Legacy and novel brominated flame retardants in interior car dust – Implications for human exposure

Environmental Pollution 230 (2017) 871e881

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Legacy and novel brominated flame retardants in interior car dust e Implications for human exposure* Athanasios Besis a, *, Christina Christia a, b, Giulia Poma b, Adrian Covaci b, Constantini Samara a a b

Environmental Pollution Control Laboratory, Department of Chemistry, Aristotle University of Thessaloniki, GR-54124 Thessaloniki, Greece Toxicological Center, University of Antwerp, Universiteitsplein 1, 2610 Wilrijk-Antwerpen, Belgium

a r t i c l e i n f o

a b s t r a c t

Article history: Received 27 April 2017 Received in revised form 8 July 2017 Accepted 11 July 2017

Brominated flame retardants (BFRs) are organobromine compounds with an inhibitory effect on combustion chemistry tending to reduce the flammability of products. Concerns about health effects and environmental threats have led to phase-out or restrictions in the use of Penta-, Octa- and Deca-BDE technical formulations, increasing the demand for Novel BFRs (NBFRs) as replacements for the banned formulations. This study examined the occurrence of legacy and NBFRs in the dust from the interior of private cars in Thessaloniki, Greece, aged from 1 to 19 years with variable origin and characteristics. The determinants included 20 Polybrominated Diphenyl Ethers (PBDEs) (Di-to Deca-BDEs), four NBFRs such as Decabromodiphenylethane (DBDPE), 1,2-bis(2,4,6-tribromophenoxy)ethane (BTBPE), 2-ethylhexyl2,3,4,5-tetrabromobenzoate (TBB), and bis(2-ethylhexyl)-3,4,5,6-tetrabromophthalate (TBPH), three isomers of hexabromocyclododecane (HBCD), and tetrabromobisphenol A (TBBPA). The concentrations of P PBDE ranged from 132 to 54,666 ng g
1 being dominated by BDE-209. The concentrations of P20
1 and were dominated by DBDPE, the major substitute of BDE4NBFRs ranged from 48 to 7626 ng g 209. HBCDs ranged between <5 and 1745 ng g
1, with alpha-HBCD being the most prevalent isomer Finally, the concentrations of TBBPA varied from <10 to 1064 ng g
1. The concentration levels and composition profiles of BFRs were investigated in relation to the characteristics of cars, such as year of manufacture, country of origin, and interior equipment (type of car seats, electronic and electrical components, ventilation, etc.). The average daily intakes of selected BFRs (BDE-47, BDE-99, BDE-153, BDE-209, TBB, BTBPE, TBPH, DBDPE, HBCDs and TBBPA) via ingestion and dermal absorption were estimated for adults and toddlers. The potential health risk due to BFRs was found to be several orders of magnitude lower than their corresponding reference dose (RfD) values. © 2017 Elsevier Ltd. All rights reserved.

Keywords: Brominated flame retardants (BFRs) Cars Indoor dust Human exposure

1. Introduction Brominated flame retardants (BFRs) are a diverse group of organic compounds extensively used in a wide range of products, including textiles, furniture foams, carpets, electrical devices, cables, building materials, insulators, etc., to reduce product flammability and to assure compliance with safety regulations (Kalachova et al., 2012). Up to 75 different BFRs have been produced, including: polybrominated diphenyl ethers (PBDEs), hexabromocyclododecanes (HBCDs), tetrabromobisphenol A (TBBPA)

*

This paper has been recommended for acceptance by Dr. Chen Da. * Corresponding author. E-mail address: [email protected] (A. Besis).

http://dx.doi.org/10.1016/j.envpol.2017.07.032 0269-7491/© 2017 Elsevier Ltd. All rights reserved.

and the novel BFRs (NBFRs) including the decabromodiphenyl ethane (DBDPE), 1,2-bis (2,4,6-tribromophenoxyethane) (BTBPE), 2-ethylhexyl-2,3,4,5-tetrabromobenzoate (TBB), bis-(2ethylhexyl)-3,4,5,6-tetrabromo-phthalate (TBPH) (Table S1) (Abdallah et al., 2015b; Al-Omran and Harrad, 2016 a,b; Covaci et al., 2011). Numerous studies have shown that these BFRs are ubiquitous substances in the environment, potentially toxic and can accumulate in tissues of living organisms (Harley et al., 2010; Main et al., 2007; Abdallah and Harrad, 2011). Also, they are considered as endocrine disruptors and causing neuro developmental disorders (Herbstman et al., 2010; Lilienthal et al., 2006). Concerns about health effects and environmental threats have led to strict bans and phase-out of PBDEs and in restrictions in the use of HBCDs (Directive ECC, 2003; European Court of Justice, 2008). The European Union (EU) banned the Penta- and Octa-BDE

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A. Besis et al. / Environmental Pollution 230 (2017) 871e881

formulations in 2004, and the Deca-BDE formulation in 2008 (Besis and Samara, 2012; Kemmlein et al., 2009), also Penta-, Octa-, and partly Deca-BDE mixtures were banned by the Stockholm's convention (2009) and UNEP (2009). The Penta and Octa formulations were voluntarily withdrawn from the US marketplace by their manufacturers at the end of 2004, leaving only the Deca formulation currently being marketed for use in commercial products in the United States (Besis and Samara, 2012). In December of 2009, the EPA announced the phase out of Deca-BDE (USEPA, 2009b). These new regulations have fueled the production and marketing of the novel BFRs (NBFRs). Ingestion of indoor dust is one of the main pathways of human exposure to PBDEs and potentially to BFRs (Ali et al., 2013; Besis and Samara, 2012; Cunha et al., 2010; Liagkouridis et al., 2014), especially for children, and toddlers because of crawling and hand-tomouth behavior (Johnson-Restrepo and Kannan, 2009; JonesOtazo et al., 2005). Dust is a complex heterogeneous mixture of particle matter derived from biological materials (as skin, human and animal hair, cells, fungal spores), fibers, cooking and heating emissions, cigarette smoke, soil brought in from outdoors and it can be found in all human indoor environments (Wilford et al., 2005). Numerous studies have investigated the presence of BFRs in indoor dust from houses and offices, but fewer in dust collected from cars (Ali et al., 2013, 2016; Harrad and Abdallah, 2011; Lagalante et al., 2009). While the average time that a human spends in automobiles is only 5.5% of the time spent indoors (Mandalakis et al., 2008), the median levels of BFRs in dust are higher in car dust than in house dust (Harrad et al., 2008b). Although many automobiles contain tinted glass to reduce light transmission and filter out short wavelength radiation automobiles can heat up to 90  C (Lagalante et al., 2011). High temperatures could cause higher release of BFRs from materials (seat polyurethane foams, electronic parts) in cars than in other indoor environments (Cetin and Odabasi, 2011). There are no previous studies investigating the presence of BFRs in car dust in Greece. The present study reports the levels of 20 congeners of PBDEs, HBCDs, TBBPA and four NBFRs (DBDPE, BTBPE, TBPH, TBB) in car dust samples taken in 2016 from 30 private cars in Thessaloniki, Greece. The construction year of the sampled cars ranged from 1997 to 2015 and the interior characteristics were recorded for further statistical correlation. The aims of our work were: (1) to explore any temporal changes in the profile of BFRs in cars produced before and after 2008, that is the year of the ban of the Deca-BDE, (2) to correlate the measured concentrations with the equipment in the interior cars, (3) to compare the BFR levels in car dust between cars produced in different continents (Europe, Asia and USA), and (4) to estimate the human exposure via car dust ingestion and dermal absorption. 2. Materials and methods 2.1. Sample collection Car dust samples were collected from 30 private cars in Thessaloniki, Greece, including gasoline- and diesel-engine automobiles with 1100e3000 cc. The average age of cars was 8 ± 5 years ranging from 1 to 19 years. Cars were classified according to the country of manufacture as following: Europe (n ¼ 15), Asia (n ¼ 9) and USA (n ¼ 6). Information concerning the origin of each car was provided by the official representatives of the corresponding car brands in Greece, nevertheless the possibility for some car parts to have been manufactured in another country, with different restrictions concerning the use of BFRs, cannot be excluded. All cars were sampled outdoors, under direct sun radiation, in parking and car washing spots. Potential influential variables on BFR

concentrations are provided in Supplementary Material (Table S2). The sampling protocol was based on previously reported methods in literature (Gevao et al., 2006; Harrad et al., 2008b; Kalachova et al., 2012). Car dust samples were collected using an 1800 W vacuum cleaner while paper bags were used. Before and after sampling, the metallic tubes of the vacuum cleaner were cleaned thoroughly using acetone to prevent cross contamination and a new paper bag was used. In each car, dust was collected by uniformly sweeping the front and back seats, the trunk, and the dashboard for 3 min. After sampling, each sample was sieved through a 1000 mm stainless steel sieve (Suzuki et al., 2006), onto a solvent rinsed aluminum foil to remove large particles and the sieved dust samples were sealed in plastic zip lock bags, and stored at
20  C. Questionnaires were completed by the car owners and the information is provided in Table S2. 2.2. Analytical procedure 2.2.1. Extraction and clean-up 2.2.1.1. PBDEs and NBFRs. PBDEs and NBFRs were simultaneously extracted from car dust samples using ultrasound extraction (ULTRAsonik, NEY 104H) in combination with agitation (Vortex, Heidolph) (Ali et al., 2011, 2013). Before extraction, 1 g of sieved dust (<1000 mm particle size) was spiked was spiked with 100 mL of 13 internal standard (IS) C-PBDEs containing BDE-15, -28,
47,
99,
153,
154,
183,
204,
207 and
209 at concentrations of 10e50 pg mL
1 to estimate the analytical recoveries. Then, 40 mL solvent mixture of n-hexane/DCM (1:1, v/v) was added and the sample was ultrasonicated for 20 min. The sample was vortexed for 5 min before and after the extraction. The extract volume was reduced to 1 mL in a rotary evaporator and applied to a glass column filled with neutral silica gel (particle size: 0.070e0.200 mm) at the bottom (1 g), 40% acidic silica gel (3 g), and Na2SO4 (0.5 g) on top. Each sample was eluted with 25 mL solvent mixture of n-hexane/DCM (3:2, v/v) and the sample was evaporated to 0.5 mL. For the solvent exchange, 1 mL of isooctane was added in the sample and the volume of the sample was further reduced to 0.5 mL and transferred into a glass vial. This was further evaporated to approximately 5 mL under a gentle nitrogen steam, and 20 mL of recovery standard (RS) (13C12 - BDE-139; 100 pg ml
1) was added. 2.2.1.2. HBCDs and TBBPA. Due to limitations in the available sample amount, HBCDs and TBBPA were analyzed only in 25 out of 30 samples. For the extraction of HBCDs and TBBPA, 30 mg of sieved dust (<1000 mm particle size) was weighted and spiked with 100 mL of IS of 13C-HBCD (alpha-, beta- and gamma-) and 13C-TBBPA. A solvent mixture of 2.5 mL containing n-hexane/acetone (3:1, v:v) was added and the sample was vortexed for 1 min, ultrasonicated for 5 min and centrifuged for 5 min (rotation speed 2500 g). The procedure was repeated twice and the pooled supernatants were loaded on Silica Bond Elut prepacked cartridges (500 mg, 3 mL) topped with 100 mg silica acid 44%. Target compounds were eluted using 12 mL of DCM. This fraction was evaporated near dryness under a gentle stream of N2, redissolved in 100 mL of methanol and transferred to injection vial for further analysis. 2.2.2. Instrumental analysis 2.2.2.1. PBDEs and NBFRs. The analysis of PBDEs and NBFRs was performed by 6890N Agilent gas chromatography and an Agilent 5973 mass spectrometer operating in electron impact ionization and selected ion monitoring (SIM) mode as described by Besis et al. (2014). A DB5-MS capillary column (15 m  0.25 mm x 0.1 mm) was used with helium as a carrier gas (1 mL min
1 flow rate). The

A. Besis et al. / Environmental Pollution 230 (2017) 871e881

temperature of the transfer line, ion source and quadrupole filter were set at 300  C, 230  C and 150  C, respectively. The identification of the BFR compounds was based on their retention times, target and qualifier ions, and were quantified using the IS calibration procedure. Further details of method are provided in Supplementary material (S1).

873

method limits of quantification (LOQs) were calculated as 3  standard deviation (SD) of the blank values and divided by the amount of dust used for analysis (30 mg) and were 5 ng g
1 for HBCDs and 10 ng g
1 for TBBPA. The recoveries of labeled IS ranged between 50 and 75%. 2.4. Calculation of daily intake

2.2.2.2. HBCDs and TBBPA. The determination of HBCDs and TBBPA was performed by an Agilent 1100 Series liquid chromatography equipped with autosampler and vacuum degasser as described by Dodson et al. (2012). A Luna C18 (2) reversed phase (RP) analytical column (150 mm  2 mm i.d., 3 mm particle size, Phenomenex) was used for the separation of HBCD isomers. A mobile phase of (A) ammonium acetate 2 mM in water/methanol (1:1 v/v) and (B) methanol at a flow rate of 0.250 mL min
1 was applied for elution of HBCD isomers. The elution gradient started at 75% (B), held for 2 min, then increased linearly to 100% until 9 min, held until 12 min followed by a linear decrease to 70% (B) over 0.5 min and held for 7.5 min. The target analytes were baseline separated with retention times of 4.0, 6.0, 6.8, 7.4 min for TBBPA, a-, b-, g-HBCD, respectively. MS analysis was performed using Agilent 6410 triple quadrupole MS system operated in the electrospray negative ionization mode (ESI). Nitrogen was used as drying gas at a flow of 10 mL min
1 and heated to 300  C. Nebulizer pressure was 35 psi and capillary voltage 4000 V. HBCD isomers were identified by isotopic dilution. MS/MS detection operated in MRM mode based on m/z 640.6 to 81 and m/z 652.6 to 81 for the native and 13C-HBCD labeled isomers, respectively. Fragmentor voltage and collision energy were set as 80 and 15 V respectively. For quantitative determination of TBBPA, the MRM: m/z 544 to 81 and m/z 652 to 81 for the native and 13CTBBPA were used. 2.3. Quality assurance and quality control 2.3.1. PBDEs and NBFRs A pre-cleaned dust sample, prepared by triplicate extraction with n-hexane/DCM/acetone (3:1:1, v/v) and vacuum-drying, was analyzed to assess blank levels of PBDEs and NBFRs, and to estimate the IS recoveries. The levels of target compounds in blank samples were in most cases lower than the instrument detection limit (LOD) and data were thus not blank-corrected. The LODs of target compounds were calculated as the concentration corresponding to signal/noise ratio of three (S/N ¼ 3). For PBDEs, LODs ranged between 0.01 and 13 ng (BDE-209) and for NBFRs ranged between 0.1 and 9 ng (DBDPE). The recovery efficiency of the extraction procedure was evaluated by the analysis of three blank dust samples spiked with known amounts of a PBDE standard containing 23 congeners (BDE-7, -15,
17,
28,
49,
47,
66,
77,
100,
99,
128,
153,
154,
138,
156,
183,
184,
191,
197,
196,
207,
206, and
209) and a NBFR standard containing DPDBE, BTBPE, TBPH, and TBB. The recoveries of 13C-PBDEs ranged between 86% for 13C-BDE-15 and 96% for 13C-BDE-204). The recoveries of NBFRs ranged between 80% for TBPE and 94% for TBB. The average recoveries of IS calculated in all the samples were 85 ± 11% (13C-BDE -15), 89 ± 9% (13C-BDE -28), 87 ± 10% (13C-BDE -47), 91 ± 9% (13C-BDE -99), 89 ± 6% (13C-BDE -153), 87 ± 9% (13C-BDE -154), 90 ± 10% (13C-BDE -183), 92 ± 8% (13C-BDE -204), 92 ± 10% (13C-BDE e207) and 88 ± 13% (13C-BDE -209). No target BFRs were detected in solvent blanks.

The average daily intake of PBDEs, NBFRs, HBCDs and TBBPA through dust ingestion (ADDingest) and dermal absorption (ADDdermal) was calculated by the following equations (USEPA, 1989, 2001):

ADDingest ¼ Cdust  IngR  EF  ED  ET=ðBW  ATÞ

(1)

ADDdermal ¼ Cdust  SA  DA  AF  EF  ED  ET=ðBW  ATÞ

(2)

where, C is the concentration of the contaminant in car dust (ng g
1); IngR is the daily ingestion of dust (assumed 0.05 g day
1 for adults and 0.06 g day
1 for toddlers, de Wit et al., 2012; USEPA, 2008 and 2009); EF is the exposure frequency throughout a year (assumed 350 days year
1; USEPA, 1989; and USEPA, 2001); ED is the life-time exposure duration (assumed 30 years for adults and 2 years for toddlers); ЕΤ is the time daily spent in cars (assumed 59 min day
1 for both, adults and toddlers, (de Wit et al., 2012; Harrad and Abdallah, 2011); BW is body weight (assumed 70 kg for adults and 12 kg for toddlers); AT is the number of days over which the exposure is averaged (365 days per year  ED for noncarcinogenic effects, USEPA, 2013); SA is the exposed body surface area (assumed 4615 cm2 for adults and 2564 cm2 for toddlers, Abdallah et al., 2009); DA is the dust adhered to skin (0.01 mg cm
2 for adults and 0.04 mg cm
2 for toddlers, Abdallah et al., 2009); and AF is the fraction of contaminant adsorbed by skin. Although the higher brominated BDEs are probably absorbed less readily from the gastrointestinal tract than Penta-BDEs and HBCDs (de Wit et al., 2012), in absence of better estimates for absorption by humans, we assumed 100% absorption of all BFRs. AF values for skin absorption of solvent-diluted BFRs were obtained from literature (33%, 34%, 37% and 8% for BDE-47, -99,
154 and
209, respectively, Abdallah et al., 2015a, b; 46% and 40% for HBCDs and TBBPA, respectively, Pawar et al., 2017; 11% and 8% for TBB and TBPH, respectively, Knudsen et al., 2016). Due to lack of other data, for DBDPE and BTBPE, we assumed the same AF values with BDE-209 and BDE-183 (8% for DBDPE that replaced BDE-209 and 13% for BTBPE that replaced, BDE-183). All other exposure factors were based on EPA (USEPA, 2008, 2009, 2011, 2013b). 2.5. Statistical analysis Descriptive analysis was computed using Microsoft Excel 2007 and SPSS (version 20.0). Spearman correlation coefficient was calculated to evaluate the relation among the concentrations of BFRs, while the Principal Component Analysis (PCA) was used to identify the major factors explaining the variability in the data set. Non-parametric ManneWhitney U test was used to study correlations between BFRs levels and characteristics from the sampling environments (p < 0.05 was considered statistically significant). 3. Results and discussion

2.3.2. HBCDs and TBBPA Four procedural blanks were analyzed, two per batch of samples, and certified reference material SRM 2585 from NIST was used to test the accuracy. Concentrations were blank-corrected and

3.1. Concentrations of PBDEs, NBFRs, HBCDs and TBBPA The BFRs quantified in the car dust samples included twenty

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A. Besis et al. / Environmental Pollution 230 (2017) 871e881

Table 1 Concentrations (ng g
1) of Polybrominated Diphenyl Ethers (PBDEs), NBFRs, HBCDs and TBBPA in car dust. Mean

Median

Min.

Max.

DF (%)

PBDEs (n¼30) BDE-15 BDE-17 BDE-28 BDE-71 BDE-47 BDE-66 BDE-77 BDE-100 BDE-119 BDE-99 BDE-85 BDE-154 BDE-153 BDE-138 BDE-183 BDE-181 BDE-203 BDE-207 BDE-206 BDE-209

4.41 3.90 14.2 41.3 375 13.5 7.22 55.8 13.3 476 2.12 76.9 71.5 16.4 150 33.9 3.52 138 111 7624

0.514 0.319 0.616 0.771 9.08 0.372 0.497 1.43 0.891 12.2 0.396 2.16 3.44 1.25 1.39 2.24 1.58 26.9 19.7 2830

0.09 0.09 0.16
55.6 87.6 332 1019 8964 288 133 1286 285 11,304 34.3 1973 1661 385 1206 239 23.3 1949 1225 37,729

100 100 100 97 100 93 93 97 87 100 93 100 100 97 97 97 87 100 100 100

∑20PBDEs

9231

2888

132

54,666

NBFRs (n¼30) TBB BTBPE TBPH DBDPE

38.3 96.8 233 1145

15.2 31.7 88.7 848

0.511 0.541 7.75 33.2

244 855 1553 5186

∑4NBFRs

1513

1188

47.8

7626

HBCDs (n¼25) alpha-HBCD beta-HBCD gamma-HBCD

178 28.4 58.2

90.3 15.8 46.4

< LOQ < LOQ < LOQ

1288 294 260

∑HBCDs

265

155

< LOQ

1745

44.8

e

< LOQ

1064

100 100 100 100

76 60 72

TBBPA (n¼25) TBBPA

16

LOD ¼ Limit of detection; LOQ ¼ Limit of quantification; DF ¼ Frequency of detection.

PBDE congeners, four NBFRs, a-, b-, and g-HBCDs, and TBBPA. The concentrations measured are summarized in Table 1, while the levels of individual compounds are provided in Table S3. ConcenP trations of 20PBDEs varied among sampled cars ranging from 132 to 54,666 ng g
1 (median and mean concentrations 2888 and 9231 ng g
1, respectively). Concentrations of four targeted NBFRs P ( 4NBFRs) ranged between 48 and 7626 ng g
1 (median and mean concentrations 1188 and 1513 ng g
1, respectively). Concentrations of HBCDs varied from <5 to 1745 ng g
1, while median and mean concentrations were 256 and 155 ng g
1, respectively. TBBPA was detected only in four samples with concentration levels varied from <10 to 1064 ng g
1. Due to the low detection frequency, TBBPA was further not used in the statistical analysis.

3.2. Comparison with other studies The literature values of the seven PBDE congeners most commonly found in car dust (BDE-28, -47,
100,
99,
154,
153, and
209) are provided in Table 2. The median concentration of P 7PBDE in the present study was lower than those reported in car dust in UK (Harrad et al., 2008b) and USA (Batterman et al., 2009; Lagalante et al., 2009, 2011), while lower values have been reported for car dust in Sweden (Thuresson et al., 2012), Nigeria (Harrad et al., 2016; Olukunle et al., 2015), Egypt (Hassan and Shoeib, 2015), Kuwait and Pakistan (Gevao et al., 2016; Ali et al., 2013), Saudi Arabia (Ali et al., 2016), Czech Republic (Kalachova et al., 2012), Portugal (Cunha et al., 2010) and Germany (Brommer et al., 2012) (Table 2). Variable particle size fractions were used in different studies to measure the BFR levels in car dust, e.g. <500 mm (Ali et al., 2013), <250 mm (Hassan and Shoeib, 2015; Ali et al., 2016), 25e500 mm (Harrad et al., 2008b, 2016; Brommer et al., 2012), and 20e250 mm (Lagalante et al., 2009, 2011). Wei et al. (2009), reported that PBDE concentrations and particle size were inversely related in car dust samples (n ¼ 1), but not in house dust samples (n ¼ 2). Chao et al. (2014) found no significant correlation between PBDE levels and particle size of floor and electronic dust. Cao et al., 2013, reported that >50% of total PBDEs in office dust was associated with particles <50 mm, whereas less than 20% with particles >200 mm. Nonetheless, Cao et al. (2015), found insignificant correlation between HBCD concentrations and dust particle size. Al-Omran and Harrad

Table 2 Concentrations of PBDEs measured in car dust in different countries (ng g
1). Country

Size fraction (mm) BDE28

UK (Birmingham) (n ¼ 20)* USA (n ¼ 12)* USA (Pennsylvania, New Jersey) (n ¼ 60)* Portugal (Porto) (n ¼ 9)*

25e500 NR 20e250 <150 whole 20e250 NR NR 25e500 <500 <500 <250 <150 25e500 <250 <250 <1000

USA (Pennsylvania) (n ¼ 66)* Sweden (Stockholm) (n ¼ 4)* Czech Republic (Prague) (n ¼ 27)* Germany **(n ¼ 12) Kuwait (Kuwait) (n ¼ 15)* Pakistan (Faisalabad) (n ¼ 15)* Egypt (Cairo) (n ¼ 9)* Nigeria (Makurdi) (n ¼ 12)* Nigeria (Lagos) (n ¼ 16)* Saudi Arabia (Jeddah)* (n ¼ 15) Kuwait* (n ¼ 19) Greece (Thessaloniki) (n¼30)*


*Median values; ** Mean; NR: not reported; LOD: Limit Of Detection.

BDE47

BDE99

BDE100

BDE153

BDE154

BDE183

BDE209

54 1800 880 88 78 590 7.4 2.2 17 5.8 1.2 5.7 68 28 10 4.9 9.1

100 2600 1130 23 110 613 11 <0.10 32 8.5 1.7 23 14 49 9.0 10 7.0

17 790 211 126 25 79 e <0.10 e 1.5 0.30 4.8 17 12 2.0 1.4 1.2

11 77 163 12 13 72 3.1 <0.30 e 1.5 0.90 e 16 9.0 1.5 1.4 3.4

11 120 105 23 4.0 10 e <0.30 e 1.1 0.30 e 19 3.6 1.5 0.8 2.2

7.8 73 73 6 4.0 2.9 2.2 <0.80 3.7 1.0 1.2 5.8 25 8.8 2.0 2.3 1.4

100,000 3100 48,100 1119 460 8120 1300 169 940 665 625 1540 122 780 200 480 2830

P

7PBDE

100,201 8573 50,708 1401 690 9491 1324 171 993 685 631 1579 282 892 310 501 2888

References Harrad et al., 2008b Batterman et al., 2009 Lagalante et al., 2009 Cunha et al., 2010 Lagalante et al., 2011 Thuresson et al., 2012 Kalachova et al., 2012 Brommer et al., 2012 Ali et al., 2013 Ali et al., 2013 Hassan and Shoeib, 2015 Olukunle et al., 2015b Harrad et al., 2016 Ali et al., 2016 Gevao et al., 2016 This study

A. Besis et al. / Environmental Pollution 230 (2017) 871e881

875

Table 3 Concentrations of NBFRs and HBCDs measured in car dust in different countries (ng g
1). P

P HBCDs

100 e 99 1300 1095 545 e 280 e e e e

100 e 99 e 1926 598 61 327 e e e e

e 54 33 e

848

1188

Country

Size fraction (mm)

TBB

BTBPE

TBPH

DBDPE

UK (Birmingham) (n ¼ 20)* Sweden (Stockholm) (n ¼ 4)* Czech Republic (Prague) (n ¼ 27)* Germany (n ¼ 12)** Kuwait (Kuwait) (n ¼ 15)* Pakistan (Faisalabad) (n ¼ 15)* Egypt (Cairo) (n ¼ 9)* Saudi Arabia (Jeddah)* (n ¼ 15) USA (n ¼ 12)* Kazakhstan (n ¼ 11)* Nigeria (n ¼ 10)* France (n ¼ 7)* Korea (n ¼ 19)* Greece (Thessaloniki) (n¼30)*

25e500 NR NR 25e500 <500 <500 <250 <250 NR <500 <500 <500 <212 <1000

e e e e 275 13 58 12 e e e e

e e e e 5.7 25 2.4 3.0 e e e e

e e e e 550 15 0.6 32 e e e e

15

32

89

4NBRs

e 38 e e 2036 297 5189 297 155

TBBPA

References

e e e e e e e e 110** 1** 2** 43** 127** 45**

Harrad et al., 2008b Thuresson et al., 2012 Kalachova et al., 2012 Brommer et al., 2012 Ali et al., 2013 Ali et al., 2013 Hassan and Shoeib, 2015 Ali et al., 2016 Batterman et al., 2009 Abdallah et al., 2016 Abdallah et al., 2016 Abdallah et al., 2016 Barghi et al., 2017 This study

*Median values; **Mean; NR: not reported.

(2016a), observed higher concentrations of the more volatile BFRs (Tri-to HeptaeBDEs) in the 25e63 mm particle fraction, although the levels of the less volatile BFRs (BDE-209, BTBPE, EH-TBB, DBDPE) did not differ significantly between the three particle fractions studied (25e63, 63e125 and 125e250 mm). Correspondingly, BFR concentrations were higher in finer particles of elevated surface dust, that had been in direct contact with materials containing BFRs, in contrast to floor dust (Al-Omran and Harrad, 2016b). It appears that other factors in addition to particle size, such as specific surface area, organic content of particles, and the potential presence of weathered appliances fragments (Al-Omran and Harrad, 2016a; Cao et al., 2013; Wei et al., 2009). P Median concentration of 4NBFRs in car dust samples was lower than those found in Kuwait (Ali et al., 2013), while lower concentrations have been reported in Egypt (Hassan and Shoeib, 2015), in Pakistan (Ali et al., 2013), in Czech Republic (Kalachova et al., 2012) and in Saudi Arabia (Ali et al., 2016) (Table 3). P Median concentration of HBCDs was higher than the value reported for car dust in Sweden (Thuresson et al., 2012), in Czech Republic (Kalachova et al., 2012) and in Egypt (Hassan and Shoeib, 2015) and lower than the value reported for car dust in France (Abdallah et al., 2016), in Kazakhstan (Abdallah et al., 2016), in

Fig. 1. Concentrations of

P

20PBDEs,

Nigeria (Abdallah et al., 2016) and in Korea (Barghi et al., 2017) (Table 3). For TBBPA mean concentration was lower than the value reported for car dust in USA (Batterman et al., 2009), in France (Abdallah et al., 2016) and in Korea (Barghi et al., 2017) (Table 3). 3.3. Evaluation of results according to car specific parameters P The highest 20PBDEs concentration was found in sample C29, a 15 years old car with fabric seats, whereas the lowest concentration was found in sample C22 manufactured in 2011 (Table S3). P For 4NBFRs, the lowest concentration was found in sample C26, P produced in 2002, and the highest concentrations for 4NBFRs and P HBCDs were found in sample C2 (year of production 2012) and C4 (manufactured in 2013), respectively. The concentrations of S20PBDEs, S4NBFRs and SHBCDs in car dust as a function of the year of manufacture are presented in Fig. 1. Interestingly, a decreasing trend was observed for S20PBDEs, whereas S4NBFRs and SHBCDs exhibited an increasing tendency. P More specifically, the mean concentrations of 20PBDEs in cars manufactured before 2008 were more than 2.5 times higher than in cars manufactured after 2008 particularly in European and Asian cars (Fig. 2), although, in all cases, the differences were not

P P HBCDs and 4NBFRs in car dust as a function of the year of manufacture.

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companies. P No significant differences were found for 20PBDEs and P HBCDs between cars with natural ventilation and air conditionP ing, except for 4NBFRs, that was almost three times higher in cars with A/C ventilation (p < 0.05). For all BFRs, no significant statistical differences were observed between cars with fabric or leather seats, as well as between vehicles with diesel or petrol engines. Also, no statistically significant correlation was observed with the number of electronics suggesting that BFRs are practically everywhere in car interior parts (i.e. flooring, dashboard, seats, panels, coatings, cables, insulation etc.). 3.4. Fingerprinting of BFRs

P P P Fig. 2. Concentrations of 20PBDEs, HBCDs and 4NBFRs in car dust collected from cars manufactured before and after 2008. The asterisks indicate statistically significant differences at p < 0.05.

P statistically significant (p > 0.05). On the contrary, 4NBFRs and P HBCDs exhibited higher mean concentrations in cars manufactured after 2008, but the difference was statistically significant only P for 4NBFRs in European and Asian cars (Fig. 2). Mean concentrations of PBDEs and NBFRs in cars manufactured before 2008 by Asian and European companies were comparable, decreasing by a factor of 5 and 2, respectively, for PBDEs and increasing by a factor of 4 and 9 for NBFRs in cars manufactured after 2008 (p < 0.05) (Fig. 2). For HBCDs, the mean concentration was two times higher in European cars manufactured after 2008 vs. cars manufactured before 2008 (p > 0.05) and no significant differences were observed for cars produced from Asian and American

The dominant PBDE congener in car dust samples was BDE-209 P (91% on average of 20PBDEs), followed by BDE-207 (1.9%), BDE-99 (1.9%), BDE-206 (1.4%), and BDE-47 (1.3%) (Fig. 3A). The prevalence of BDE-209 is in accordance with other studies worldwide (Ali et al., 2016; Harrad et al., 2008a, 2008b; Kalachova et al., 2012; Kang et al., 2011) highlighting the extensive use of Deca-BDE in the automotive industry. BDE-209 is the main congener in Deca-BDE technical mixtures that account for more than 80% of the total PBDEs production worldwide (BSEF, 2009). Moreover, its stronger affinity to the surface of solid particles and the lower volatility in comparison to other PBDEs enhance its prevalence in car dust (Wilford et al., 2005). The relatively high contents of BDE-207 (1.9%) and BDE-206 (1.4%) can be explained by their occurrence in Deca-BDE technical formulations, although at lower proportions in comparison to BDE209 (the proportion BDE-209:BDE-207:BDE-206 is 96.8:0.24:2.19% in Saytex 102E and 91.6:2.19:5.13% in Bromkal 82-0 DE, La Guardia et al., 2006). The relatively high contribution of BDE-47 (1.3%) and
99 (1.9%) are indicative of emissions from materials treated with Penta-BDE technical formulations (DE-71 and Bromkal 705DE). The BDE-47:BDE-99 ratio (calculated from median values) in this study was 0.78, very close to the corresponding ratio in DE-71 Penta-BDE formulation (0.79; La Guardia et al., 2006). Higher BDE47:BDE-99 ratios (up to 15) were observed in few samples, often in combination with high BDE-209 concentrations, thus suggesting possible thermal or photolytic degradation of BDE-209 to lower congeners (Bezares-Cruz et al., 2004). The predominant compound of NBFRs in car dust samples was P DBDPE (74% on average of 4BFRs), followed by TBPH, BTBPE, and ΤВВ with 17, 5.6, and 3.2%, respectively (Fig. 3B). The profile of NBFRs in Greek car dust is similar to those found in Saudi Arabia car dust (DBDPE > TBPH > TBB > BTBPE) (Ali et al., 2016) and in Kuwait car dust samples (DBDPE > TBPH > TBB > BTBPE) (Ali et al., 2013), whereas a different profile (DBDPE > BTBPE > TBPH > TBB) was found in Pakistani car dust (Ali et al., 2013) and in Egypt (TBB > BTBPE > TBPH) (Hassan and Shoeib, 2015). The high levels of DBDPE are indicative of its extensive use in the car materials. DBDPE is the substitute of BDE-209; hence its larger abundance compared to other NBFRs is not surprising. In addition, DBDPE has similar physicochemical properties compared to BDE209, with slightly higher hydrophobicity due to the ethane bridge between the two rings. It is contained in commercial mixtures Saytex® 8010 and Firemaster® 2100, both alternatives of Deca-BDE (Covaci et al., 2011). TBPH has been produced to substitute Penta-BDEs in polyurethane foams and is contained, together with TBB, in the commercial mixtures Firemaster 550 (ratio 1:4) and Firemaster BZ-54 (ratio 1:2.5) (Andersson et al., 2006). In this study, the TBPH:TBB ratio ranged from 1 to 80, which resemble closely the range (0.25e50) reported by Ali et al. (2013, 2016). This indicates other sources of TBPH such as DP-45, which contains ~100% TBPH, as well

A. Besis et al. / Environmental Pollution 230 (2017) 871e881

877

Fig. 3. Profiles of PBDEs, NBFRs and HBCDs.

as PVC and neoprene, in which TBPH is used as plasticizer (Andersson et al., 2006; Dirtu et al., 2012). Another plausible explanation could be a more rapid photodegradation of TBB relative to TBPH (Davis and Stapleton, 2009). With the exception of four cars, in which g-HBCD had a higher P contribution, the main contributor to HBCD was a-HBCD with 67% vs. 11 and 22% for b-HBCD and g-HBCD, respectively (Fig. 3C). In contrast, Kalachova et al. (2012) reported that mean and median concentrations of g-HBCD were almost three times higher compared to a-HBCD. A plausible explanation for the dominance of a-HBCD in car dust samples could be its presence in a higher percentage already in specific consumer products as obtained after a thermal pre-treatment (Harrad et al., 2009; Kajiwara et al., 2009;

Zhao et al., 2010). Moreover, Harrad et al. (2009), reported rapid photolytically-mediated shift from g-HBCD to a-HBCD that seems a likely explanation for the HBCD isomerization in dust.

3.5. Correlation analysis Spearman correlation coefficients among PBDEs congeners, P their substitutes NBFRs, and HBCDs were computed to investigate the possibility of common sources in the interior of cars (Table S4). Correlation coefficients among Tri-to Hexa-BDEs were high (0.800e0.979, p < 0.01), suggesting common origin. High correlation coefficients were also found among the higher brominated BDE congeners Octa-to Deca-BDEs (0.640e0.904, p < 0.01).

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Fig. 4. Principal component analysis (PCA): Score plot (A) and loading plot (B) for BFRs in car dust samples.

Fig. 5. Average daily amounts of exposure (ng kg
1 day
1) to BDE-47, -99, -153 -209, ea84eb2a2d33a1c0461a319b528270cf (ADDingest.) and dermal contact (ADDdermal) for adults and toddlers (box: 25e75%; (
) mean; (˥ ˩) min-max).

A. Besis et al. / Environmental Pollution 230 (2017) 871e881

879

Interestingly, Di-to Hepta-BDEs and Nona- and Deca-BDEs appeared to be negatively correlated with BTBPE, TBPH, and DBDPE (from
0.424 to
0.663, p < 0.01) suggesting that the phasing out of PBDEs was associated with their replacement by certain NBFRs. BDE-209, in particular, appeared to correlate negatively with its known alternate DBDPE (
0.461, p < 0.01), as well as with BTBPE (
0.492, p < 0.01). This scenario also found in car dust by Ali et al. (2016). Finally, HBCDs were negatively correlated with Octa-BDE (
0.414, p < 0.01), while positively with TBB (0.428, p < 0.05).

concerning the time spent in cars, and the ingestion rates for adults and toddlers. Nevertheless, it is worth to note that the median dust ingestion intake of BDE-209 estimated in this study for adults (8.09E-02 ng kg
1 day
1) is lower to those previously calculated for ingestion of central A/C filter dust in occupational indoor environments in Thessaloniki (2.08E-01 ng kg
1 day
1; Besis et al., 2014). This is in agreement with de Wit et al. (2012), who concluded that the intakes of PBDEs and HBCDs from car dust ingestion in Sweden were smaller than those from dust ingestion in office/day care centres or homes.

3.6. Principal component analysis

4. Conclusions

The principal component analysis (PCA) was used to further analyze the BFR profiles in car dust samples (Fig. 4). Three principal components (PCs) were found to explain 78% of the total variance in the original data set. PC1, accounting for 45%, showed high loadings for main components of Penta-BDE mixtures (Tri-to Hexa-BDEs) (Fig. 4). PC2, explaining 19% of the total variance, exhibited high loadings for NBRFs and HBCDs. PC3, explaining 15% of the total variance, showed high loadings with the higher brominated BDE congeners (Octa-to Deca-BDEs). The score plots (Fig. 4B) indicate a quite reasonable grouping of cars manufactured before 2008 and those manufactured after 2008. Nevertheless, some of the dust samples exhibited very similar BFR profiles despite the different period of car manufacture. This is probably an indication of delayed conformity of the automotive industries to regulations for phasing out PBDEs,, of different origin of different car parts, and/or come back of the banned compounds in the consumer products through recycling. 3.7. Human exposure via dust ingestion and dermal absorption To estimate the potential health risk due to BFRs, noncarcinogenic risk assessment is typically conducted using the hazard quotient (HQ), which is the ratio between the estimated exposure and the oral reference dose (RfD, ng kg
1 day
1), according to the following equation (3):

HQ ¼ ðADDÞ=RfD

P P P Median concentrations of HBCDs and 20PBDE, 4NBFRs, TBBPA were 2,888, 1,188, 155 and 45 ng g
1, respectively. The profile of PBDEs in dust samples suggested that Penta- and Deca-BDE technical mixtures were mostly used as BFRs in car equipment, together with DBDPE, the substitute of BDE-209. Finally, alphaHBCD was the most prevalent isomer for HBCDs. PBDEs mean concentration was higher in cars that manufactured before 2008 in comparison with cars that manufactured after 2008, while for NBFRs the mean concentration was higher in cars manufactured before 2008 in comparison with cars manufactured after 2008. The intakes of BFRs via ingestion and dermal absorption were several orders of magnitude lower than their corresponding RfD values. In general, the intake of BFRs was higher via dust ingestion than via dermal absorption. Also, regardless the route of exposure, it was higher for toddlers than for adults. Acknowledgements Dr. Giulia Poma acknowledges a post-doctoral fellowship from the University of Antwerp, while Mrs. Christina Christia was supported by an Erasmus Plus exchange program. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2017.07.032.

(3)

where, ADD is the average daily dose (ng kg
1 day
1), either through ingestion (ADDingest.) (Eq. (1)) or dermal contact (ADDder
1 day
1). mal) (Eq. (2)), and RfD is the oral reference dose (ng kg The available RfD values for BFRs are provided in Table S5. The daily doses of PBDEs (BDE-47, -99,
153,
209), NBFRs (TBB, BTBPE, TBPH, DBDPE), HBCDs and TBBPA calculated for adults and toddlers are presented in Fig. 5 and Table S6. As expected, the highest exposure via dust ingestion was calculated for BDE-209 followed by its alternative DBDPE. Correspondingly, the highest exposure via dermal contact was calculated P for BDE-209 followed by its alternative DBDPE and HBCDs. For all BFRs, exposure via dust ingestion was higher than exposure via dermal absorption in agreement with other studies (JohnsonRestrepo and Kannan, 2009). Also in agreement with other studies, exposure of toddlers appeared to be higher than those of adults (Ali et al., 2013; Jones-Otazo et al., 2005). Apparently, the estimated ADD values of BFRs via both exposure routes were significantly lower than their RfD values (Table S5). It is worth to note that dust ingestion represents only one of the human's exposure pathways to BFRs, while detailed information on this group of contaminants for the typical Greek diet is not available. It is difficult to compare our ADD values to results from other studies due to the large variation among different studies

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