Environment International 64 (2014) 61–68

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Polybrominated diphenyl ethers (PBDEs), hexabromocyclododecane (HBCD) and “novel” brominated flame retardants in house dust in Germany H. Fromme ⁎, B. Hilger, E. Kopp, M. Miserok, W. Völkel Bavarian Health and Food Safety Authority, Department of Chemical Safety and Toxicology, Pfarrstrasse 3, D-80538 Munich, Germany

a r t i c l e

i n f o

Article history: Received 5 July 2013 Accepted 25 November 2013 Available online 22 December 2013 Keywords: PBDE HBCD BFR Dust Intake Exposure

a b s t r a c t Brominated flame retardants (BFRs) are used in a wide variety of products such as electronic devices, upholstery and carpets and in insulation boards. The study presented here aimed to quantify the amounts of BFRs in house dust in Germany. For this purpose 20 residences' dust samples were collected from vacuum cleaner bags and analysed with LC–MS/MS and simultaneously with GC/MS. Using GC/MS, the median (95th percentile) concentrations of PBDEs (sum of tetra- to hepta-congeners), BDE 209, Σ-HBCD (sum of three congeners), and decabromodiphenylethane (DBDPE) were 42 ng/g (230 ng/g), 950 ng/g (3426 ng/g), 335 ng/g (1545 ng/g), and 146 ng/g (1059 ng/g), respectively. Using LC–MS/MS some “novel” flame retardants were found in median concentrations of 343 ng/g (bis(2-ethyl-1-hexyl)tetrabromophthalate, TBPH), and 28 ng/g (tetrabromobisphenol A, TBBPA). Whilst 1,2-bis-(2,4,6-tribromophenoxy)ethane (BTBPE) and 2-ethyl-1-hexyl-2,3,4,5-tetrabromobenzoate (EH-TBB) could not be detected. Based on these measurements an exposure assessment for the sum of tetra- to heptabrominated congeners, BDE 209, and Σ-HBCD resulted in a “high” daily intake for toddlers (based on 95th percentiles) of 1.2 ng/kg b.w., 0.69 ng/kg b.w., and 8.9 ng/kg b.w., respectively. For TBPH the “high” intake was calculated at 4.1 ng/kg b.w. and for DBDPE at 5.3 ng/kg b.w. A clear tendency was observed to apply “novel” BFRs in Germany. Moreover, the results suggest that the recent exposure to PBDEs and HBCD via house dust in Germany is well below the levels that are associated with health effects. For the “novel” brominated flame retardants such an assessment is not possible due to limited toxicological information. © 2013 Elsevier Ltd. All rights reserved.

1. Introduction Nowadays, mainly four different groups of chemicals are used in large quantities to prevent fire hazards by raising the ignition temperature of the polymer, reducing the rate of burning, reducing flame spread or reducing smoke generation (Morose, 2006). Besides inorganic salts (e.g. antimony trioxide, aluminium hydroxide, borat), phosphorus compounds (e.g. organophosphates, halophosphates, phosphine oxides, and red phosphorus), nitrogen-based compounds (e.g. melamine and melamine derivatives) and halogenated substances such as halogenated paraffins, chlorinated alicyclic compounds, and brominated aromatic compounds are used. The group of brominated flame retardants (BFRs) consists of diverse chemicals, such as polybrominated diphenyl ethers (PBDEs), a class of substances with theoretically 209 individual congeners depending on their location and number of bromine atoms, hexabromocyclododecane (HBCD), tetrabromobisphenol A (TBBPA), 1,2-bis-(2,4,6-tribromophenoxy)ethane (BTBPE), 2-ethyl1-hexyl-2,3,4,5-tetrabromobenzoate (EH-TBB), and bis(2-ethyl-1-hexyl) ⁎ Corresponding author. E-mail address: [email protected] (H. Fromme). 0160-4120/$ – see front matter © 2013 Elsevier Ltd. All rights reserved. http://dx.doi.org/10.1016/j.envint.2013.11.017

tetrabromophthalate (TBPH). Whilst PBDEs are used as additive flame retardants in electronic devices such as computers and TV sets as well as in upholstery and carpets, HBCD is primarily used in insulation boards (e.g. polystyrene foam) and textiles, and TBBPA as a reactive flame retardant in electrical equipment (e.g. printed circuit boards) (de Wit et al., 2006; EFSA, 2011a,b; NICNAS, 2001). Only limited information is available on the current global market volume, but approximately 311,000 metric tonnes of BFRs were used worldwide in 2005, which is about 21% of the total consumption of flame retardants (Harju et al., 2008). Since August 2004 the use and import of products containing more than 0.1% of technical penta-, and octabromodiphenyl ethers have been banned in the European Union (RoHS, 2003). Following a decision made by the European Court of Justice these restrictions are now also valid for deca-BDE. Moreover, HBCD was listed in annex XIV of REACH as substances subject to authorisation and the restrictions on pentaand octa-BDE were provided in annex XVII (EC, 2011). In the US, some states have banned penta- and octa- forms of PBDE and the producers have decided to phase out the production and use of deca-BDE voluntarily by the end of 2012 (State of California, 2003). In addition to these shifts in the market, there has been an increase in the use of nonPBDE BFR as an alternative to the discontinued PBDEs.

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A number of studies have revealed that PBDEs and some non-PBDE brominated BFRs like HBCD are persistent, bioaccumulative and globally distributed in the environment and have been found in different environmental media and in biota (Covaci et al., 2006; Darnerud et al., 2001; Frederiksen et al., 2009; Kolic et al., 2009; Law et al., 2008; Marvin et al., 2011; Raab et al., 2008; Rice et al., 2002; Sjödin et al., 2008a; Stapleton et al., 2008; Toms et al., 2009). Additionally, “novel” BFRs such as EH-TBB and TBPH have been found under occupational conditions in electrical and electronic waste recycling sites [e.g.(Ali et al., 2011a; Rosenberg et al., 2011)] as well in atmospheric and marine environmental media from Asia up to remote areas like the Arctic (Möller et al., 2011). Nevertheless, toxicological knowledge is still limited to some PBDE and HBCD congeners (EFSA, 2011a,b; HC, 2006; NICNAS, 2001). Overall, the main targets are the liver, thyroid hormone homeostasis and the reproductive and nervous systems, and effects on neurodevelopment have been identified as critical endpoints. For the other “novel” BFRs toxicological data are very limited at present. There are some studies available showing that dietary intake is a main source of human exposure to PBDEs (Bakker et al., 2008; Fromme et al., 2009; Gómara et al., 2006; Harrad et al., 2004; Knutsen et al., 2008) and HBCD (Goscinny et al., 2011; Knutsen et al., 2008; Roosens et al., 2009; Törnkvist et al., 2011). Additionally, non-dietary human exposure via inhalation and oral ingestion of house dust can make a significant contribution to the total intake under some circumstances, particularly for toddlers (Abdallah et al., 2008a,b; D'Hollander et al., 2010; Fromme et al., 2009; Harrad et al., 2010; Lorber, 2008; Roosens et al., 2010). House dust originates from a number of sources and is a sink for different organic compounds, varies substantially in its chemical and biological compositions and is a heterogeneous material. Therefore, the analysis of chemicals in this matrix is a good indicator of the contamination occurring over a long period of time (Butte and Heinzow, 2002). Nevertheless, the contribution of house dust to the total exposure of humans is associated with a number of uncertainties resulting from the different sampling methods and sample preparation techniques as well as from the models used for quantifying the amount of dust ingested. For example Björklund et al. (2012) compared two sampling methods using vacuum cleaner bags or a researcher-collected method. They found a good correlation for ΣOctaBDE and ΣDecaBDE, but not for ΣPentaBDE or HBCD. In addition house dust is a complex matrix since composition may vary in a wide range. This and some drawbacks regarding sample work up and degradation of compounds as previously discussed by e.g. Kopp et al. (2012) complicate the analysis of BFRs in house dust samples. Therefore one aim of the study was the application of two different analytical methods namely a common GC–MS-method well-known for the detection of BFRs, especially PBDE, and a more innovative one based on LC–MS/MS with a minimized sample work up. A second goal was the determination of some BFRs in house dust to get an indicator of the use of these chemicals in Germany and to assess the daily nondietary exposure to these substances via the ingestion of dust. An overview of the target parameters is given in Table 1.

2. Material and methods The study was carried out in the city of Munich and in nearby suburban and rural areas of southern Germany. In 20 residences house dust was taken from the bags of the vacuum cleaners regularly used for cleaning the residences. The samples were collected in accordance with the German guideline VDI 4300-8 (VDI, 2001), sieved b 63 μm using a Vibratory Sieve Shaker AS 200 (Retsch GmbH, Haan, Germany) and stored frozen until analysis.

Table 1 Brominated flame retardants measured in this study. Substance

Abbreviation

CAS-No.

Tetrabromobisphenol A Hexabromocyclododecanes (α, β, γ)

TBBPA HBCD

2,4,4′-Tribromodiphenyl ether 2,2′,4,4′-Tetrabromodiphenyl ether 2,2′,4,4′,5-Pentabromodiphenyl ether 2,2′,4,4′,6-Pentabromodiphenyl ether 2,2′,4,4′,5,5′-Hexabromodiphenyl ether 2,2′,4,4′,5,6′-Hexabromodiphenyl ether 2,2′,3,4,4′,5′,6-Heptabromodiphenyl ether Decabromodiphenyl ether 2,2′,4,4′,5,5′-Hexabromobiphenyl 1,2-Bis(2,4,6-tribromphenoxy)ethane 2-Ethyl-1-hexyl-2,3,4,5-tetrabromobenzoate Bis(2-ethyl)tetrabromophthalate Decabromodiphenylethane

BDE 28 BDE 47 BDE 99 BDE 100 BDE 153 BDE 154 BDE 183 BDE 209 BB 153 BTBPE EH-TBB TBPH DBDPE

79-94-7 3194-55-6 25637-99-4 41318-75-6 5436-43-1 60328-60-9 189084-64-8 68631-49-2 207122-15-4 207122-16-5 1163-19-5 36355-01-8 37853-59-1 183658-27-7 26040-51-7 84852-53-9

2.1. Chemicals and standard solutions The following standards were purchased from Campro (Campro Scientific GmbH, Berlin, Germany): TBBPA (50 μg/ml in methanol), 13 C 12 -TBBPA (50 μg/ml in methanol), α-HBCD (50 μg/ml in toluene), 13C12-α-HBCD (50 μg/ml in toluene), β-HBCD (50 μg/ml in toluene), 13C12-β-HBCD (50 μg/ml in toluene), γ-HBCD (50 μg/ml in toluene), 13C12-γ-HBCD (50 μg/ml in toluene), BTBPE (50 μg/ml in nonane), 13C12-BTBPE (50 μg/ml in nonane), EH-TBB (50 μg/ml in toluene), TBPH (50 μg/ml in toluene), BDE-28 (50 μg/ml in nonane), 13 C 12 -BDE-28 (50 μg/ml in nonane), BDE-47 (50 μg/ml in nonane), 13C12-BDE-47 (50 μg/ml in nonane), BDE-99 (50 μg/ml in nonane), 13C12-BDE-99 (50 μg/ml in nonane), BDE-100 (50 μg/ml in nonane), 13C12-BDE-100 (50 μg/ml in nonane), BDE-153 (50 μg/ml in nonane), 13C12-BDE-153 (50 μg/ml in nonane), BDE-154 (50 μg/ml in nonane), 13C12-BDE-154 (50 μg/ml in nonane), BDE-183 (50 μg/ml in nonane), 13C12-BDE-183 (50 μg/ml in nonane), BDE-209 (50 μg/ml in nonane), 13C12-BDE-209 (25 μg/ml in nonane), BB-153 (50 μg/ml in nonane) and 13C-BB-153 (50 μg/ml in nonane). All solvents were purchased in HPLC quality from Promochem (Promochem LCG Standards GmbH, Wesel, Germany). Chemicals were acquired in the highest purity available from Sigma-Aldrich (Sigma-Aldrich Chemie GmbH, Taufkirchen, Germany). 2.2. LC–MS/MS method The APCI–LC–MS/MS method was previously described in detail by Kopp et al. (2012). In brief, after addition of 13C-internal standards (500 ng/ml) to each sample containing 100 mg of dust diluted in a solvent mixture (2 ml) of 1 part methanol, 1 part acetonitrile and 2 parts isopropanol for extraction the samples were extracted in an ultra sonic bath at 60 °C for 30 min. After centrifugation the supernatant was filtered with OPTI-Flow syringe filters (regener. cellulose, 0.45 μm, 13 mm; WICOM Germany GmbH, Heppenheim, Germany) and 100 μl were injected into an integrated on-line clean-up device with a trap column (Pursuit XRs 5, C18, 5 μm, 30 × 2.0 mm; Varian, Darmstadt, Germany). The different BFRs were separated on an analytical column (Pursuit XRs 5, C18, 5 μm, 150 × 2.0 mm) by gradient elution applying an ammonium acetate buffer, 2 mM and methanol. Target analytes were then detected using a Triple-Quad mass spectrometer (Q-Trap 5500, AB Sciex Germany GmbH, Darmstadt, Germany) equipped with an atmospheric pressure chemical ionization probe (APCI). Measurements were conducted in the multiple reaction monitoring (MRM) mode with negative ionisation. The limits of quantification (LOQ), defined as a signal to noise ratio N10, were as follows: 0.6 ng/g for TBBPA, 1.0 ng/g for each HBCD congener, 1.5 ng/g for TBPH, 3.0 ng/g for EH-TBB, 1.0 ng/g for BDE 99, 10 ng/g for BB 153, 2 ng/g for BDE

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100, −153, −154, −183, 10 ng/g for BTBPE, BDE 209, 80 ng/g for BDE 28, and 80 ng/g for BDE 47. More details are discussed in chapter 2.6 of Kopp et al., 2012. 2.3. GC/MS method For the analysis approx. 1 g of dust was taken and after addition of C12-BDE-28, − 47, − 99, − 100, − 153, − 154, − 183, and − 209, 13 C14-DBDPE and 13C12-γ-HBCD, were extracted by means of PLE (pressurized liquid extraction). The extraction cell was filled with 4 g Na2SO4, 1 g dust mixed with 2 g Na2SO4 and 4 g activated silica gel, 2 g Na2SO4 and quartz sand. The extraction was conducted using n-heptane/acetone (1:1) (v/v) at 100 °C, 100 bar and with 3 static cycles. The raw extract was reduced to approx. 1 ml and cleaned up by means of column chromatography using 7 g of neutral silica gel, covered with 20 g of sulphuric acid-treated silica gel (44% by weight) (column diameter 2 cm). The column was conditioned using 70 ml of n-heptane. The PBDEs and DBDPE (F1) were eluted using 260 ml of n-heptane, the HBCD (F2) was eluted using 130 ml of n-heptane/dichloromethane (1:1) (v:v). Both of the fractions were reduced to 1 ml in the parallel evaporator. The PBDE fraction (F1) was cleaned up using 5 g of basic aluminium oxide covered by 2 g of Na2SO4. The column was conditioned using 20 ml of n-hexane, the pre-elution was conducted using 50 ml of nhexane/dichloromethane (98:2) (v/v), and the PBDEs and DBDPE were eluted using 50 ml of n-hexane/dichloromethane (1:1) (v/v). The solvent was removed in the parallel evaporator and the residue was absorbed in 50 μl of toluene. Analogously the HBCD fraction (F2) was cleaned up using 4 g of activated Florisil® covered with 2 g Na2SO4 (column diameter 1 cm). The column was conditioned using 20 ml of n-hexane and interfering matrix components were separated by means of pre-elution using 25 ml of n-pentane, 15 ml of n-hexane/toluene (9:1) (v:v) and 10 ml of n-hexane. The HBCD was eluted using 50 ml of n-hexane/ dichloromethane (1:1) (v/v). The solvent was removed in the parallel evaporator and the residue was absorbed in 50 μl of toluene. The cleaned extracts were analysed using a GC/MS-QP2010plus (Shimadzu Corporation, Duisburg, Germany). The GC was equipped with a PTV injector and a 15 m DB5-MS capillary column (0.25 mm inner diameter, 0.1 μm film thickness; Agilent J&W Scientific, Waldbronn, Germany). Helium was used as the carrier gas and the ionisation was achieved by means of electron impact with 20 eV of ionisation energy and a 20 μA emission current at ion-source and interface temperatures of 250 °C and 280 °C, respectively. After the injection the PTV injector was heated to 325 °C at a rate of 300 °C/min and maintained at this temperature. The column oven was maintained at 110 °C with an increased inlet-pressure of 4 bar for 2 min, then with a constant column flow of 1.2 ml/min it was heated to 180 °C at a rate of 30 °C/min, to 260 °C at a rate of 20 °C/min and to 325 °C at a rate of 10 °C/min and was maintained at this temperature for 6 min. The analytes were detected by means of selected ion monitoring (SIM) and by testing the isotope ratios of the isotope pattern of molecule-specific fragments created as a result of bromine substitution (tolerance range ± 15%) (see Table S 1). LOQs of the analytical method were 0.1 ng/g for BDE 28, 1.0 ng/g for BDE 47 and BDE 99, 2 ng/g for tetra- to deca-BDE, and 0.05 ng/g for DBDPE. 13

2.4. Calculation of daily intake Taking into account the concentrations ascertained in house dust, the intake of brominated flame retardants is to be estimated for this route of exposure. For the PBDEs an exposure dose was estimated taking into account the specific resorption rates ascertained for each congener in animal experiments (bioavailability). Huwe et al. (2008) estimated the absorption rates for PBDEs in house dust following oral ingestion. The rates are

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33% (BDE 28), 69% (BDE 47), 44% (BDE 99), 78% (BDE 100), 73% (BDE 153), 19% (BDE 154), 48% (BDE 183), and 4% (BDE 209). For HBCD there is evidence of extensive absorption following oral ingestion, with certain differences between the isomers (EFSA, 2011b). As insufficient information about bioavailability is available for the other brominated flame retardants, an absorption of 100% after oral ingestion is assumed in these cases and for HBCD. In principle two scenarios were examined, an “average” intake on the basis of median concentrations in house dust, and a “high” intake on the basis of 95th percentiles. In addition, the exposure was determined for the group of adults and that of toddlers, as it is well-known that the latter group is able to ingest larger amounts of house dust by hand-to-mouth contact. Various assumptions have to be made for the calculation. For instance, it is assumed that the toddler has a body weight of 12 kg and that the adult has a body weight of 70 kg (EFSA, 2012). The average daily intake of house dust is assumed to be 30 mg for the adult and 60 mg for the toddler (US-EPA, 2011). 2.5. Statistical analysis Data were analysed using the statistical software package SPSS 13.0. Unless otherwise stated, all values below the quantification limit were set half the quantification limit for statistical analysis. Correlations were evaluated with the Spearman rank correlation coefficient. 3. Results 3.1. Analysis of the PBDEs and HBCD The statistical parameters regarding the LC–MS/MS and GC/MS measurements of the PBDEs in house dust are given in Table 2. As a result of the higher limits of detection it was not possible to detect BDE 28 and BDE 47 using the LC–MS/MS method. With this method BDE 209 and BDE 99 could be determined in 100% and 95% of the samples respectively, the other congeners in only 25 to 40% of the samples. Using the GC/MS method all analytes could be detected in the house dust samples, whilst BDE 28 was found only in 55% of the samples. As is also visible from the

Table 2 BFR-concentrations detected in house dust measured with APCI–LC–MS/MS and (GC/MS) in 20 dust samples in ng/g. N N LOQ

Mean

Median

95th percentile

Range

Polybrominated diphenyl ether BDE 28 0 (11) (0.2) BDE 47 0 (20) (11.7) BDE 99 19 (20) 21.7 (19.6) BDE 100 7 (20) 3.5 (3.3) BDE 153 8 (20) 4.6 (4.5) BDE 154 5 (20) 2.5 (2.1) BDE 183 7 (20) 27.9 (90.7) a Σ-BDE (132) BDE 209 20 (20) 386 (1233)

(0.1) (5.7) 9.4 (9.2) b2.0 (1.6) b2.0 (2.1) b2.0 (1.1) b2.0 (9.3) (42) 419 (950)

(0.5) (42.8) 70.1 (73.6) 11.4 (12.3) 17.8 (12.3) 9.2 (6.7) 129 (152) (230) 845 (3426)

(b0.1–1.0) (1.3–52.3) b1.0–84.1 (1.7–79.7) b2.0–15.8 (0.3–12.4) b2.0–20.5 (0.2–25.6) b2.0–9.5 (0.1–6.8) b2.0–394 (0.7–1494) (6–1546) 11–929 (10–3748)

Hexabromocyclododecane α-HBCD 20 293 β-HBCD 20 72 γ-HBCD 20 255 Σ-HBCD 20 (20) 620 (516)

180 35 114 345 (335)

986 238 486 1782 (1545)

32–1063 8–496 14–2563 53–4041 (12–2740)

28.0 b10 b3.0 343 (146)

105 26 12.5 811 (1059)

2.9–233 b10–34 b3.0–13.6 25–2274 (47–1570)

Other brominated flame retardants TBBPA 20 44.1 BTBPE 7 10 EH-TBB 8 4.2 TBPH 20 436 DBDPE – (17+) (323) BB 153 0

Concentrations b LOQ were set =0.5 ∗ LOQ. +of 18 samples. a Sum of tetra to hepta brominated congeners.

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aforementioned table, median levels in the lower ng/g range were found for the all congeners. Some higher concentrations in the dust were found only for BDE 183. We observed a strong dominance of BDE 209 in the house dust samples. Using the GC/MS method the total concentration of the 6 tetra- to heptabrominated PBDE congeners (i.e. excluding BDE 209) was in the range of 6 ng/g to 1546 ng/g, with a median of 20 ng/g. Using LC–MS/MS all of the HBCD congeners could be detected in all samples tested. The highest median levels were found for the α-isomer, at 180 ng/g, followed by 114 ng/g for the γ-isomer and 35 ng/g for β-HBCD. Comparable concentrations were found for the sum of the three HBCDs: a median of 345 ng/g using the LC–MS/MS method and 335 ng/g using GC/MS with a regression coefficient of R2 = 0.97 and a slope of 0.68 which shows systematically lower levels measured by the GC–MS method. In part this could be explained with different response factors (rf) of the diastereomers as previously described by Abdallah et al. since quantitation was performed with the γdiastereomer which shows a lower response (rf of about 0.7) compared to the α-diastereomer which was dominantly found in the dust samples as shown by LC–MS/MS. The two methods applied were compared by calculating correlations, though taking into account only the results that were above the LOQ. In the statistical analysis significant correlations were found between the levels for the sum of the HBCD (r: 0.983, p b 0.001), BDE 99 (r: 0.973, p b 0.001), BDE 100 (r: 0.949, p = 0.001), BDE 153 (r: 0.891, p = 0.003), BDE 183 (r: 0.970, p b 0.001), and BDE 209 (r: 0.879, p b 0.001). Only for BDE 154 no statistically significant correlation could be found. In Figs. S 1 to S 6 results of the linear regression analysis were given. Nevertheless, in contrast to BDE 99 (R2 = 0.95) for BDE 209 a regression of R 2 = 0.77 was found if the results of GC/MS and LC–MS/MS were compared. The slope of the regression line was 1.0 for BDE 99 and BDE 100, 0.8 for BDE 153, but only 0.3 and 0.2 and for BDE 183 and 209, respectively. It has to be considered that the number of measurements above LOQ using LC–MS/MS was only 7 for BDE 100, 8 for BDE 153, and 7 for BDE 183. With regard to Σ-HBCD the slope of the regression equation was 1.4 with an intercept of −140. In addition, a significant correlation was found between the sum of HBCD and BDE 209 in the dust samples (r: 0.531, p = 0.016), but not between HBCD and the other PBDE congeners. For the assessment of both methods applied the SRM 2585 dust sample was analysed by LC–MS/MS. The data are presented in Table S 1 and results for BDEs vary from 94% for BDE 28 to 111% for BDE 183. BDE 47 was quantitated with 495 ng/g which corresponds to 100% to the reported value of 497 ng/g. Some other BFRs such as HBCD were quantitated in other studies (see Tables S 2a,b). Using the GC/MS method recoveries from the reference material varied from 9 to 27% for the different PBDE congeners. 3.2. Analysis of the “novel” brominated flame retardants The results of the measurements of the other (“novel”) brominated flame retardants are given in Table 2. Using LC/MS–MS it was possible to detect BTBPE and EH-TBB in 35% and 40% of the samples, respectively; hexabromobiphenyl (BB 153) could not be detected in any sample above the LOQ of 10 ng/g. TBBPA and TBPH could be detected in all samples with median values of 28 ng/g and 343 ng/g, respectively. Using the GC/MS method it was also possible to detect DBDPE as a significant flame retardant in the house dust samples with median levels of 146 ng/g (range: 47 ng/g–1570 ng/g). 3.3. Daily intake of adults and toddlers Given the aforementioned assumptions regarding the ingestion of dust, the concentrations measured in the house dust in the residences lead to the daily intake levels given in Table 3. It becomes clear that the highest intake via this route of exposure must be expected in toddlers. The intake of ΣHBCD stands at 0.76 ng/kg b.w. (adult) and

8.9 ng/kg b.w. (toddler) in the “high” intake scenario, thus exceeding the intake of the PBDEs substantially. Despite high levels in house dust, the intake of BDE 209 is only 0.06 ng/kg b.w. (adult) and 0.69 ng/kg b.w. (toddler) in the aforementioned scenario owing to the very low absorption in the gastro-intestinal tract. In the case of the “novel” brominated flame retardants, an intake via this route must be expected in particular for TBPH and DBDPE. For toddlers the TBPH and DBDPE “high” intake is 4.1 ng/kg b.w. and 5.3 ng/kg b.w., respectively. 4. Discussion As described before the aims of the presented study were to get concentrations for some BFRs especially “novel” ones in house dust samples and to compare the more common GC–MS method with a LC–MS/MS method with reduced sample work up as previously published (Kopp et al., 2012). For some BFRs really good correlations with R2 N 0.9 and slopes near one could be observed in the presented study. For some BFRs such as BDE 209 only poor values were obtained. As previously published by Sahlström et al. (2012) a comparable discrepancy for BDE 209 and other BFRs may occur if different methods were used. At the moment we have no reasonable explanation for the differences observed in our study but we believe that is important to show such data to activate discussion and additional studies to find solutions and explanations for the observed discrepancy. For BDEs the results analysing the SRM 2585 with both methods presented here compared to NIST original data and two other working groups are presented in Table S 2a. Based on the results from the analysis of the SRM 2585 by LC–MS/MS (Tables S 2a,b) we believe that the LC–MS/MS method provides reliable data especially for BDEs and in addition to the reduced sample work up the LC–MS/MS-method is the method of choice for a fast measurement of selected BFRs in dust samples. Nevertheless GC–MS based methods not only provide slightly higher deviations but are also reliable in the same manner as shown in Table S 2a. The presented GC–MS method provides poor recoveries but the use of stable isotope labelled internal standard compensates the poor recovery rates resulting in deviations from certified concentrations of SRM 2585 from 76 to 116%. It is obvious that the application of such standards is necessary for quantitation of BFRs in house dust samples but it does not explain the different concentrations observed for some BFRs in house dust samples using different methods. Therefore it is necessary to develop, compare and standardize methods from sampling to analysis for quantitation of compounds in complex matrices such as house dust as recently reviewed by Mercier et al. (2011). On the whole, our study shows a low level of exposure for the tri- to heptabrominated diphenyl ethers. Higher levels were only observed in house dust for BDE 209, HBCD and the “novel” brominated flame Table 3 Calculation of the daily intake via dust consumption for toddlers and adults in pg/kg body weight (“average” intake on the basis of median concentrations in house dust, and a “high” intake on the basis of 95th percentiles). Adult

Σ-BDEb,c BDE 209b Σ-HBCDe TBBPAe BTBPEe EH-TBBe TBPHe DBDPEb a b c d e

Toddler

Reference dosea

“Average” intake

“High” intake

“Average” intake

“High” intake

18 16 148 12 2 1 147 63

99 59 764 45 11 5 348 454

210 190 1725 140 25 8 1715 730

1150 685 8910 525 130 63 4055 5295

100,000d 7,000,000 – – – – – –

Derived by US-EPA. Calculated from GC/MS data. Sum of tetra to hepta brominated congeners; 100% absorption from the intestine. Assumed that all congeners have a reference dose of 100 ng/kg b.w. Calculated from LC/MS data.

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retardants, TBPH and DBDPE. This indicates that the bans on penta- and octa-PBDEs, which have been in place for a long time, are effective and that substitution processes for these brominated compounds have begun. When comparing the LC–MS/MS and GC/MS methods we were able to show good levels of correlation in the measurement results on the whole, though the GC/MS procedure is advantageous for the PBDEs owing to its greater sensitivity. Nevertheless, it has to be taken into account that the GC/MS method produced higher levels of BDE 209 than the LC–MS/MS method.

studies conducted in Germany. For instance, median concentrations of BDE 209 ranging between 63 ng/g and 704 ng/g were found by Knoth et al. (2003), by Sjödin et al. (2008a), and by Gabrio et al. (2008). In contrast, considerably higher median concentrations were found in the UK (Harrad et al., 2008; Pless-Mulloli et al., 2006; Sjödin et al., 2008b). One possible explanation for this is Great Britain's apparently large share of the consumption of BDE 209 in the European Union (Harrad et al., 2008). The amount of BDE 209 as a percentage of the total concentration of all PBDEs is approx. 20% to 63% in North America, approx. 71% to 90% in Sweden and Germany, but 98% in Great Britain.

4.1. PBDEs in dust

4.2. HBCD in dust

A variety of studies have been conducted throughout the world examining house dust for the presence of the various PBDE congeners. In so far as results were reported on the individual congeners, selected study findings are given in Fig. 1. Beside differences in the dust levels due to different use patterns, time trends have to be considered. Regarding this, our results (GC–MS data) correspond very well with those of current measurements in Germany (Fromme et al., 2009; Sjödin et al., 2008b), whereas an older study still found a level of exposure for the tetra- to hepta-brominated PBDEs which was higher by a factor of 2.8 (Knoth et al., 2003). In Europe, higher concentrations than those found in our study were measured in dust of UK homes (Harrad et al., 2008; Pless-Mulloli et al., 2006; Sjödin et al., 2008b). The highest concentrations in Europe were reported from Sweden by Karlsson et al. (2007). In this study 5 homes were examined and very high concentrations of BDE 99 in particular were found. This congener accounted for approx. 60% of the total concentrations of the tetra- to heptabrominated congeners in the dust. Low concentrations were observed in Romania (Dirtu et al., 2010). The very high concentrations of PBDEs in the dust in North American indoor environments are particularly striking. Here the highest concentrations were found in 12 private homes in Michigan (Batterman et al., 2009) and in 49 residences in California (Zota et al., 2008). This can be explained in the light of the different fire protection regulations and the associated very different amounts of the PBDEs used. Furthermore, a correlation was observed with the presence of computers or new carpets in the rooms (Schecter et al., 2005). For BDE 209, too, the concentrations measured by us with GC/MS are in the range that was observed in the same region analysed previously (Fromme et al., 2009). This also applies to other

Studies which have already reported results on HBCD in dust are listed in Table 4. Abb et al. (2011), for example, who only determined γ-HBCD in 24 dust samples from Germany and 2 from the USA, found concentrations that are comparable with our study. No other findings from Germany have been published so far. Studies from the UK (Abdallah et al., 2008a), Canada (Abdallah et al., 2008a), and the USA (Abdallah et al., 2008a; Stapleton et al., 2008) also revealed a similar concentration range with median concentrations for the sum of the HBCDs of 354 to 730 ng/g. In contrast, considerably higher concentrations of 1300 ng/g (median) were described by Abdallah et al. (2008b), who conducted their study in 45 homes in the West Midlands and in Basingstoke, UK, in 2006/2007. In this study a maximum value of 140,000 ng/g was measured in one house. The authors provided no explanation for this exceptional value. Somewhat lower HBCD concentrations were reported by D'Hollander et al. (2010), who examined 43 houses in Flanders, and by Roosens et al. (2009), who investigated 12 university students' apartments in Belgium (median values: 13 ng/g and 114 ng/g respectively). Takigami et al. (2009) observed very different HBCD concentrations, 13,000 ng/g and 140 ng/g, in two wooden houses in Hokkaido, Japan. Depending on the commercial product utilised and the specific processing conditions, a different diastereomer pattern in the product and consequently in the dust must be expected (Abdallah et al., 2008a). Against this background the γ-diastereomer dominated in the UK with 56% and 69% (Abdallah et al., 2008a) and in the USA with 69% (Abdallah et al., 2008a), whilst in Belgium and in our study γ-HBCD only accounted for 27% and 33% of the total concentrations, respectively. In the two last-named studies α-HBCD was the dominant diastereomer.

USA (Stapleton et al., 2005) Canada (Wilford et al., 2004) USA (Wu et al., 2007) USA (Zota et al., 2008) USA (Sjödin et al., 2008b) USA (Harrad et al., 2008) USA (Batterman et al., 2009) USA (Wei et al., 2009) USA (Johnson et al., 2010) New Zealand (Harrad et al., 2008) China (Huang et al., 2010) Germany (Knoth et al., 2003) Sweden (Karlsson et al., 2007) UK (Pless-Mulloli et al., 2006) UK (Harrad et al., 2008) UK (Sjödin et al., 2008b) Germany (Sjödin et al., 2008b) Germany (Fromme et al., 2009) Belgium (D Hollander et al., 2010) Romania (Dirtu et al. 2012) Denmark (Vorkamp et al., 2011) Germany (this study)

100 x ng/g

BDE 47 BDE 99 BDE 100 BDE 153 BDE 183 0

10

20

30

40

50

60

70

80

ng/g dust Fig. 1. Polybrominated diphenyl ether (PBDE) (tetra to hepta congeners) in house dust (medians).

90

100

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H. Fromme et al. / Environment International 64 (2014) 61–68

Table 4 Median concentrations of non-PBDE brominated flame retardants detected in house dust of residences in ng/g. Reference

n

α-HBCD

β-HBCD

γ-HBCD

Σ-HBCD

TBBPA

BTBPE

Europe This study Sahlström et al. (2012) Abb et al. (2011) Abdallah et al. (2008a) Abdallah et al. (2008b) Geens et al. (2009) Roosens et al. (2009) D'Hollander et al. (2010) Ali et al. (2011b)

20 6 26 31 45 18 16 43 39

180 93

35 25

345 152

28

b10 b8

170 380

66 93

114 39 166 440 670

69

14

31

114 130

North America Abdallah et al. (2008a) Abdallah et al. (2008b) Stapleton et al. (2008) Batterman et al. (2009)

13 8 19 20

80 300

Asia Takigami et al. (2009) Wang et al. (2010)

2 27

a b c

300 230

TBPH

DBDPE

Study site

b3.0 53

343 430

146 1700

13

153

Germany Sweden Germany/USA UK UK Belgium Belgium Belgium Belgium

1.0

48 730 1300

62a 10

78c

28 72

EH-TBB

390 640 354b

2.0

48b

322b

234b

130b

57

140/13,000

490/520 6.5

2733

USA Canada USA USA

Japan China

n = 35. Geometric mean. TBBPE-DBPE.

4.3. “Novel” flame retardants in dust Table 4 shows the concentrations of “novel” BFRs found in the dust of homes published so far. Most studies looked at TBBPA with median concentrations between 10 ng/g and 78 ng/g, which were in a range that is comparable with our study (28 ng/g). However, there were some large ranges in the individual studies, with maximum values up to 2500 ng/g in the USA (Batterman et al., 2009), 1481 ng/g in Belgium (Geens et al., 2009), 470 ng/g in Germany/USA (Abb et al., 2011), and 382 ng/g in the UK (Abdallah et al., 2008b). The lowest maximum value with 233 ng/g was observed in our study. The median concentration of DBDPE found here is comparable with those in studies from Belgium (Ali et al., 2011b) and the USA (Stapleton et al., 2008). In contrast, Sahlström et al. (2012) and Wang et al. (2010) describe considerably a higher mean concentration of 5195 ng/g (470 ng/g–24,000 ng/g) and 2733 ng/g (100 ng/g–47,000 ng/g) in 6 and 27 dust samples from Sweden and China, respectively. The Chinese authors explain the high concentrations by the fact that DBDPE is the most frequently used BFR in China with utilisation rates of 80% at present. Similar to our results, Ali et al. (2011b) found low concentrations of BTBPE and TBB in the dust of 39 Belgian homes. However, in that study very high maximum values of 1019 ng/g and 436 ng/g respectively were also reported, whilst our study was not able to find any values above the detection limits of 10 ng/g and 3 ng/g respectively. Stapleton et al. (2008), too, describe higher median concentrations in dust in the “main living areas” of 20 homes in the USA. In this case the results show a broad concentration range of 4.7 ng/g–654 ng/g (BTBPE) and b6.6 ng/g–15,030 ng/g (EH-TBB). Moreover, in 6 Swedish apartments concentrations for BTBPE and EH-TBB of b4 to 550 ng/g and 25 ng/g to 440 ng/g were found (Sahlström et al., 2012). Results on TBPH in house dust have so far only been reported from Belgium (Ali et al., 2011b), the USA (Stapleton et al., 2008), and Sweden (Sahlström et al., 2012). Whilst in the first study a considerably lower median concentration was observed compared to our data, the broad range from b 2.0 ng/g to 5004 ng/g is similar to the range presented here with 25 to 2274 ng/g. Even higher concentrations, but with a similar median, were reported in the American study with a range of 3.0 ng/g to 10,630 ng/g. In Sweden the TBPH values ranged from 260 ng/g to 950 ng/g. In conclusion it becomes clear that measurement results in indoor dust may vary over several orders of magnitude due to the use of

different BFRs, the different amounts used and their different spectrum of application. For indoor living environments in Germany contamination from both BDE 209, HBCD and in particular from TBPH and DBDPE must be expected. Owing to the limited number of samples and the focus on indoor living environments, our study findings can only be used as an initial rough estimate of the current situation. Further more extensive studies, also in other indoor environments are therefore necessary. 4.4. Comparison of the PBDE intake with toxicological values So far only the US Environmental Protection Agency (US-EPA) has derived a so-called reference dose (RfD) for four PBDEs, for its Integrated Risk Information System (IRIS) (IRIS, 2012). For BDE 99 the RfD is 100 ng/kg b.w. As the toxicological point of departure a value of 0.29 mg/kg b.w. was selected as the Benchmark Dose Lower Confidence Limit (BMDL), the value at which behavioural changes were observed in the further development of mice after the oral administration of a single dose after birth. Taking into account a safety factor of 3000 (factor of 10 for extrapolating animal data to humans, 10 for intra-individual variability, 3 for extrapolating the administration of a single dose to lifelong exposure and 10 for the limited availability of toxicological data) a RfD of 100 ng/kg b.w. is obtained. Using the same safety factors and the same species the US-EPA derived a RfD of 100 ng/kg b.w. and 200 ng/kg b.w. for BDE 47 and BDE 153, respectively, with reference to effects associated with the development of the nervous system. Our “average” and “high” intake scenarios for adults and toddlers can be compared with the RfDs obtained by the US-EPA. For the sake of simplicity the bioavailability in humans is assumed to be equivalent to that found in animal experiments. Furthermore it is assumed that the aforementioned toxicological assessment values which were determined for BDE 99, the BDE with the greatest risk to health, also apply to all congeners from BDE 47 to BDE 183 and for the sum of all the congeners. In the “high” intake scenario we find exposure amounting to 0.1% (adults) and 0.7% (toddlers) of the RfD for the sum of the tetra- to heptabrominated congeners. For BDE 209 only a low systemic toxicity could be observed in the case of chronic exposure, the main target organs of which were the liver and the thyroid (IRIS, 2012). Here NOAEL (No-Observed-Adverse-EffectLevel) values between 1120 mg/kg b.w. and 2550 mg/kg b.w. were observed. Three studies conducted so far examined the sensomotoric development of rats or mice which had been administered BDE 209 orally

H. Fromme et al. / Environment International 64 (2014) 61–68

in the particularly critical period for this endpoint (several days after birth). Only in one target study a NOAEL of 2.2 mg/kg b.w. was derived. The US-EPA used this NOAEL as the point of departure for deriving a RfD of 7000 ng/kg b.w. using a safety factor of 300. When taking into account the “high” intake scenario, an exposure amounting to 0.008% of the RfD for BDE 209 is derived for adults and 0.09% of the RfD for children.

4.5. Comparison of the intake of other BFRs with toxicological values No scientific institutions have yet derived values for the tolerable daily intake (TDI) or comparable scales of values for HBCD. However, it is possible to estimate a margin of exposure (MOE) on the basis of the EU Risk Assessment Report (EU, 2008) and the more recent scientific literature (e.g. Marvin et al., 2011). In order to be able to estimate the risks of the general population over their entire lifetimes, the NOAEL of 10 mg/kg b.w. from a more recent Japanese two-generation reproductive toxicity study can be used (Ema et al., 2009). In this study a dose-related decrease in the fertility index in the F0 generation and a significant decrease in the number of primordial follicles in the ovary were observed. A comparison of our data with this NOAEL shows that there is an MOE of 13,100,000 for adults and 1,100,000 for toddlers in the “high” intake scenario. In addition, it seems reasonable to take into account the results of a further study with regard to toddlers. There, male mouse pups were dosed once by gavage at neonatal day 10 (Eriksson et al., 2006). At the age of three months, the mice were assessed for spontaneous behaviour and learning and memory capabilities. Based upon significantly altered spontaneous behaviour including hyperactive condition and reduced habituation a LOAEL (Lowest-Observed-Adverse-Effect-Level) of 0.9 mg/kg b.w. was estimated. In this case the MOE for the group of toddlers with regard to dust ingestion is still 100,000 (“high” intake scenario). There are hardly any toxicological data available on the two other “novel” brominated flame retardants detected in higher concentrations in house dust, TBPH and DBDPE. For instance, in the context of the High Production Volume Challenge Program in the USA a report was submitted about a study on rats which was conducted in 1988 and where TBPH was administered for 4 weeks. In this study no evidence of systemic toxicity and mortality, and no adverse effects on the reproductive organs were found. A NOAEL of 223 mg/kg b.w. was assessed based on decreased body weight, decreased liver enzyme levels, and a decrease in calcium and phosphorus levels in high dose females (HPV, 2004). No other studies are available, so it is currently not possible to make a valid estimate of the intake via house dust with toxicological assessment values. In this connection it seems particularly important for toxicological data on these two flame retardants to be determined in the future.

Conflict of interest The authors declare that they have no competing interests.

Acknowledgement The project was funded by the Bavarian Ministry of the Environment and Public Health.

Appendix A. Supplementary data Tables S 1, S 2a, S 2b, and Figures S 1 to S 6 were given as supplementary information. This material consists of results of the measurements of the reference material and of the regression calculation for BFRs. Supplementary data associated with this article can be found, in the online version, at http://dx.doi.org/10.1016/j.envint.2013.11.017.

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