Polyurethane foam-based passive air sampling for simultaneous determination of POP- and PAH-related compounds: A case study in informal waste processing and urban areas, northern Vietnam

Chemosphere 247 (2020) 125991

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Polyurethane foam-based passive air sampling for simultaneous determination of POP- and PAH-related compounds: A case study in informal waste processing and urban areas, northern Vietnam Hoang Quoc Anh a, b, Isao Watanabe a, Nguyen Minh Tue c, d, Le Huu Tuyen d, Pham Hung Viet d, Ngo Kim Chi e, Tu Binh Minh b, Shin Takahashi a, * a

Center of Advanced Technology for the Environment (CATE), Graduate School of Agriculture, Ehime University, 3-5-7 Tarumi, Matsuyama, 790-8566, Japan Faculty of Chemistry, VNU University of Science, Vietnam National University, Hanoi, 19 Le Thanh Tong, Hanoi, 100000, Viet Nam Center for Marine Environmental Studies (CMES), Ehime University, 2-5 Bunkyo-cho, Matsuyama, 790-8577, Japan d Centre for Environmental Technology and Sustainable Development (CETASD), VNU University of Science, Vietnam National University, Hanoi, 334 Nguyen Trai, Hanoi, 100000, Viet Nam e Institute of Natural Products Chemistry, Vietnam Academy of Science and Technology, 18 Hoang Quoc Viet, Hanoi, 100000, Viet Nam b c

h i g h l i g h t s

g r a p h i c a l a b s t r a c t


PCBs, BFRs, and PAHs were comprehensively monitored in Vietnamese air samples.
Concentrations of pollutants decreased: PAHs > MePAHs > PCBs > PBDEs > NBFRs.
The abundance of mono- and di-CBs (notably CB-11) was found in Vietnamese air.
BaP-EQs and WHO-TEQs were higher in waste processing sites than urban sites.

a r t i c l e i n f o

a b s t r a c t

Article history: Received 2 December 2019 Received in revised form 14 January 2020 Accepted 20 January 2020 Available online 21 January 2020

Polychlorinated biphenyls (PCBs), polybrominated diphenyl ethers (PBDEs), hexabromobiphenyl (BB153), novel brominated flame retardants (NBFRs), and unsubstituted/methylated polycyclic aromatic hydrocarbons (PAHs/Me-PAHs) were simultaneously monitored in the air samples collected from Vietnamese urban and vehicular waste processing areas by using polyurethane foam-based passive air sampling (PUFePAS) method. Concentrations (pg m3) of organic pollutants decreased in the order: PAHs (median 29,000; range 5100e100,000) > Me-PAHs (6000; 1000e33,000) > PCBs (480; 170 e1100) > PBDEs (11; 5.3e86) > NBFRs (0.20; n. d. e 51) > BB-153 (n.d.). The difference in total PCB and PBDE concentrations between the urban and waste processing air samples was not statistically significant. Meanwhile, levels of PAHs, Me-PAHs, benzo [a]pyrene equivalents (BaP-EQs), and toxic equivalents of dioxin-like PCBs (WHO-TEQs) were much higher in the waste processing sites. This is the first report on the abundance of mono- and di-CBs (notably CB-11) in the air from a developing country, suggesting their roles as emerging and ubiquitous air pollutants. Our results have indicated potential sources of specific organic pollutants such as dioxin-like PCBs, PAHs, and Me-PAHs from improper treatment of end-of-life vehicles and other vehicle related materials (e.g., waste oils and rubber tires), as well as current emission of PCBs and PBDEs in the urban area in Vietnam. Further atmospheric monitoring

Handling editor: R Ebinghaus Keywords: PUFePAS PCBs PBDEs PAHs Informal waste processing Urbanization

* Corresponding author. Center of Advanced Technology for the Environment, Graduate School of Agriculture, Ehime University, 3-5-7 Tarumi, Matsuyama, 7908566, Japan. E-mail address: [email protected] (S. Takahashi). https://doi.org/10.1016/j.chemosphere.2020.125991 0045-6535/© 2020 Published by Elsevier Ltd.

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H.Q. Anh et al. / Chemosphere 247 (2020) 125991

studies should be conducted in this developing country that cover both legacy and emerging contaminants with a focus on areas affected by rapid urbanization and informal waste processing activities. © 2020 Published by Elsevier Ltd.

1. Introduction Semivolatile organic compounds (SVOCs) including various persistent organic pollutants (POPs) and polycyclic aromatic compounds (PACs) are usually monitored in the air due to their volatile, persistent, and toxic nature (Bogdal et al., 2013; Bao and Zeng, 2014; Melymuk et al., 2014). The conventional dynamic/active air sampling (AAS) techniques with high or low volume air samplers (HVAS or LVAS) can yield accurate air concentrations of target pollutants based on their amounts accumulated in the sorbent bed and clearly defined air volume sucked by the pump (Marc et al., 2016). However, the AAS techniques have major disadvantages as expensive operation and relatively complex instrument that prevent large-scale and frequent monitoring, particularly in developing countries and remote areas (Shoeib and Harner, 2002; Pozo et al., 2006; Tuduri et al., 2012; Bogdal et al., 2013; Marc et al., 2015, 2016). To overcome such constraints, diffusion/passive air sampling (PAS) techniques have been introduced and applied in local, national, regional, and even global monitoring studies on POPs and PACs (Bogdal et al., 2013; Melymuk et al., 2014). One of the most popular PAS methods is based on polyurethane foam disks (PUFePAS) with main features such as cheap, simple, and applicable to a wide range of SVOCs (Bogdal et al., 2013; Melymuk et al., 2014; Bidleman and Melymuk, 2019). Despite its wide application, there is no consistent method to estimate sampling rate and/or sampled air volume of PUFePAS, especially for novel and highly volatile pollutants (Chaemfa et al., 2009; Harner et al., 2013; Bohlin et al., 2014; Melymuk et al., 2014; Herkert et al., 2018). Vietnam is a Southeast Asian developing country and has increasing concerns about air pollution (WHO, 2018); but comprehensive and systematic monitoring studies on organic air pollutants are still very limited in this country. To our knowledge, there are few studies that applied PUFePAS method for monitoring of SVOCs in the Vietnamese air (Tue et al., 2013; Wang et al., 2016; Anh et al., 2019a). These investigations have indicated that the air pollution caused by organic pollutants in Vietnam is likely associated with urbanization processes and informal processing activities of modern waste such as electrical and electronic waste (e-waste) and end-of-life vehicles (ELV) (Tue et al., 2013; Wang et al., 2016; Anh et al., 2019a). However, these studies have at least one of the following limitations: applying only one generic sampling rate of 3.5 m3 d1 to calculate air volume and concentrations of a variety of pollutants with different physicochemical properties, analyzing limited number of compounds, or having inadequate sensitivity, which may lead to some incomprehensive evaluations. In the present study, we report results of PUFePAS sampling method for the simultaneous determination of 209 PCBs; 36 polybrominated diphenyl ethers (PBDEs), hexabromobiphenyl (BB153), and 3 novel brominated flame retardants (NBFRs) such as pentabromoethylbenzene (PBEB), 1,2-bis-(2,4,6-tribromophenoxy) ethane (BTBPE), and decabromodiphenyl ethane (DBDPE); and 18 polycyclic aromatic hydrocarbons (PAHs) and 12 methylated derivatives (Me-PAHs) in the air samples collected from some representative locations in northern Vietnam, which have been affected by urbanization and improper waste processing activities. To our knowledge, this is the first study to investigate full congenerspecific profiles of PCBs in the PUFePAS air samples with a focus

on low-chlorinated PCBs with one or two chlorine atoms, by applying an appropriate calculation method for air concentrations of highly volatile compounds. This initial database emphasizes the need of comprehensive and detailed studies on air pollution monitoring, which consider not only “legacy” pollutants such as technical PCB congeners, PBDEs, and 16 priority PAHs, but also “emerging” and infrequently monitored contaminants like unintentionally produced PCBs, NBFRs, and Me-PAHs. 2. Material and methods 2.1. Sample collection The air samples were collected between JanuaryeMarch 2013 (average temperature 20  C) and SeptembereNovember 2015 (average temperature 27  C) in three locations in northern Vietnam: an urban area in Hanoi City, an informal ELV processing area in Bac Giang Province, and a waste recycling (WR) cooperative in Thai Nguyen Province. The samplers were deployed in different micro-environments of both indoor, outdoor, and semi-open spaces including: urban houses (UIeH1, 2, 3), urban office (UI-OF), urban outdoors in a residential area (UO-R) and a mixed land use area (UO-M); semi-open ELV workshops (ELV-1, 2, 3) and rural outdoor (ELV-0); and some WR facilities for plastic recycling (PR), oil refining (OR-1, 2), and rubber melting (RM-1, 2). Detailed information on the studied areas and sampling sites are presented in Table S1 and Fig. S1 of Supplementary data, and were described in our previous studies (Takahashi et al., 2017; Anh et al., 2019a, b, c). The PUF disks (136 mm diameter, 13 mm thickness, and 0.0140 g cm3 density; INOAC Corporation) were washed by Soxhlet extraction with acetone for 24 h, dried in a vacuum desiccator, wrapped in aluminum foil, and sealed in polyethylene ziplock bag until deployment. Each sampling chamber comprises two stainless steel bowls (upper 26 cm and lower 20 cm in diameter) with a 2-cm gap between the two bowls to facilitate the air flow. The samplers were allowed to hang free at 2e3.5 m above the ground over periods of 4e7 weeks. The heights of samplers were based on filed surveys for appropriate hanging points and were in line with those applied in various studies using the same sampling method (Sun et al., 2016; Khan et al., 2017; Yadav et al., 2017; Okeme et al., 2018; Cetin et al., 2019). After the sampling periods, the samplers were disassembled and the PUF disks were collected and stored at 20  C until analysis. Further information on the sampling method was described in our previous studies (Tue et al., 2013; Anh et al., 2019a). 2.2. Chemical analysis The PUF disks were Soxhlet extracted with acetone for 16 h and the extracts were evaporated and solvent-exchanged into 10 mL hexane (Anh et al., 2019a). For PCB and BFR analysis, a portion of 4 mL extract was spiked with 13C12-PCBs and 13C12-BDE-209 (Wellington Laboratories), and monofluorinated FBDEs (AccuStandard) as surrogates. The crude extract was purified by using a multi-layer silica gel column and an activated silica gel column with elution solvents as mixtures of 25% and 10% dichloromethane (DCM) in hexane, respectively. The eluate was concentrated and

H.Q. Anh et al. / Chemosphere 247 (2020) 125991

spiked with internal standards of 13C12-PCBs (Wellington Laboratories) and FBDE-154 (AccuStandard) before injection into GCeMS. For PAH/Me-PAH analysis, a portion of 2 mL extract was spiked with deuterated surrogate standards (Cambridge Isotope Laboratories) and applied to an activated silica gel column. The column was washed by hexane to remove aliphatic compounds and the target compounds were eluted by mixture of 50% DCM in hexane. The eluate was concentrated and spiked with chrysene-d12 (Cambridge Isotope Laboratories) as internal standard before GCeMS analysis. Solvents and chemicals were analytical grade and supplied by Wako Pure Chemical Industries. PCBs were determined by a gas chromatograph (6890 N; Agilent Technologies) connected to a double-focusing mass spectrometer (JMS-800D; JEOL) at electron impact (EI) ionization mode and selected ion monitoring (SIM) mode. BFRs and PAHs/Me-PAHs were analyzed by a GCeMS system (GCMSeQP2010 Ultra; Shimadzu) at negative chemical ionization (NCIeSIM) mode and EIeSIM mode, respectively. The instrumental conditions were described in our previous studies (Anh et al., 2018, 2019c, d, e) and are summarized Table S2. 2.3. Quality assurance and quality control The recovery tests were performed by triplicate analysis of PUF disks spiked with native standards of all the target compounds. Travel blank and procedural blank samples were analyzed simultaneously with real samples to check for contamination of the whole procedure. There were no significant levels of almost all the analytes in procedural and travel blanks. Concentrations of the target compounds were corrected by average levels in the procedural blanks. The method detection limits (MDLs) were derived based on the instrumental detection limits, final extract volume, and volume of air samples. The MDL of a compound with nonnegligible blank level was calculated by average blank plus 3 times the standard deviation. In addition, a detected compound must have a signal-to-noise ratio 3. The MDLs ranged from 0.030 to 0.50 pg m3 for PCBs, from 0.050 to 1.0 pg m3 for BFRs, and from 3.0 to 10 pg m3 for PAHs and Me-PAHs (see details in Table S6). Due to the large number of target compounds, concentrations below the MDLs were assigned as zero to avoid overestimation of sum concentrations. Recoveries of native standards in test samples and surrogate standards in real samples were within the range of 60%e120% with relative standard deviations RSD <15% (Table S3). 2.4. Data analysis Air concentrations of PCBs, BFRs, PAHs, and Me-PAHs were derived by their amounts accumulated in the PUF disks and an effective air volume. The effective air volume is a function of sampling rate, deployed duration, PUF disk parameters, nature of target compounds, and meteorological conditions (Tuduri et al., 2012; Harner et al., 2013). The equations used to calculate air concentrations of organic pollutants are tabulated in Table S4. Based on these equations, we applied the template for calculating PUF disk sample air volumes proposed by Harner (2016) with corrected parameters of PUF disks and average temperature. It should be noted that the current PUFePAS method has focused on vapor phase despite the fact that most of SVOCs are distributed between gaseous and particulate phases. Although some previous studies indicated similar PUFePAS sampling rates of gas- and particle-phases for POPs and PACs (Bohlin et al., 2010; Melymuk et al., 2011; Harner et al., 2013), the performance of PUFePAS samplers for particle-associated compounds as well as phase distribution characteristic of these pollutants have not been well characterized (Melymuk et al., 2014). Therefore, in the current study, derived concentrations of air pollutants were considered as

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bulk concentrations with equal sampling rates for both gas- and particle-phases. Microsoft Excel (Microsoft Office, 2010) and Minitab 16® Software (Minitab Inc.) were used for statistical analysis. The nonparametric Mann-Whitney U test was applied to find out the difference in air contamination levels between the studied locations. Potential relationships among the target compounds were assessed by using Pearson correlation analysis. The significance level was set at p < 0.05. 3. Results and discussion 3.1. Derivation of air concentrations of organic pollutants by PUFePAS method Air concentrations of the target compounds were calculated based on their amounts (in pg) measured in the PUF disks (see details in Table S5) and effective air volume (in m3). The effective air volume for all most all the compounds ranged from 130 to 200 m3 per sampler, except for some low-chlorinated PCBs. The PUFePAS derived concentrations of 279 organic pollutants (including 13 and 2 co-eluted peaks for PCBs and PAHs/Me-PAHs, respectively) are presented in Table S6 and a summary is shown in Table 1. There are several ways to elucidate sampling rate and/or sampled air volume in the PUFePAS method: (1) calibrating with gas-phase concentrations derived by active air sampling; (2) using performance reference/depuration compounds spiked into PUF disks to estimate sampling rate of target compounds based on depuration rate of reference compounds; (3) applying linear kinetic sorption models that are sampler- and compound-specific; and (4) just applying a generic sampling rate (e.g., 3.5 or 4 m3 d1) (Tuduri et al., 2012; Harner et al., 2013; Melymuk et al., 2014). Methods (1) and (2) are more practical and site-specific, which consider the meteorological effects. However, each way has its own limitations: (1) AAS instrument and well-designed calibration studies are needed; (2) some depurated compounds do not represent a huge number of pollutants; and (4) the important factor of pollutants’ nature is not accounted. In our previous studies, we applied a generic sampling rate of 3.5 m3 d1 for all the target compounds including PCBs, PBDEs, PAHs, and other SVOCs (Tue et al., 2013; Anh et al., 2019a). In the current work, we calculated air concentrations of pollutants by using method (3) in order to provide more accurate PUFePAS monitoring results, as compared with those obtained by method (4). Based on our calculation, almost all the compounds with adjusted PUF-air partition coefficients K’PUF > 106 have similar effective air volume and sampling rate (e.g., 3.5e4 m3 d1) during sampling duration of 4e7 weeks. These observations were in accordance with those reported elsewhere, which indicated that SVOCs with octanol-air partition coefficients KOA > 108.5 have PUFePAS sampling rates of 3e5 m3 d1 during about 3-month period (Harner et al., 2004; Jaward et al., 2004; Wilford et al., 2004). Except for more volatile pollutants such as mono-to triCBs and di-BDEs, the specific sampling rates of almost PCBs and PBDEs were in good agreement with the generic sampling rate of 4 m3 d1 and with those derived by AAS calibration or depuration methods (Chaemfa et al., 2008; Bohlin et al., 2014; Herkert et al., 2018). The PUFePAS sampling rates and linear uptake phase of mono- and di-CBs have not been fully characterized. Herkert et al. (2018) reported that sampling rates for mono- and di-CBs were 5.08 ± 1.30 and 4.80 ± 1.23 m3 d1, respectively; but that study used only 4 depuration compounds (e.g., 13C12-CB-9, -15, 32, and CB30) to calibrate sampling rates for 10 PCB homologs. The decline in sampling rates and linear uptake ranges with decreasing chlorination degree and KOA was documented only for tri-to hepta-CBs (Shoeib and Harner, 2002; Melymuk et al., 2011). Our estimates

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Table 1 Concentrations (pg m3) of PCBs, PBDEs, BB-153, NBFRs, PAHs, and Me-PAHs in the air samples from northern Vietnam.

Mono-CBs Di-CBs a CB-11 Tri-CBs Tetra-CBs Penta-CBs Hexa-CBs Hepta-CBs Octa-CBs Nona-CBs Deca-CB Ʃin-PCBs b Ʃdl-PCBs c WHO-TEQs d Ʃ208PCBs e Ʃ209PCBs S35PBDEs f BDE-209 S36PBDEs g BB-153 PBEB BTBPE DBDPE S18PAHs S12Me-PAHs S30PAHs BaP-EQs h a b c d e f g h i

UI-H1

UI-H2

UI-H3

UI-OF

UO-R

UO-M

ELV-0

ELV-1

ELV-2

ELV-3

PR

OR-1

OR-2

RM-1

RM-2

38 34 260 39 19 8.2 3.8 0.95 n.d. i n.d. n.d. 14 1.4 0.048 140 410 5.4 5.1 10 n.d. 0.12 n.d. n.d. 9800 2000 12,000 88

51 230 280 110 30 8.4 2.5 0.53 n.d. n.d. n.d. 21 1.0 0.036 430 710 4.4 11 16 n.d. n.d. n.d. n.d. 11,000 1000 12,000 20

37 37 100 31 15 8.7 4.3 1.0 n.d. n.d. n.d. 12 1.6 0.047 130 230 14 28 42 n.d. 0.11 0.31 n.d. 13,000 1500 14,000 28

46 150 200 90 29 13 4.9 0.52 n.d. n.d. n.d. 23 1.7 0.060 330 530 6.3 25 31 n.d. 0.20 n.d. n.d. 10,000 2000 12,000 200

76 250 300 120 35 9.2 3.2 0.46 n.d. n.d. n.d. 26 0.95 0.028 500 810 5.0 5.0 10 n.d. 0.19 n.d. n.d. 18,000 3000 21,000 41

45 22 99 53 48 43 19 3.3 0.12 n.d. n.d. 39 10 13 230 330 13 28 41 n.d. 0.32 0.20 n.d. 41,000 5000 46,000 360

36 8.7 79 11 14 16 6.4 3.2 n.d. n.d. n.d. 12 3.6 5.2 96 170 3.5 7.0 10 n.d. 0.18 0.17 n.d. 28,000 6000 34,000 260

60 110 140 70 50 45 9.3 1.1 n.d. n.d. n.d. 36 6.8 4.8 340 480 3.8 7.0 11 n.d. 0.18 n.d. n.d. 29,000 7500 36,000 170

130 180 160 170 94 80 26 5.9 1.7 0.78 0.19 70 20 79 690 850 8.4 15 24 n.d. 0.25 n.d. 11 50,000 33,000 83,000 480

140 68 110 52 42 42 13 1.9 0.64 n.d. 0.12 31 9.0 25 360 470 18 16 34 n.d. 0.30 n.d. n.d. 47,000 15,000 62,000 460

37 170 140 70 17 26 22 3.1 n.d. n.d. n.d. 30 14 0.44 340 480 1.5 4.0 5.5 n.d. n.d. n.d. n.d. 5100 1300 6400 120

330 290 180 170 61 41 30 9.2 1.0 n.d. n.d. 60 9.3 14 930 1100 2.2 3.8 6.0 n.d. n.d. 0.66 n.d. 57,000 17,000 74,000 400

150 150 120 100 47 34 25 9.0 0.76 n.d. n.d. 47 8.3 14 510 640 1.8 3.5 5.3 n.d. n.d. n.d. n.d. 40,000 13,000 53,000 300

51 94 68 49 21 10 4.8 1.2 n.d. n.d. n.d. 15 2.2 7.8 230 300 5.3 80 86 n.d. n.d. n.d. 51 100,000 10,000 110,000 910

52 200 160 76 21 9.5 3.0 0.86 n.d. n.d. n.d. 15 1.1 0.049 360 520 2.6 7.4 10 n.d. n.d. n.d. n.d. 61,000 7400 69,000 460

Dichlorobiphenyls excluding CB-11. Indicator PCBs (CB-28, -52, 101, 118, 138, 153, and 180). Dioxin-like PCBs (CB-77, -81, 126, 169, 105, 114, 118, 123, 156, 157, 167, and 189). TEQs of dl-PCBs (fg TEQ m3) based on WHO 2005 TEFs proposed by Van den Berg et al. (2006). Total PCBs excluding CB-11. Total of 35 di-to nona-BDEs. Total of 36 di-to deca-BDEs. BaP-EQs (pg BaP-EQ m3) based on TEFs proposed by Nisbet and Lagoy (1992). Not detected.

have revealed that the average sampling rates of mono- and di-CBs were 3.0 ± 0.5 and 3.3 ± 0.4 m3 d1 over linear uptake ranges of 10 days and 4 weeks, respectively (see details in Fig. S2). Meanwhile, the linear regression between the effective air volume and sampling duration of heavier congeners was observed until 8 weeks or even more. The sampling rates of PAHs and Me-PAHs varied greatly by one to two orders of magnitude (e.g., 0.2e40 m3 d1) between compounds and studies (Melymuk et al., 2011; Bohlin et al., 2014; Strandberg et al., 2018). The linear uptake ranges of low-molecularweight PAHs such as naphthalene and acenaphthylene have been estimated to be 1e2 weeks (Bohlin et al., 2014). This situation suggests the need of further calibration studies for low chlorinated PCBs and light PACs to correctly evaluate their sampling rates and linear uptake periods by the PUFePAS method. The deployment of several PUFePAS samplers with different sampling durations at each studied station is recommended to obtain more accurate air concentrations of highly volatile pollutants.

3.2. Concentrations and profiles of PCBs Concentrations of total PCBs (S209PCBs) ranged from 170 to 1100 (median 480) pg m3 with the most dominant congener as CB-11 (median 140; range 68e300 pg m3), which accounted for 16%e63% (average 32%) of S209PCBs (Fig. 1). Median concentration of CB-11 was higher in the urban samples (230 pg m3) than the ELV and WR sites (140 pg m3) while concentrations of S208PCBs (total PCBs excluding CB-11), indicator PCBs (Sin-PCBs), and dioxinlike PCBs (Sdl-PCBs) were generally higher in the ELV and WR sites (Table 1). The highest PCB concentrations were found in the air

samples collected from an oil refining facility (OR-1; 1100 pg m3) and from an ELV workshop (ELV-2; 850 pg m3). The difference in total PCB concentrations between the urban and waste processing air samples was not statistically significant, largely due to the ubiquitous presence of PCBs in Vietnamese air and/or small sample set of data. The toxic equivalents (WHO-TEQs) derived by using concentrations of dl-PCBs and the toxic equivalency factors (WHO, 2005 TEFs; Van den Berg et al., 2006) in the air samples from the waste processing sites (median 7.8; range 0.049e79 fg TEQ m3) were significantly higher than those detected in the urban samples (0.050; 0.028e13 fg TEQ m3) (p < 0.05). Apart from CB-11, the most important contributors to total PCBs were other di-CBs (average 23%), mono-CBs (15%), and tri-CBs (14%). These low-chlorinated congeners (i.e., mono-to tri-CBs) together accounted for 66%e94% (average 85%) of total PCBs. Proportions of medium-chlorinated congeners (i.e., tetra-to hepta-CBs) were higher in the waste processing sites and a mixed land use urban site (UO-M) (average 19%) as compared to other urban sites (8.2%). Based on our initial data, we proposed a novel set of PCBs including 7 conventional indicator congeners (i.e., CB-28, -52, 101, 118, 138, 153, and 180) and 9 low-chlorinated congeners (i.e., CB-1, -2, -3, -4, -5/8, 11, 18, 31), which hereinafter referred to as atmospheric indicator PCBs (ai-PCBs). Concentrations of ai-PCBs exhibited good correlation with S209PCBs (R2 ¼ 0.957; p < 0.001; n ¼ 15) and a multiplication factor of 1.4 ± 0.1 can be used to extrapolate total PCBs from ai-PCBs (see details in Table S7). Indicator PCBs accounted for only 2.9%e12% (average 5.8%) of total PCBs in the air samples, although these congeners have been suggested to exhibit about 20%e30% of total

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inconsistent comparisons, suggesting the need for a systematic study (e.g., a PUFePAS monitoring network) to characterize spatiotemporal trend and potential sources of this emerging compound.

3.3. Concentrations and profiles of PBDEs and other BFRs

Fig. 1. Concentrations and profiles of PCBs in the air samples from northern Vietnam (Ʃ208PCBs and Di-CBs* indicate total PCBs and dichlorobiphenyls excluding CB-11, respectively).

PCBs in other environmental media and biological samples (Froescheis et al., 2000; Hoai et al., 2010; Klees et al., 2015; Anh et al., 2019d), suggesting the need of a more relevant indicator PCB scheme for the air. A comparison on the occurrence of PCBs in the air from different locations in the world observed by the PUFePAS method is presented in Table S8. It is relatively difficult to facilitate relevant comparison because the numbers of monitored PCB congeners varied between studies and CB-11 is infrequently examined. PCB concentrations (including CB-11) in our air samples were within the range reported for indoor air of homes and outdoor air in some urban and rural areas in the US (Ampleman et al., 2015). PCB concentrations (excluding CB-11) detected in the present study (median 340; range 96e930 pg m3) were comparable to those found in several urban and rural sites from Argentina (Tombesi et al., 2014; Pegoraro and Wannaz, 2019) and Ghana (Hogarh et al., 2018); but still lower than levels measured in Turkey (Cetin et al., 2017a), India (Pozo et al., 2011; Devi et al., 2014), and China (Hogarh et al., 2012). Because PUFePAS studies for monitoring of CB-11 in the air are relatively scarce, we compared our results with data from both passive and active sampling methods for this unique congener. Concentrations of CB-11 found in our samples (68e300 pg m3; sampling year 2013e2015) were within the ranges documented in the ambient air of Jeonju, South Korea (14.67e244.19 pg m3; HVAS; Kim et al., 2003) and a background site in Ningbo, China (14e235 pg m3; HVAS; 2012e2015; Mao et al., 2019). Our detected values were higher than those measured in northern Japan (1.9e37 pg m3; LVAS; 2005e2011; Anezaki and Nakano, 2014); in Philadelphia metropolitan area, US (4e44 pg m3; PUFePAS; 2005; Du et al., 2009); and in Antarctica (5.16e31.4 pg m3; PUFePAS; 2009e2010; Li et al., 2012). However, the differences in sampling methods and time of sampling make

Concentrations of S36PBDEs in our air samples ranged from 5.3 to 86 (median 11) pg m3 with BDE-209 as the most predominant congener (median 7.4; range 3.5e80 pg m3) that accounted for 67% ± 12% of total PBDEs (Fig. 2). The highest PBDE concentrations were found in the air around a rubber melting kiln (RM-1; 86 pg m3) and in an urban house (UIeH3; 42 pg m3). The difference in PBDE concentrations between the urban and waste processing sites was not statistically significant. Levels of PBDEs in the air samples from Hanoi urban area of this study (23; 10e42 pg m3) were quite similar to those recorded in some urban and suburban sites of Hanoi in 2008 (4.6e58 pg m3; Tue et al., 2013). Elevated concentrations of PBDEs were measured in some e-waste recycling workshops in northern Vietnam (620e720 pg m3; Tue et al., 2013), which were about one to two orders of magnitude higher than levels detected in the ELV processing sites of this study. Concentrations of BB-153 were lower than the MDL of 0.050 pg m3 in all the samples. PBEB and BTBPE were respectively found in 9 and 4 out of the 15 samples at low concentrations (<0.050e0.32 and < 0.050e0.66 pg m3, respectively). DBDPE was detected at considerable levels in a rubber melting kiln (RM-1; 51 pg m3) and an ELV workshop (ELV-2; 11 pg m3). Concentrations of PBDEs in the Vietnamese air were generally lower than those reported for different functional areas in Bloomington, US and Toronto, Canada (Vernier et al., 2016); Guangzhou,

Fig. 2. Concentrations and profiles of BFRs in the air samples from northern Vietnam (Ʃ35PBDEs indicates di-to nona-BDEs; other BFRs includes PBEB, BB-153, BTBPE, and DBDPE).

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China (Ding et al., 2016); Chennai, New Delhi, Mumbai, and Kolkata,  rdoba, Argentina (Pegoraro and India (Chakraborty et al., 2017); Co Wannaz, 2019); and Dilovasi, Turkey (Cetin et al., 2019) (Table S9). Our values were comparable to or higher than levels found in the urban air of Nepal (Yadav et al., 2017), Kuwait (Gevao et al., 2006), and Czech Republic (Vernier et al., 2016); and remote area of King George Island, Antarctica (Li et al., 2012). Previous studies have indicated e-waste processing activities as potential sources of PBDE emissions into the atmosphere (Muenhor et al., 2010; Tue et al., 2013; Xu et al., 2015); however, information about the release of PBDEs from ELV processing sites is relatively limited. Cahill et al. (2007) reported that outdoor air concentrations (gas and particulate) of PBDEs around an automotive shredding and metal recycling (ASMR) facility in California, US were 810 and 390 pg m3 in “normal operation” and “no activity” periods, respectively. High concentrations of PBDEs (mean 326; range 167e510 pg m3; HVAS) were also found in the vicinity of a large ASMR facility in Brisbane, Australia (Hearn et al., 2013). Air concentrations of PBDEs in the workshops of our study were much lower than those reported for some ASMR facilities in developed countries, mainly due to lower scale and productivity of the Vietnamese ones. Besides, the difference in sampling methods (PUFePAS vs. HVAS) with different particle sampling effectiveness may affect analytical results, especially for highly brominated compounds.

3.4. Concentrations and profiles of PAHs and Me-PAHs Concentrations of PAHs (median 29,000; range 5100e100,000 pg m3) were about 2e10 times higher than those of Me-PAHs (6000; 1000e33,000 pg m3). Levels of both PAHs and Me-PAHs in the air around the waste processing sites were significantly higher than the urban sites (p < 0.05). Information about bulk air (gas and particle) concentrations of PAHs in Vietnam is still limited, especially for residential areas and working places. Concentrations of 47 PAHs (including Me-PAHs) in several roadside sites in Hanoi were 480,000 ± 300,000 and 63,000 ± 82,000 pg m3 for gaseous and particulate phases, respectively, suggesting traffic activities as important sources of PAH emission in Vietnamese urban areas (Kishida et al., 2008). Hong et al. (2016) reported that concentrations of 16 PAHs and 26 alkylated PAHs in the Vietnamese air were 42,800 (15,200e87,000) and 26,500 (8030e79,900) pg m3, respectively. To our knowledge, there is no study to investigate air concentrations and profiles of PAHs and Me-PAHs in living and working areas in Vietnam. For unsubstituted PAHs, almost all the urban samples (except for UO-M) and a plastic recycling facility (PR) showed the abundance of phenanthrene (Phe) and fluorene (Flu); whereas samples from other waste processing sites and a mixed land use area in Hanoi were dominated by Phe, fluoranthene (Flt), and pyrene (Pyr) (Fig. 3). For Me-PAHs, methylated derivatives of anthracene (Ant) and Phe were the major compounds in all the samples. Toxic equivalents to benzo [a]pyrene (BaP-EQs) in our air samples ranged from 20 to 910 (median 260) pg m3 with the most important contributors as BaP, dibenz [a,h]anthracene (DA), and benz [a]anthracence (BaA). The occurrence of PAHs in the air from different locations in the world observed by the PUFePAS method is summarized in Table S10. It is interesting that air concentrations and profiles of PAHs were relatively uniform, regardless of geographic variation. Previous studies have indicated that PAHs are more concentrated in the air of industrial and urban areas (Santiago and Cayetano, 2007; Pozo et al., 2012; Alvarez et al., 2016; Cetin et al., 2017a, b; Pegoraro and Wannaz, 2019). It should be noted that the workshops of this study are located in rural or even remote areas, and the abundance

Fig. 3. Concentrations and profiles of PAHs and Me-PAHs in the air samples from northern Vietnam.

of PAHs in the air of these sites is closely related to ELV processing activities. PAHs containing 3 and 4 fused rings such as Phe, Flu, Flt, Pyr, and chrysene were the most dominant compounds detected in the air. Naphthalene and its derivatives are more volatile and concentrations of these compounds are omitted in our work as well as in many earlier studies, probably due to immediate equilibrium and/or significant blank levels. In addition, Me-PAHs are rarely monitored, but their concentrations are elevated in some ELV workshops and oil refining facilities of this study. The abundance of alkylated PAHs as compared to the parent compounds was also observed in the air from several Asian countries such as China, India, Japan, South Korea, and Vietnam (Hong et al., 2016). Further comprehensive and detailed studies are needed to characterize the occurrence of highly volatile PAHs and substituted PAHs, which may exhibit high abundance but are infrequently monitored.

3.5. Potential emission sources Log-transformed concentrations of PCBs, PBDEs, PAHs, and MePAHs were analyzed by Pearson correlation analysis to evaluate their potential emission sources (Table S11). Mono-CBs and tetra-to hepta-CBs, including dl-PCBs, were highly correlated; whereas CB-

H.Q. Anh et al. / Chemosphere 247 (2020) 125991

11 exhibited moderate correlation with other di-CBs and tri-CBs only. CB-11 together with other low-chlorinated PCBs have been identified as impurities in paint pigment manufacturing and they were found in various consumer products (Rodenburg et al., 2010; Anezaki and Nakano, 2014; Guo et al., 2014; Shang et al., 2014; Vorkamp, 2016). The abundance of CB-11 in the air samples of this study, especially for the urban air, suggests their emission sources from color-printed products and decorative paints. Proportions of medium chlorinated PCBs (i.e., tetra-to hepta-CBs) were higher in the waste processing sites compared to residential areas, indicating current emissions from improper treatment of PCB-containing materials such as capacitor and transformer oils, heat transfer fluids, lubricants, adhesives, sealants, rubber products, and wire and cable coatings (Erickson and Kaley II, 2011). Our previous study also found that floor dust from similar ELV workshops in Bac Giang was contaminated by technical PCB mixtures rich in penta- and hexa-CBs such as Aroclor 1254, Kanechlor 500, or Sovol (Takahashi et al., 2017). PAHs and Me-PAHs were strongly correlated and their profiles suggest the dominance of pyrogenic sources compared to petrogenic sources. The ratios of Flt/(Flt þ Pyr) from 0.50 to 0.77 (average 0.57) indicate that atmospheric emissions of PAHs in these locations were mainly attributed to biomass, wood, and coal combustion (Yunker et al., 2002). In some ELV workshops and oil refining facilities, elevated concentrations of Me-PAHs were found with potential petrogenic sources such as fuels, lubricants, and engine oils, which are applied in different vehicular parts, together with their sources from bituminous coals, crude oils and refined products (Zakaria et al., 2002; Richter-Brockmann and Achten, 2018). PCBs were applied as lubricants and petroleum additives (Erickson and Kaley II, 2011). Lubricating oils in automobiles were also estimated as potential sources of PCBs in soils, sediments, and road dusts from Hanoi urban area (Toan et al., 2007; Hoai et al., 2010; Anh et al., 2019d). It is interesting that Me-PAHs showed significant correlation with several PCB homologs (e.g., mono- and tetra-to hepta-CBs), probably due to their co-existence in lubricating oils and other petroleum products. However, further confirmation by

7

analyzing PCBs and PACs in source-related samples such as vehicular exhaust, industrial emissions, and petroleum products, is needed. PBDEs did not show any correlation with PCBs and PAHs, implying their widespread and specific applications in electrical and electronic equipment, construction materials, and vehicles (Li et al., 2017). Based on these results of source apportionment, we found that primitive ELV processing activities in Vietnam are potential sources of several classes of organic micro-pollutants, notably technical PCBs and PAH-related compounds. The recycling technology and waste management practices in these informal facilities should be improved, e.g., prohibition of uncontrolled burning, collection and disposal of waste oils and other hazardous materials by appropriate systems or by professional waste treatment services. 3.6. Implications for human exposure The inhalation daily intake doses (IDs) of PCBs, PBDEs, and PAHs were estimated for adults in the urban area and workers in the waste processing sites based on measured concentrations of these pollutants with an inhalation rate of 16 m3 d1 (US EPA, 2011) and an average body weight of 60 kg. In the waste processing workshops, time fractions 8/24 and 16/24 were applied for working and living areas, respectively. The median values and ranges of IDs of total PCBs and selected PBDEs and PAHs are summarized in Table 2, in the comparison with respective reference doses (RfD). The IDs of total PCBs estimated for urban adults and workers were comparable, but the workers may receive higher doses of dl-PCBs and WHOTEQs. The IDs of PBDEs and low-molecular-weight PAHs (e.g., acenaphthene, anthracene, and fluorene) were not significantly different between urban adults and workers. However, IDs of some high-molecular-weight PAHs such as Flt, Pyr, and BaP were one to two orders of magnitude higher in workers compared to normal adults. In our previous study, we estimated median ID of 16 ng kg1 d1 of PAHs and Me-PAHs in settled dust for ELV dismantling workers (Anh et al., 2019c). Based on our data, a median ID of 12 ng kg1 d1 of PAHs and Me-PAHs was estimated for the

Table 2 Inhalation daily intake doses (pg kg1 d1) of PCBs, PBDEs, and PAHs estimated for urban adults and workers in waste processing areas, northern Vietnam. Compounds

Urban adults Median

PCBs WHO-TEQs BDE-47 BDE-99 BDE-153 BDE-209 Anthracene Acenaphthene Fluorene Fluoranthene Pyrene Benzo [a]pyrene a b c d e f g h i j k l m

1.3  1.3  1.2  n.a. a n.a. 4.8  1.2  1.0  6.1  2.7  1.9  7.2 

102 105 101

100 102 102 102 102 102 100

Workers in waste processing areas Min 6.1  7.5  9.1  n.a. n.a. 1.3  8.9  3.9  4.0  1.5  4.5  n.a.

Max 101 106 102

100 101 101 102 102 101

2.2  3.5  3.5  1.6  n.a. 7.5  2.2  1.9  9.5  2.4  2.1  4.9 

Median 102 103 101 101 100 102 102 102 103 103 101

7.5  1.9  1.3  n.a. n.a. 1.9  3.0  7.2  5.5  2.1  1.6  2.4 

Not available because of non-detected compounds. RfD proposed by the US Agency for Toxic Substances and Disease Registry (ATSDR, 2000). Tolerable daily intake proposed by the World Health Organization (Van Leeuwen et al., 2010). RfD based on neurobehavioral effects (US EPA, 2008a). RfD based on neurobehavioral effects (US EPA, 2008b). RfD based on neurobehavioral effects (US EPA, 2008c). RfD based on neurobehavioral effects (US EPA, 2008d). RfD for no observed effects (US EPA, 1990a). RfD based on hepatotoxicity (US EPA, 1990b). RfD based on hematological effects (US EPA, 1990c). RfD for effects on hepatic and urinary system (US EPA, 1990d). RfD for effects on urinary system (US EPA, 1990e). RfD based on neurobehavioral changes (US EPA, 2017).

101 103 101

100 102 101 102 103 103 101

Min 5.7  9.3  1.2  n.a. n.a. 1.6  1.8  5.7  3.9  1.2  9.2  1.6 

RfD Max

101 104 101

100 102 101 102 103 102 101

1.3 7.9 3.6 1.3 4.7 8.4 4.9 1.0 7.8 3.1 2.9 3.6

           

102 103 101 101 102 100 102 102 102 103 103 101

2  104 1e4 c 105 d 105 e 2  105 7  106 3  108 6  107 4  107 4  107 3  107 3  105

b

f g h i j k l m

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workers, suggesting that inhalation of contaminated air is an important source of human non-dietary exposure to PAH-related compounds. In general, the IDs of airborne PCBs, PBDEs, and PAHs in the studied locations of this study were much lower than the RfDs (Table 2), implying insignificant non-cancer risk. However, it should be noted that these RfD values were originally proposed for oral exposure and the reference concentrations (RfC) and quality guidelines for the air environment of almost all the compounds investigated here are still unavailable. The US Environmental Protection Agency proposed a RfC of 2000 pg m3 for BaP based on developmental effects (i.e., decreased embryo/fetal survival) (US EPA, 2017). Air concentrations of BaP (n.d. e 245 pg m3) or even BaP-EQs (20e910 pg m3) in our study were lower than this RfC. The World Health Organization established BaP concentrations for excess lifetime cancer risk of 106, 105, and 104 as 12, 120, and 1200 pg m3, respectively (WHO, 2000). Based on these criteria, BaP-EQ concentrations in several waste processing workshops of this study may exhibit risk level over 105, partially indicating a potential cancer risk (WHO, 2008; Health Canada, 2010). In addition, recent studies have focused on human exposure and toxic effects of low-chlorinated and pigment-derived PCBs (Ampleman et al., 2015; Takeuchi et al., 2017; Pencikova et al., 2018). CB-11 was among the most dominant congeners that inhaled in greatest amounts by mothers in urban and rural cohorts in East Chicago, Indiana and Columbus Junction, Iowa, US (Ampleman et al., 2015). CB-11 was found as major PCB congener in human serum (Koh et al., 2015) and its hydroxylated metabolites (e.g., 4-OH-CB-11) can induce a significantly higher estrogen receptor mediated activity as compared to the parent compound (Pencikova et al., 2018). Other low-chlorinated PCBs (e.g., CB-1, -8, -28, 31, 35) have been identified as aryl hydrocarbon receptor (AhR) agonists (Takeuchi et al., 2017; Pencikova et al., 2018). This situation suggests an urgent need to conduct a more comprehensive monitoring and risk assessment study on PAH-related compounds and low-chlorinated PCBs, especially for highly urbanized and informal waste processing areas in developing and newly industrialized countries like Vietnam. More attention should be paid to promote human awareness and health protection measures against toxic effects of airborne micro-pollutants in such areas. 4. Conclusions and perspectives This study is among the first to create a comprehensive and detailed picture on typical organic micro-pollutants, including both legacy and emerging contaminants, in the air samples collected from urban and vehicular waste processing areas in northern Vietnam, by using a cost-effective and highly applicable monitoring scheme. Our results have revealed the widespread occurrence of PCBs, PBDEs, PAHs, and Me-PAHs in the Vietnamese air with specific sources of technical and dioxin-like PCBs and PAH-related compounds from ELV processing activities, and pigment-derived PCBs from human activities in the urban area. Regarding the abundance and toxicity of low-chlorinated PCBs and highly volatile PACs, additional analytical efforts are needed to improve sampling protocols (e.g., determination of sampling rate and linear uptake phase by the PUFePAS), chemical analysis (e.g., combination of screening, targeted, and non-target analysis), and toxicity evaluation by using in vitro bioassays. The inhalation daily intake doses of organic micro-pollutants in the studied areas is generally lower than the reference doses, except for concentrations of BaP in few locations (e.g., a mixed land use area in Hanoi, an ELV workshop, and a rubber melting kiln) may exceed critical level of 120 pg m3 corresponding to cancer risk over 105. Further intensive studies on airborne micro-pollutants should be conducted in Vietnam, a

typical newly industrialized country with increasing concerns about air pollution. CRediT authorship contribution statement Hoang Quoc Anh: Methodology, Formal analysis, Writing original draft. Isao Watanabe: Methodology, Formal analysis. Nguyen Minh Tue: Conceptualization, Formal analysis. Le Huu Tuyen: Conceptualization, Resources. Pham Hung Viet: Conceptualization, Resources. Ngo Kim Chi: Conceptualization, Resources. Tu Binh Minh: Conceptualization, Resources. Shin Takahashi: Supervision, Conceptualization, Resources, Writing - review & editing. Acknowledgements This study was supported in part by Grants-in-Aid for Scientific Research (B: 16H02963) and Fund for the Promotion of Joint International Research (Fostering Joint International Research (B)) (18KK0300) from the Japan Society for the Promotion of Science (JSPS); and the Environment Research and Technology Development Fund (3K153001) from the Japanese Ministry of the Environment. We would like to thank all the people who supported us in sampling activities. Appendix A. Supplementary data Supplementary data to this article can be found online at https://doi.org/10.1016/j.chemosphere.2020.125991. References Alvarez, A.L., Pozo, K., Paez, M.I., Estellano, V.H., Llanos, Y., Focardi, S., 2016. Semivolatile Organic Compounds (SVOCs) in the atmosphere of Santiago de Cali, Valle del Cauca, Colombia along north-south transect using polyurethane foam disk as passive air samplers. Atmos. Pollut. Res. 7, 945e953. Ampleman, M.D., Martinez, A., DeWall, J., Rawn, D.F.K., Hornbuckle, K.C., Thorne, P.S., 2015. Inhalation and dietary exposure to PCBs in urban and rural cohorts via congener-specific measurements. Environ. Sci. Technol. 49, 1156e1164. Anezaki, K., Nakano, T., 2014. Concentration levels and congener profiles of polychlorinated biphenyls, pentachlorobenzene, and hexachlorobenzene in commercial pigments. Environ. Sci. Pollut. Res. 21, 998e1009. Anh, H.Q., Tomioka, K., Tue, N.M., Tri, T.M., Minh, T.B., Takahashi, S., 2018. PBDEs and novel brominated flame retardants in road dust from northern Vietnam: levels, congener profiles, emission sources and implications for human exposure. Chemosphere 197, 389e398. Anh, H.Q., Tomioka, K., Tue, N.M., Tuyen, L.H., Chi, N.K., Minh, T.B., Viet, P.H., Takahashi, S., 2019a. A preliminary investigation of 942 organic micropollutants in the atmosphere in waste processing and urban areas, northern Vietnam: levels, potential sources, and risk assessment. Ecotoxicol. Environ. Saf. 167, 354e364. Anh, H.Q., Tomioka, K., Tue, N.M., Suzuki, G., Minh, T.B., Viet, P.H., Takahashi, S., 2019b. Comprehensive analysis of 942 organic micro-pollutants in settled dusts from northern Vietnam: pollution status and implications for human exposure. J. Mater. Cycles Waste Manag. 21, 57e66. Anh, H.Q., Tue, N.M., Tuyen, L.H., Minh, T.B., Takahashi, S., 2019c. Polycyclic aromatic hydrocarbons and their methylated derivatives in settled dusts from end-of-life vehicle processing, urban, and rural areas, northern Vietnam: occurrence, source apportionment, and risk assessment. Sci. Total Environ. 672, 468e478. Anh, H.Q., Watanabe, I., Tomioka, K., Minh, T.B., Takahashi, S., 2019d. Characterization of 209 polychlorinated biphenyls in street dust from northern Vietnam: contamination status, potential sources, and risk assessment. Sci. Total Environ. 652, 345e355. Anh, H.Q., Minh, T.B., Tran, T.M., Takahashi, S., 2019e. Road dust contamination by polycyclic aromatic hydrocarbons and their methylated derivatives in northern Vietnam: concentrations, profiles, emission sources, and risk assessment. Environ. Pollut. 254, 113073. Atsdr, 2000. Toxicological Profile for Polychlorinated Biphenyls (PCBs). Accessed date. (Accessed 30 May 2019). https://www.atsdr.cdc.gov/toxprofiles/tp17.pdf. Bao, L.J., Zeng, E.Y., 2014. Field application of passive sampling techniques for sensing hydrophobic organic contaminants. Trends Environ. Anal. Chem. 1, e19ee24. Bidleman, T.F., Melymuk, L., 2019. Forty-five years of foam: a retrospective on air sampling with polyurethane foam. Bull. Environ. Contam. Toxicol. 102,

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