipid concentration over time can be expressed by Eq. (1)(Lorber, 2008).

δCPBDE

δt¼ IPBDE tð Þ � AFPBDE

BL tð Þ−KPBDE � CPBDE tð Þ ð1Þ

where CPBDE is the compound specific concentration in lipids (ng g−1 lw); IPBDE is the daily intake of the target BFR (ng day−1); AFPBDEis the absorption fraction (unitless); BL is body lipid mass (g) andKPBDE is the compound specific first order dissipation rate (day−1).

If KPBDE is assumed to be constant over time then Eq. (1) can besolved into:

CPBDE tð Þ ¼ CPBDE 0ð Þ � e −KPBDE�tð Þ þ IPBDE tð Þ � AFPBDEð ÞBL tð Þ

� �� 1−e −KPBDE�tð Þ� �

KPBDE

24

35ð2Þ

where CPBDE (0) is the studied PBDE body lipid concentration at time 0(initial concentration before intake).

Assuming a constant dose over time at constant body lipid mass, thesteady state PBDE lipid concentration can be calculated fromEq. (3). It is

132 M.A.-E. Abdallah, S. Harrad / Environment International 63 (2014) 130–136

stressed that the assumption of steady state conditions is an inherentuncertainty with this approach.

CPBDE ¼ IPBDE � AFPBDEð ÞBL � KPBDE

: ð3Þ

2.6. Quality assurance/quality control

Good recoveries (68–106%) of the 13C-labeled internal standardswere obtained for all the studied compounds (Table SI-4). Further eval-uation of themethod extraction/clean up performancewas achieved viaspiking milk samples (n = 6) with 13C-BDE-154 prior to freeze dryingand excellent recoveries (N90%) were obtained (Table SI-5).

No target compounds were detected in method blanks (n = 5;consisting of 2 g pre-extracted anhydrous sodium sulfate treated exact-ly as a sample) or field blanks (n = 5; consisting of ~2 g of brokenpieces of the glass milk containers treated exactly as a sample). There-fore, there was no need for blank correction of concentrations andmethod limits of detection (LOD) and quantification (LOQ) were esti-mated based on 3:1 and 10:1 S:N ratios respectively.

The accuracy and precision of the analytical method applied forPBDE determination were assessed via replicate analysis (n = 10) ofNIST SRM 2585. The results obtained compared favorably with thereported reference values (Table SI-6a).

3. Results and discussion

3.1. Concentrations of Σtri-hexaBDEs in UK human milk

While none of the investigated hepta- to nona-BDE congeners wasabove LOQ, BDE-47 was quantified in all the analyzed samples contrib-uting 34–74% to Σtri-hexa BDEs (Table 1). The predominant BDE conge-ners in the studied human milk were in the order BDE-47 N BDE-153 N BDE-99. These 3 congeners constituted an average of 85% of thequantified Σtri-hexa BDEs in the studied samples. This is in agreementwith previous reports of PBDEs in human milk from various countries(Frederiksen et al., 2009). Interestingly, a higher average level of BDE-153 (1.1 ng g−1 lw) than that of BDE-99 (0.71 ng g−1 lw) was ob-served (Table 1). While this differs from the relative contribution ofthese 2 PBDE congeners in the commercial pentaBDE formulations (LaGuardia et al., 2006), several authors have reported higher levels ofBDE-153 than BDE-99 in human milk (Ben Hassine et al., 2012; Dunnet al., 2010; Frederiksen et al., 2009). In addition, a recent study has re-ported BDE-153 as the dominant congener in 5 human breastmilk sam-ples from California (Park et al., 2011). Furthermore, a study of PBDEs inhumanmilk from the Faroe Islands also reported predominance of BDE-153 (Fangstrom et al., 2005). However, such high levels of BDE-153could not be associated with high consumption of seafood diet in thestudied population, indicating that dietary exposure was not the reasonfor the elevated BDE-153 concentrations in breast milk. Therefore, we

Table 1Statistical summary of PBDE concentrations (ng g−1 lw) in human milk samples (n = 35) fro

BDE-47 BDE-49 BDE-85 BDE-99

Average 3.30 b0.05 0.08 0.71SDa 3.25 0.08 0.15 0.67Median 2.80 b0.05 b0.05 0.69DFb (%) 100 20 46 94LOQ 0.043 0.045 0.051 0.055Minimum 0.17 b0.05 b0.05 b0.0625th %ile 0.78 b0.05 b0.05 0.2075th %ile 5.15 b0.05 0.09 0.85Maximum 14.65 0.45 0.83 3.43

a Standard deviation.b Detection frequency.c Not applicable.

hypothesize that the relatively higher contribution of BDE-153 to Σtri-hexa BDEs in human milk samples than expected from the pentaBDEtechnical mixture may be attributed to 2 main factors:

First, the high bioaccumulation potential of BDE-153 in lipids (as ev-idenced by a half-life of 6.5 years compared to 1.8 and 2.9 years forBDE-47 and BDE-99 respectively) (Geyer et al., 2004) which indi-cates that over time, BDE-153 will become the predominant conge-ner in the body.Second, the possible production of BDE-153 as a result of BDE-209metabolic stepwise meta–meta debromination (Roberts et al.,2011). This stepwise debrominationwas previously observed in per-egrine falcon eggs from California, where BDE-153 was the domi-nant congener only in eggs with high levels of BDE-209 (Holdenet al., 2009). Interestingly, while concentrations of BDE-153 in thisstudy were significantly (r = 0.443; p b 0.01) correlated withthose of BDE-209, no other statistically significant (p b 0.05) correla-tionwas observed between BDE-209 levels and any of the PBDE con-geners or Σtri-hexa BDEs in the analyzed samples. This furthersupports the hypothesis that metabolic degradation of BDE-209yields the highly bioaccumulative BDE-153 resulting in elevatedconcentrations of the latter in human milk.

While the levels of Σtri-hexa BDEs in this study (Table 1) are slightlylower than those reported in UK humanmilk samples collected in 2003(n = 54, average = 6.3 ng g−1 lw), these concentrations are still at thehigh end of those reported from other European, Asian, African andAustralasian countries (Table 2). On the other hand, Σtri-hexa BDEs inUK human milk are substantially lower than those reported from USAand Canada (Table 2)which is in agreementwith the farmore extensiveproduction and use of the pentaBDE technical formulation in NorthAmerica than elsewhere (BSEF, 2013).

3.2. Concentrations of BDE-209 in UK human milk

BDE-209 was above LOQ in 69% of the studied milk samples rangingfrom b0.06–0.92 ng g−1 lw (Table 1). To the authors' knowledge, thispaper is the first to report concentrations of BDE-209 in UK humanmilk. Interestingly, these levels are at the lower end of BDE-209 concen-trations reported in human milk from other European countries(Table 2) despite the substantially higher levels of this BFR reported inUK indoor dust compared to the rest of Europe (Harrad et al., 2010)and the reported higher usage of BDE 209 in theUK than other EU coun-tries (EU Risk Assessment Report, 2002). This may indicate that whileindoor dust ingestion is themajor pathway of external human exposureto BDE-209 (Harrad et al., 2008a; Lorber, 2008), the high levels of thiscompound in indoor dust do not significantly contribute to humanbody burdens. Our research group have recently reported on the verylow bioaccessibility (~14%) of BDE-209 in indoor dust across thehuman gastrointestinal tract (GIT) following oral ingestion (Abdallah

m Birmingham, UK.

BDE-100 BDE-153 BDE-154 ∑tri-hexa BDE-209

0.45 1.10 0.30 5.95 0.310.39 1.05 0.30 5.35 0.300.38 0.91 0.21 5.00 0.25

89 97 77 100 690.053 0.058 0.059 N/Ac 0.062

b0.05 b0.06 b0.06 0.2 b0.060.12 0.35 0.07 1.70 b0.060.70 1.43 0.55 9.55 0.581.86 4.57 11.10 26.10 0.92

Table 2Average concentrations of PBDEs (ng g−1 lw) in human milk samples from different countries.

Location Year Number ∑tri-hexa BDEs BDE-209 Reference

UK 2010 35 5.9 0.3 (This study)China 2011 103 0.8 1.8 Shi et al. (2013)Tunisia 2010 36 8.3 N/Aa Ben Hassine et al. (2012)New Zealand 2009–2010 33 4.7 0.4 Coakley et al. (2013)Norway 2003–09 393 2.7 0.6 Thomsen et al. (2010)India 2009 45 1.1 0.4 Devanathan et al. (2012)Korea 2008–09 21 2.7 N/A Kim et al. (2011)Taiwan 2007–2008 46 3.4 0.5 Chao et al. (2010)Australia 2007 10 7.6 0.3 Toms et al. (2009)Italy 2005–07 13 1.3 N/A Alivernini et al. (2011)Belgium 2006 22 3.0 5.9 Roosens et al. (2010)Sweden 1996–2006 276 3.4 N/A Lignell et al. (2011)France 2004–06 93 2.5 1.6 Antignac et al. (2009)North Carolina, USA 2004–2006 303 51 N/A Daniels et al. (2010)Spain 2005 9 2.1 2.5 Gomara et al. (2011)Canada 2003 10 50.4 0.4 She et al. (2007)Texas, USA 2002 47 34.0 0.9 Schecter et al. (2003)UK 2001–03 54 6.3 N/A Kalantzi et al. (2004)

a N/A not analyzed.

35

40

day

-1)

Σtri-hexa BDEs

BDE-209

133M.A.-E. Abdallah, S. Harrad / Environment International 63 (2014) 130–136

et al., 2012), consistentwith animal studies reporting lowbioavailability(4–26%) of BDE-209 (Huwe and Smith, 2007; Sandholm et al., 2003).Such poor uptake of BDE-209 from the GIT, combined with its veryshort human half-life (t0.5 = 7 days) (Geyer et al., 2004) and its prefer-ential partitioning to serum rather than milk fat (Mannetje et al., 2012)may result in the apparently low influence of BDE-209 concentrations inindoor dust on UK adult body burdens.

3.3. Nursing infants' dietary intake of PBDEs via breast milk

Breast milk is a recognized medium for direct transfer of POPs tonursing infants. To estimate the nursing infants' dietary intake of thestudied BFRs via breast milk, Eq. (4) was used.

Di ¼ CPBDE � FlipidBw

ð4Þ

where Di is the estimated dietary intake (ng kg−1 bw day−1); CPBDE isthe concentration of target PBDE in milk (ng g−1 lw); Flipid is the dailylipid intake via breast milk (g day−1) and Bw is the body weight(4.14 kg) (U.S. EPA, 2002.). The infant's daily lipid intake via breastmilk (Flipid) was calculated using U.S. EPA guidelines (U.S. EPA, 2002)which suggest an average intake of 702 mL milk per day for a1 month old infant weighing 4.14 kg. The median lipid content of theanalyzed milk samples was 3.47 g lipid per 100 mL of breast milkresulting in a daily lipid intake of 24.4 g lipid day−1.

Table 3 shows the estimated dietary intake of target PBDEs via breastmilk using different exposure scenarios (in which exposure factors (e.g.dust ingestion rate) were held constant but using different PBDE con-centrations (e.g. 25th percentile) derived from our breast milk data).

Table 3Estimated exposurea (ng (kg bw)−1 day−1) of a 1 month old infant to the target BFRs viabreast milk under different scenariosb.

25th %ile Average Median 75th %ile

BDE-47 4.6 19.3 16.3 30.3BDE-99 1.2 4.2 4.0 5.1BDE-100 0.7 2.7 2.2 4.2BDE-153 2.1 6.5 5.3 8.4BDE-154 0.4 1.7 1.3 3.2Σtri-hexa BDEs 10.0 34.9 29.4 56.4BDE-209 b0.1 1.8 1.2 3.4

a Values below LOQ were assumed to be 1/2 LOQ.b Based on an average bodyweight of 4.14 kg and a daily lipid intake of 24.4 g lipid day−1

(U.S. EPA, 2002).

While the estimated average UK infant exposure to Σtri-hexa BDEs ismuch lower than that in North America (Park et al., 2011), a 1 month-old infant in the UK is still more exposed to Σtri-hexa BDEs than in sev-eral other European countries via breast milk (Roosens et al., 2010).Interestingly, the average exposure of a nursing infant to Σtri-hexa BDEsvia breast milk exceeded upper-bound dietary intakes of UK adults andtoddlers (UK Food Standards Agency, 2006) (Fig. 1), while for BDE-209,dietary exposurewas themost significant exposure pathway for toddlers.

The low concentrations of BDE-209 in the studied milk sam-ples resulted in much lower exposure of UK nursing infants tothis contaminant than the USEPA reference daily dose (RfD) of7 μg kg bw−1 day−1. Similarly, our estimated UK infant daily in-takes (Table 3) are lower than the USEPA reference doses forBDE-47 (100 ng kg bw−1 day−1for neurodevelopmental toxicity)and Σtri-hexa BDEs (2000 ng kg bw−1 day−1for liver toxicity)(U.S.EPA, 2008). However, the median level of Σtri-hexa BDEs in thisstudy (4.98 ng kg−1 lw) is slightly higher than that associated with con-genital cryptorchidism (4.16 ng kg−1 lw; p b 0.01) in Danish–Finnishnewborn boys (Main et al., 2007) and generally in linewith levels associ-ated with irregular menstruation periods in a Taiwanese population(Chao et al., 2010). Another study from Taiwan showed a significant cor-relation between PBDE levels in breast milk (average = 3.93 ng g−1 lw)and decreased birth weight, length, chest circumference, and Quetelet'sindex of infants (Chao et al., 2007). Recently, an inverse association be-tween Σ4PBDEs (47, 99, 100 and 153) in human milk from Sweden(n = 364; average = 3.1 ng g−1 lw) and birth weight was observed in

0

5

10

15

20

25

30

Adult Toddler Breast-fed infant

Die

tary

exp

osur

e (n

g kg

-1 b

w

* From reference (UK Food Standards Agency, 2006); ** This study.

Fig. 1. Average estimates of dietary exposure (ng (kg bw)−1 day−1) of UK adults*,toddlers* and breast-fed infants** to PBDEs. *From UK Food Standards Agency (2006);** This study.

134 M.A.-E. Abdallah, S. Harrad / Environment International 63 (2014) 130–136

a multivariate model including gestational length (β = −106;p = 0.04). The observed association was stronger for boys than forgirls (Lignell et al., 2013). Potential mechanisms underlying the ob-served negative association between prenatal exposure to PBDEs andbirthweight are unknownbutmay involve effects on the endocrine sys-tem (Lignell et al., 2013). Both animal and epidemiological studies haveshown that PBDEs can alter thyroid hormone levels (Chevrier et al.,2010; Stapleton et al., 2011), which may influence fetal growth. Fur-thermore, two studies from Spain and Taiwan reported an associationof borderline statistical significance between PBDE levels in humanmilk and psychomotor development in infants after multivariate step-wise linear regression analyses (Chao et al., 2011; Gascon et al., 2012).The lack of significant direct correlation was attributed to unmeasuredBDE-209 metabolites, including hydroxylated PBDEs, which are moretoxic, more stable, and more likely to cross the placenta and to easilyreach the brain than BDE-209 (Gascon et al., 2012).

While this does not provide solid evidence on the potential healtheffects associated with the reported levels of PBDEs in human milkdue to the lack of relevant studies in the UK, our results certainly raiseconcerns about potential adverse effects resulting from exposure of in-fants and mothers to PBDEs. Although breastfeeding mothers shouldbe encouraged and supported due to the well-documented beneficialeffects of breast feeding, scientific studies ought to characterize andmeasure the contaminants in breast milk so that protective measuresmay be provided, if necessary, to avoid any potential harmful effectson the mother or the newborn.

3.4. Comparison of PBDEs intake to human body burdens

To convert daily adult intakes of BFRs via different exposure path-ways to expected body burdens, the bioaccessible fractions of each tar-get compound (Abdallah et al., 2012) were used in Eq. (3) to substitutefor AFPBDE in case of exposure via dust ingestion or diet, while theinhalable fraction was assumed to be 100% bioavailable. The body lipidmass was estimated based on a 25% body fat for an average adult

Table 4Comparison of predicted adult body burdens arising from average and median dailyexposures# to major target PBDEs with observed levels in human milk samples.

BDE-47 BDE-99 BDE-100 BDE-153 BDE-154 Σ5BDEs BDE-209

Average intake⁎ (ng day−1)Dusta 1.10 1.80 0.24 0.31 0.17 3.70 4270Dietb 35 30 5.60 7.00 2.80 80 310Airc 0.90 0.60 0.14 0.05 0.03 1.70 9.40

Median intake⁎ (ng day−1)Dusta 0.29 0.67 0.08 0.12 0.01 1.20 2975Dietb 35 30 5.60 7.00 2.80 80 310Airc 0.20 0.30 0.04 0.01 0.01 0.55 7.40

Average predicted body burdens (ng g−1 lw)Dust 0.06 0.05 0.01 0.02 0.01 0.14 0.34Diet 3.33 1.39 0.38 1.15 0.16 6.40 0.03Air 0.11 0.05 0.01 0.01 0.01 0.20 0.01Sum 3.49 1.49 0.40 1.19 0.18 6.74 0.38

Median predicted body burdens (ng g−1 lw)Dust 0.01 0.02 0.00 0.01 0.00 0.04 0.24Diet 3.33 1.44 0.38 1.15 0.16 6.45 0.03Air 0.03 0.03 0.00 0.00 0.00 0.06 0.00Sum 3.36 1.48 0.39 1.16 0.16 6.55 0.27

Observed body burdens (ng g−1 lw)Average 3.28 0.71 0.45 1.09 0.28 5.92 0.31Median 2.77 0.68 0.38 0.9 0.21 4.98 0.24

# Values below LOQ were assumed to be 1/2 LOQ.⁎ Based on average adult dust ingestion rate of 20 mg day−1 (Jones-Otazo et al., 2005),

average inhalation rate of 20 m3 day−1 (Currado and Harrad, 1998) and average adultweight of 70 kg.

a Estimated from Harrad et al. (2008a).b Estimated from UK Food Standards Agency (2006).c Estimated from Harrad et al. (2006) and Stapleton et al. (2009).

weighing 70 kg (U.S. EPA, 1997). Finally, KPBDE was calculated as0.693/t0.5; where t0.5 is the half-life of the studied BFR in the bodylipid compartment (Geyer et al., 2004).

In general, good agreement was observed between the predictedand the observed body burdens of main target PBDEs (Table 4) giventhe simplicity of the model used (e.g. only one body compartmentwas studied), the dearth of information regarding the half-lives of dif-ferent PBDE congeners in various compartments of the human body,and the uncertainty about the bioavailability of the studied compoundsfrom different exposure routes.

In addition, the PKmodel used here does not estimate human expo-sure via routes such as dermal contact and water intake. This is due tothe high uncertainty and complete absence of experimental data onthe extent of BFR absorption via dermal contact by humans coupledwith the expected minimal contribution of water intake to the overalldaily exposure to BFRs based on the very low aqueous solubility ofPBDEs.

Nevertheless, the good agreement between the predicted and ob-served results indicates that the studied exposure routes are the mainpathways driving UK adult body burdens of PBDEs. This is in line withthe findings of Lorber (2008) who studied the exposure of Americansto PBDEs and reported indoor dust ingestion as themain route of expo-sure followed by diet and inhalation. However, more research is re-quired for assessment of the bioavailability of various PBDEs viadifferent exposure routes and determination of t0.5 of PBDEs in varioushuman tissues.

Acknowledgments

The authors thank all the milk donors and the staff of BirminghamWomen's Hospital Milk Bank (Heather Barrow, Jenny Harris and AnneHemming). We also thank Kelly Hard (R & D manager at BirminghamWomen's Hospital) for helping with the ethical issues for this project.

Appendix A. Supplementary data

Specific details of analytical methodology, exposure estimation, QA/QCmeasurements and concentrations of target BFRs in each sample areavailable as Supplementary data. Supplementary data to this article canbe found online at http://dx.doi.org/10.1016/j.envint.2013.11.009.

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