Interactive comment on “ Particle-size distribution of polybrominated diphenyl ethers ( PBDEs ) and its implications for health ”

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Introduction
Polybrominated diphenyl ethers (PBDEs) are a class of organobromine compounds that are widely used as flame retardant.They are applied to a broad array of textiles and consumer products including plastics, polymers, mattresses, upholstery, carpeting, building materials and electronic equipment (de Wit, 2002;Alaee et al., 2003).Because the compounds are additive rather than chemically bound to the products, they can be released into the environment.They are persistent organic chemicals and can bioaccumulate, therefore, they have become contaminants detectable in the environment, in animals, and in humans (Su et al., 2009;Besis and Samara, 2012).Human uptake is thought to be through inhalation, dermal absorption and consumption of contaminated food (Marklund et al., 2003;Wensing et al., 2005).The primary source of exposure to humans is believed to be consumption of contaminated fish, poultry, meat and dairy products (Su et al., 2009;Besis and Samara, 2012).Occupational exposures may occur in computer, electronic warehouses and formulation facilities (Harrad et al., 2010).Inhalation exposure can take place through ambient aerosol or dust containing PBDEs.
Compared with dust PBDEs, the inhalation of ambient aerosols may be a minor pathway for human people, but it has a long-term bioaccumulation process in human body.
When PBDEs are suspended in air, they can be present as particles.Since we can not say how long PBDEs remain in the air, long-term exposure to PBDEs has a greater potential to cause health effects than does short-term exposure to low levels because of their tendency to build up in your body over many years.Growing concerns about the health impacts of PBDEs have led to decline in their production and finally banned to use in the US and Europe since 2004 (Kemmlein et al., 2009).All technical mixtures of PBDEs were also totally phased out in other regions including China (Betts, 2008).However, it is likely that long-term exposure will continue long after PBDE production has ended through emissions from PBDE-containing products that are still being used.Thus, it becomes necessary to investigate the particulate PBDEs characteristic existing in urban ambient aerosols.
Over the past decade, measurements on atmospheric PBDEs have been carried out in various areas around the world such as Turkey (Cetin and Odabasi, 2008), Japan (Kakimoto et al., 2014), Thailand (Muenhor et al., 2010) and China (Yang et al., 2013) in Asia, USA (Hale et al., 2003) and Canada (Wilford et al., 2004) in North America, Greece (Besis et al., 2015), France (Castro-Jimenez et al., 2011) and Czech (Okonski et al., 2014) in Europe and someplace in Arctic (Moller et al., 2011;Wang et al., 2005).In these studies, the particulate PBDEs were mainly investigated in the individual particle size fractions such as PM 2.5 and PM 10 , and rarely involved with size-resolved particles.Particle-size distribution of PBDEs was crucial when evaluating human health risks since the size-resolved particles dominated deposition behavior of particles in the respiratory tract.To the best of our knowledge, particle-size distribution of PBDEs Introduction

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Full was merely reported in Thessaloniki, Athens (Greece) (Mandalakis et al., 2009;Besis et al., 2015), e-waste recycling sites in close to Guangzhou (China) (B.Z. Zhang et al., 2012;Luo et al., 2014b), and Brno (Czech) (Okonski et al., 2014).These studies showed that major lighter brominated congeners existed on coarse particles, while most highly brominated congeners occurred in fine particles.In association with the fact that fine particles can easily penetrate/enter the alveolar region, fine particle-bound highly brominated congeners can travel deep into the lungs and cause serious health problems for human people (Geiser et al., 2005).To clarify this issue we first need to investigate the actual particle-size distribution of PBDEs through a long-term observation.
This investigation was conducted in urban Shanghai with the aim to evaluate the contribution of particle-size distribution of PBDEs and their deposition in human respiratory tract.Besides this, the elucidation of influence of some factors such as the volatility, the chemical affinities and releasing source onto these distributions was also attempted.

Chemicals
Standard mixtures of PBDEs

Sample collection
The sampling campaign took place on the rooftop (20 m above the ground) of No. 4 teaching building at Fudan University campus (121.50 • E, 31.30• N), approximately 5 km northeast of downtown Shanghai city (elevation about 4 m a.s.l.).A Fudan super monitoring station for atmospheric chemistry was running all year round.The site was in close proximity to digital malls, residences and the traffic around was busy due to close to sub-downtown.The main releasing sources at this site included industries emission, household heating and road transport.Details regarding the sampling site were included in our previous work (P.F. Li et al., 2011;X. Li et al., 2011).Particle samples were collected by drawing air through quartz fiber filter (Whatman QMA, ∅81 mm), using an Anderson 8-stage air sampler (Tisch Environmental Inc., USA).The flow rate was controlled at 28.3 L min −1 .The cutoff aerodynamic diameters for each stage were < 0.4, 0.4-0.7,0.7-1.1,1.1-2.1, 2. 1-3.3, 3.3-4.7, 4.7-5.8, 5.8-9.0 and > 9.0 µm.The whole observing period ranged from December 2012 to November 2013.The sampling time was 120 h for each sample batch.A total of 189 particle samples (21 sample batches containing 9 size fractions) were obtained at this site.Prior to sampling, the filters that wrapped in aluminum foils were baked at 450 • C for 12 h to remove organic materials.After sampling, loaded filters together with aluminum foils were stored at −20 • C until extraction.In addition, meteorological data during the sampling period was obtained from the Fudan atmospheric monitoring station (Lv et al., 2015).

Sample extraction
The aerosol samples were extracted by Soxhlet with a mixture of dichloromethane/hexane (1 : 1, ν/ν).The extraction time was 36 h at a constant temperature 69 • .After extraction, the samples were filtered through 0.45 µm PTFE syringe filters and concentrated using a rotary evaporator (BÜCHI Rotavapor, Switzerland) and a pure N 2 stream.Before instrumental analysis, 13 C-BDE-28, -47, -99 and 153 were added to the final extract as internal standards.Introduction

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Full  • C. High-purity Helium (99.999 %) was applied as carrier gas with a constant flow rate of 1.2 mL min −1 in column.Nitrogen gas was used as collision gas in MS.The identification and quantification of PBDEs were done according to retention times, selected precursor ions, product ions and the internal standard method relative to the closest eluting PBDE surrogate.The calibration solutions were prepared at five concentrations and contained uniform concentrations of the internal standards.For each analyte, a relative response factor was determined for each calibration level using the internal standard.The five response factors were then averaged to produce a mean relative response factor for each species.Reported analyte concentrations were corrected for internal standards recoveries.The calibration curves showed a linear response in the range 0.1-5 µg L −1 .The correlation coefficients of the calibration curve for the different PBDEs were R 2 > 0.99.

Quality control and assurance
Each batch of samples included one procedural blanks.In that case, only BDE-71, 100, 154 and 190 were commonly detected at much lower levels (< 5 %) in some samples.
The mean values of blanks were then subtracted from measured values of each sample.Method recoveries determined by spiking the sampling process (five replicates) with a standard mixture of PBDEs ranged from 75 to 175 %.The recoveries of the Introduction

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Full isotopically-labelled PBDEs internal standard were higher than 70 % for all the samples (refer to SRM 2585, NIST, Gaithersburg, MD).Repeatability was evaluated by performing four analyses of a standard PBDEs solution containing the above mentioned PBDEs and the surrogate standards in the same conditions.The relative standard deviations (RSD) of the relative response factors were below the 10 % for all PBDEs.
The method detection (DL) and quantification limits (QL) were calculated as the concentrations equivalent to three and ten times the noise of the quantifier ion for a blank sample (DL ranged from 0.05 to 0.6 pg m −3 ).For the purpose of statistical analysis, samples with concentrations under LOD were assigned concentrations equal to 0.5 LOD (Okonski et al., 2014).

Size-specific gas/particle partition
Two processes are commonly accepted for illustrating mechanisms of particle-gas partition, i.e., adsorption and absorption process.In the case of adsorption, it assumes that chemicals adsorb to active sites on the surface of the particle.The gas-particle partitioning coefficients (K p-ads ) during the adsorption process is described by (Pankow, 1987): where N s is the surface concentration of sorption sites (4 × 10 −10 mol cm −2 ), A TSP is the specific surface area of the particles, T is the ambient temperature (292 K), R is the ideal gas constant (8.31 J mol −1 K −1 ), Q L , Q v are respectively the enthalpy of desorption from the surface and the enthalpy of vaporization of the subcooled liquid (kJ mol −1 ), p • L is the vapor pressure of the subcooled liquid.In contrast with PAHs, similar situation was assumed for PBDEs that Q L − Q v ≈1× 10 4 J mol −1 (Aubin and Abbatt, 2006).After logarithmic transformation on both sides in Eq. ( 1), we can get the followed Eq. ( 2 Size dependent A TSP adopted from the results of Yu and Yu (2012) and the data was list in Table 2. Based on three modes, we then obtained Eq. ( 3) which derived from Eq. (2). log The temperature dependent p • L values of PBDE congeners was calculated using the regression parameters by (log (Tittlemier et al., 2002).In our study, the average temperature of sampling campaign was 292 K.The temperature dependent log p • L was list in Table 1.In the case of absorption, it assumes that atmospheric aerosols are coated with an organic film and chemicals can absorb into this organic phase.The gas-particle partitioning coefficients (K p-abs ) during the absorption process is described by Finizio et al. (1997): where M o and M OM are the mean molecular weights of octanol and the organic matter phase (g mol −1 ), γ o and γ OM are the activity coefficients of the absorbing compound in octanol and in the organic matter phase, respectively.f OM is the fraction of organic matter phase on particles, K OA is octanol-air partition coefficient.ρ OM is the density of octanol (820 kg m −3 at 20 • ).
With the assumption that = 1, the Eq. ( 4) can be simplified to Eq. ( 5) after logarithmic transformation on both sides: log K p-abs = log K OA + log f OM − 11.91 (5) Introduction

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Full Size-specific f OM was adopted from Yu and Yu (2012) and list in Table 2. Through calculation, we can deduce Eq. ( 6) as followed: K OA has been reported as functions of temperature (log (Harner and Shoeib, 2002).In this study, the average temperature of sampling campaign was 292 K.
The temperature dependent log K OA was list in Table 1, along with other physiochemical properties of the target PBDE congeners.Pankow (Pankow, 1994a) proposed a definition of the measured particle-gas partition coefficient ( K p-measured ) to characterize the partitioning behavior of semi-volatile organic compounds (SVOCs) between the gas and particulate phases Where P and G are PBDEs in particulate and gas phase concentration, respectively, and TSP is the total suspended particulate (µg m −3 ).After linear regression between In association with Eqs. ( 4), ( 6) and ( 7), we can investigate the sorption mechanisms governing particle-size distribution of PBDEs by comparing theoretical K p-ads and K p-abs with measured K p-measured .

Human respiratory risk assessment
Because the size-resolved particles played a key role in health risk assessment through inhalation (Luo et al., 2014a), we adopted a so-called International Commission on Radiological Protection (ICRP) model ( International human respiratory tract.The human respiratory tract can be divided into three regions: head airway (HA), tracheobronchial region (TB), and alveoli region (AR).The particle deposition efficiency (DE) in HA, TB and AR are estimated by the followed simplified Eqs. ( 8), ( 9) and ( 10): 1 + e 6.84+1.183ln D p + 1 1 + e 0.924−1.885ln D p (8) Where D p is diameter of the particle, and IF is inhalable fraction of all particles, The deposition flux (DF, pg h −1 ) of inhaled particulate PBDEs in respiratory tract is estimated by Where DE i is the particle deposition efficiency in each region for Dp i (the average diameter of each particle size fraction); C i is PBDEs concentration in particle Dp i (pg m −3 ); and V is the breathing rate.The lower and upper limit diameters of particles in this research were assumed to be 0.1 and 30 µm, respectively.The respiration rate under normal conditions was considered as 0.45 m 3 h −1 (K.Zhang et al., 2012).
In addition, we applied hazard quotient (HQ) values to assess non-cancer risk of size-resolved PBDEs through inhalation.The formula is as follows: where DI is daily intake (pg day −1 ) and calculated by multiplying deposition flux (DF: pg h −1 ) with average exposure time (ET: h day −1 ), BW is mean body weight of adult (60 kg) and RfD is reported oral reference dose for PBDEs (pg kg −1 bw day −1 ).In order to understand the impact of risk and uncertainty in size-resolved particles, we used Monte Carlo simulations to produce probability distributions of hazard levels with five thousand trials.Moreover, we used the SPSS version 22.0 (IBM company, Chicago, IL, USA) to perform Pearson correlation analysis for all data and considered p values of smaller than 0.01 or 0.05 statistically significant.

PBDEs occurrence and seasonal variation
Most PBDE congeners were detected in the vast majority of samples (Fig. 1 and Table 3).100,154 and 190 were sometimes present close to the detection limits of the method.Due to the erratic concentration of BDE 209, this compound has been removed from further analysis.The box plot in Fig. 1 summarizes the concentrations measured throughout the year and allows for easy visualization of PBDE congener groups (e.g., tri-, tetra-, penta-, hexa-and hepta-BDEs).The box contains the middle 50 % of the data, whereas the top and bottom end of the box represent the 75th and 25th percentiles of the data set, respectively.The extensions ("whiskers") at either end of the box indicate the 95 and 5 percentile and the solid spheres represent the maximum and minimum values.The median concentrations are indicated by the solid vertical lines whereas the mean concentrations are depicted by the horizontal line.In general, the size of the box and the length of the whiskers are an indicator of the variability in concentrations at a given site for a given compounds containing the same number of bromine atoms.A small box shows that the distribution is uniform over the entire sampling period and vice versa.In these groups, penta-BDEs (49.5 ± 21.5 pg m −3 ) were the dominant congeners detected in all sam-Introduction

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Full ples, followed by hexa-BDE (16.7 ± 7.8 pg m −3 ), hepta-BDE (11.2 ± 3.1 pg m −3 ), tetra-BDE (5.9±1.3 pg m −3 ) and tri-BDE (1.3±0.3 pg m −3 ).Among individual PBDEs, PBDE-47, -99 and -85 were detected in 100 % of the ambient aerosol samples, with BDE-99 and -85 being the most dominant congeners.This may be due to the fact that less brominated BDEs have longer half-lives (years) and could be formed through debromination of more brominated congeners (Bezares-Cruz et al., 2004).The observed average concentrations of particulate Σ13PBDEs were ranged from 30.6 to 141.2 pg m −3 with a mean value of 86.3 pg m −3 (Table 3).This result was consistent during summer (30.6−82.2pg m −3 ) (Table 3).Higher concentrations in winter were at least in part due to increased emissions and by the distinctive meteorological conditions including reduced mixing heights and lower precipitation depth for favoring the pollutants accumulation in the atmosphere (Volckens and Leith, 2003).In addition, the adsorption of gaseous PBDEs on particles was likely to increase during winter since the partition coefficient, K p was inversely correlated with temperature (r = −0.867,p < 0.01) (see Fig. 2).Lower concentrations in summer may have been caused by wet scavenging since some summer sampling days experienced precipitation at this site.Seasonal variations in PBDEs can also be explained by the Asian monsoon patterns.Shanghai sites are situated in a transitional zone of the northern subtropical monsoon system, where the northwesterly winter monsoon bring polluted air masses from China Mainland, while the southeasterly summer monsoon bring cleaner oceanic aerosols comes from the oceans (Western Pacific) (Shi and Cui, 2012).Moreover, higher wind speeds appeared to be typically associated with lower PBDEs concentrations (r = −0.583,p < 0.01) (Fig. 2).Higher PBDEs concentrations were associated with higher PM 2.5 level (r = 0.629, p < 0.01) and lower visibility (r = −0.686,p < 0.01).This seasonal pattern were consistent with those measured in Huaniao Island (Li et al., 2015), Qingyuan (Tian et al., 2011) and Dongguan (Zhang et al., 2009).

PBDEs size distribution and process mechanism
Among the PBDE congeners measured, we chose BDE-47, -85, -99, -138, -153 and -183 for the study on size distribution due to detection frequencies higher than other congeners.Figure 3 plots the average size distributions of these PBDEs in the continuous smoothed curves inverted from the sample data.The results showed that particulate PBDEs exhibited a bimodal distribution with a mode peak in the accumulation particle size range and the second mode peak in the coarse particle size ranges.As the number of bromine atoms in the molecule increased, accumulation mode peak intensity increased while coarse mode peak intensity decreased, which indicated that the lighter Introduction

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Full brominated congeners BDE 47 and 85 were mainly associated with particles larger than 2.1 mm, whereas the highly brominated congeners were mainly sorbed to the fine particles.The similar bimodal distribution of PBDEs also occurred in Heraklion (Mandalakis et al., 2009), Brno and Telnice (Okonski et al., 2014), Guangzhou (B.Z. Zhang et al., 2012).Differences in the particle-size distribution of individual PBDEs could reflect differences in their emission sources but there was no credible scientific evidence in support of this claim.Although published data on particle-size distribution of PB-DEs are not available for comparison, analogous trends have also been observed for other classes of organic contaminants such as PAHs.Previous field measurements by cascade impactors demonstrated that more-ring PAHs were sorbed to the fine aerosol fraction, while more volatile or low-ring species were associated with larger particles (Wang et al., 2015;Kavouras et al., 1999;Kawanaka et al., 2004;Bi et al., 2005).The reason on these is due to the different volatility of PAHs, since more volatile species are absorbed to fine aerosol and distribute in coarse particles by rapid volatilization and condensation.On the contrary, for the more-ring PAHs, due to the lower vapour pressures, the time required for this repartitioning process is much longer (Bi et al., 2005), therefore, they tend to remain in fine particles initially emitted (Duan et al., 2007).This hypothesis can explain the relatively higher abundance of more volatile PAHs in the coarse particle mode.Similarly, it can be applied the particle-size distribution of PBDEs.To further confirm this hypothesis, the geometric mass diameter (GMD) for particulate PBDEs was calculated and correlated with logarithmic subcooled liquid vapor pressures (log P L ) (Fig. 4).The mean GMD values for all PBDE congeners was ranged from 1.9 to 2.9 µm in Shanghai, which was higher than those in Greece (0.14-0.63 µm) (Mandalakis et al., 2009) and Guangzhou (0.98-1.98 µm) (Luo et al., 2014b).Moreover, there are a positive moderate correlation between GMD and log P L (r = 0.69, p < 0.01), indicating the GMD increases as the volatile of PBDE congeners increases.This phenomenon becomes more apparent in coarse size fraction with a increased positive correlation (r = 0.75, p < 0.01) (right panel in Fig. 4).This result suggests that most coarse particle-bound PBDEs contain higher volatile species such as tri-and Introduction

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Full tetra-BDEs.They are derived from the secondary distribution process, i.e., re-volatilize from fine particles and re-condensate onto coarse ones (Wang et al., 2008;La Guardia et al., 2006).Moreover, chemical affinities also played an important role in PBDEs' distribution process.Theoretically, highly brominated congeners had strong hydrophobicity and prefer to bound with small particles because they had large surface areas (Venkataraman et al., 1999).Such an explanation, however, cannot adequately account for the PBDEs distribution patterns observed in the present study.Perhaps in fact other factors, e.g.emission sources, sampling sites and weather conditions (temperature and/or relative humidity) might also influence their distributions (Zielinska et al., 2004).Although there were still difficulties in totally clarifying the size distributing mechanism of PBDEs or other SVOCs now, it was important to integrate all factors in consideration in future study.

Preliminary study on PBDEs partitioning mechanisms
Usually, two major mechanisms, i.e., adsorption and absorption played the important role in PBDEs partitioning to multimodal urban aerosols (Lohmann and Lammel, 2004).To clarify these processes, the theoretical K p-ads and K p-abs were respectively calculated based on Eqs. ( 3) and ( 6) because they involved the size-specific parameters, usually including organic matter fractions and the available adsorptive sites on aerosol particles (Pandis et al., 1992;Pankow, 1994b).The obtained theoretical K p-ads and K p-abs values were estimated and compared with measured K p-measured (from Eq. 7).Since we had no gas PBDEs concentrations, the measured K p-measured were based on a recent study in shanghai by Yang et al. (2013).In their studies, both gas and particulate PBDEs concentration were reported at an urban site about ∼ 50 km away from our site (Table 1) with the theoretical K p-ads and K p-abs derived from adsorption and absorption in our study.Note that we took < 0.4µm, 0.4−2.1 µm and > 2.1µm for Aitken, accumulation and coarse mode, respectively.The plots of measured log K p-measured vs. log p • L and log K OA were presented in Fig. 5, along with two sets of theoretical K p-ads and K p-abs based on adsorption and absorption in three modes.As presented, significant linear correlations were found between measured log K p-measured and log p • L (R 2 = 0.76) as well as measured log K p-measured and log K OA (R 2 = 0.77).For the same class compounds under equilibrium conditions by either adsorption or absorption, the slope of log-log plots of K p and p • L was expected to be close to −1 (Pankow and Bidleman, 1992) and the slope of log-log plots of K p and K OA should be close to 1 (Finizio et al., 1997).However, more gentle regression lines (slopes: −0.53, 0.68) were detected (Fig. 5), similar to slopes reported in previous studies (Cetin and Odabasi, 2008;Yang et al., 2012).The deviations were possibly caused by kinetic limitations (non-equilibrium partition), thermodynamic limitations (lack of constancy in desorption) and additional sorption (Harner and Bidleman, 1998;Cousins and Mackay, 2001;Lohmann et al., 2007).
The three mode data sets of theoretical log K p-ads and log K p-abs in Fig. 5a and b were calculated using Eqs.( 3) and ( 6), considering only adsorption mechanism or absorption mechanism, respectively.As expected, the slopes for them were all −1.Both log K p-ads considering only adsorption and log K p-abs considering only absorption were compared with measured log K p−meaused in Fig. 5.The results showed that the K p-measured values of highly brominated congeners (e.g.,  in three modes fell into the regression line of the theoretical K p-ads .(Fig. 5a), while the measured K p-measured values of lighter brominated congeners (e.g., BDE-17, −28, −47 and −66) in three modes fell into the regression line of the theoretical K p-abs (Fig. 5b).These facts revealed that adsorption on surfaces of particles appeared to be responsible for bimodal distribution of highly brominated congeners, while absorption into organic matter seemed to play an important role for lighter brominated congeners.In addition, we can find the measured K p-measured lines are respec-Introduction

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Full tively close to the theoretical K p-ads line in Aitken regression lines (Fig. 5a) and the theoretical K p-abs lines in the accumulation and coarse regression lines (Fig. 5b).This meant that the mechanisms controlling the particle-size distribution of PBDEs included adsorption to Aitken mode particles and absorption to accumulation and coarse mode particles.Adsorption is dependent on available aerosol surface area (A TSP ) and absorption on available aerosol organic mass (f OM ).Although A TSP and f OM could not be measured and empirical data of A TSP or f OM were adopted from references in this paper, we did provide a way to investigate the mechanisms for size distribution of SVOCs from the view of gas-particle partition.

Correlation analysis of PBDEs
Table 4 presents a Pearson correlation matrix among PBDE congeners based on concentrations.Significant correlation was found among the tri-BDEs (BDE-17 and -28, r = 0.75, p < 0.05), as well as penta-BDEs (BDE-100, -99 and -85, r = 0.67−0.83,p < 0.05).These high correlation values suggested tri-BDEs and/or penta-BDEs shared a common source and/or exhibited a similar distribution behavior in environment.BDE-28 significantly correlated with BDE-71 (r = 0.65, p < 0.05), and both of them also significantly correlated with penta-BDEs 100, 99 and 85.As we known, penta-BDEs were frequently detected in ambient particles around solid waste incineration plants (Dong et al., 2015).In this observation, we also found that penta-BDEs appeared high concentrations (mean: 49.5 pg m −3 ) compared with other congeners.Thus, we concluded that these penta-BDEs and correlated congeners probable came from the same source regions because there were numerous of solid waste incineration plants located in surrounding places of Shanghai.Hepta-BDE (BDE-183 and -190) correlated poorly with the other congeners, with only two exception, which were statistically significant at the r = 0.47, p < 0.01 level for BDE-183 and -138, as well as r = 0.71, p < 0.05 level for , respectively.This suggested hepta-BDEs might not originate from same sources.Other studies showed there existed possibilities of decomposition of higher brominated PBDEs to form lower brominated PBDEs in atmosphere (Eriksson Introduction

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Full  , 2004;Söderström et al., 2004;Kajiwara et al., 2008).Here we did not measure higher brominated PBDEs (octa-and deca-BDEs) and no correlation analysis were performed on them.But as seen in Fig. 1, the hexa-vs.hepta-BDEs concentrations (mean: 16.7 and 11.2 pg m −3 ) were relatively high, thus indicative of multiple releasing sources in this area, i.e., local emission, higher brominated PBDEs breakdown and long-range atmospheric transport maybe had the potential contributions to hexa-and hepta-BDEs contaminations.

Implication for health
In this section, we calculated the regional deposition flux in human respiratory tract based on Eqs. ( 8), ( 9) and (10).Figure 6 showed the deposition fluxes of size-resolved PBDEs for adult men.The total deposition fluxes of Σ13PBDE were calculated at 26.8 pg h −1 .Among these compounds, penta −BDE was the major congeners and contributed a mean value of 58 % (range: 31−70 %) to the total deposition fluxes.The percent contribution of Σ13PBDEs to the respiratory tract were 84.4 % (22.64 ng h −1 ) in the head airway, 4.6 % (1.24 ng h −1 ) in the tracheobronchial, and 11.0 % (2.93 pg h −1 ) in the alveoli regions, respectively.Moreover, we also found that coarse particles contributed major PBDEs in the head and tracheobronchial regions, while fine particles (accumulation plus Aitken mode particles) contributed lots of PBDEs in the alveoli region.As we known, the particle-size distribution of PBDEs has a decisive influence on their potential health effects.Considering that fine particles can penetrate deeper into the respiratory system compared to coarse particles, fine particle-bound PBDEs are expected to accumulate in the lower parts of the lungs and pose a greater risk to human health.We further evaluated the human health risk that caused by PBDEs by using HQ approach based on inhaled PBDEs data.Figure 7 showed that the HQ values of the individual congeners were ranging from 4.0 × 10 −7 to 6.8 × 10 −5 with the total value of 1.6 × 10 −4 for Σ13PBDE.During the assessment process, we found that the HQ values was highly dependent on the variable daily intake (DI, see Eq. 12), which was incon-12972 Introduction

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Full stant and would result in uncertainty in the risk evaluation.Taking these situations into consideration, we utilized Monte Carlo (MC) simulation to evaluate the influences of uncertainty on this exposure model and to examine whether a difference exists between the model and the calculated HQ.In the simulation, the MC procedure was repeated 5000 times with different calculated HQ.The results of the simulation were depicted in Fig. 8, which exhibited a wide gamma distribution.The 95 % percentile values of HQ were in the range 9.15 × 10 −5 -8.36 × 10 −5 , with the mean value of 8.76 × 10 −5 .In comparison with the corresponding experimental HQ data (1.6×10 −4 ), excellent agreements could be observed between the data, indicating the accuracy of our simulation.By comparison, these HQ values in the present study were much lower than the risk guideline value (1.0) recommended by the US EPA.Even under heavy exercise conditions (assuming a high breathing rate of 3 m 3 h −1 ), the estimated HQ of Σ13PBDE was only (1.17 ± 0.42) × 10 −3 , far less than 1.Thus, particulate PBDEs in Shanghai urban atmosphere posed low noncancer risk through inhalation.However, it was noteworthy here that only particle phase PBDEs participated in the assessment and we were not sure whether the risk posed by atmospheric PBDEs (gas plus particle) exceeded the threshold.Specifically, BDE-47 and BDE-99 mainly existed in gas phase, were probably more toxic and bioaccumulative than other congeners (Darnerud, 2003).Furthermore, we only measured 13 PBDE congeners, neglecting risk caused by substantial octa -BDE and deca -BDE in particles, and other ways of exposure like ingestion, dermal contact were not considered, either.Thus, it was not to suggest that the occurrence of particulate PBDEs in shanghai was not an issue; rather, we advocated further studies that measuring of more PBDE congeners in not only particle phase but gas phase in relation to health risk assessment.

Summary and conclusions
This study revealed the particle-size distribution of PBDEs in the Shanghai urban atmosphere and calculated the contribution of each size-specific particle to PBDEs deposit-Introduction

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Full The non-cancer risk of PBDEs from inhaling particles in Shanghai urban atmosphere showed somewhat low.Further studies will be focused on the investigations of fine or ultrafine particles as transporters of toxic compounds from the atmosphere to the respiratory tract, and the evaluations of exposure risks to ultrafine particles in the atmosphere.Introduction

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Full    Full  Full Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | were purchased from AccuStandard, Inc. (USA).Internal standards ( 13 C-BDE-28, 47, 99 and 153) were purchased from Cambridge Isotope Laboratories (Andover, MA).PBDE congeners are divided into 5 groups (e.g., tri-, tetra-, penta-, hexa-and hepta-BDE) based on the number of bromine atoms in the molecule.The solvents used in this research were HPLC/spectro grade and bought from Tedia Company Inc. (USA)Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | ): log K p-ads = − log p • L + log A TSP − 8.35 (Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Commission on Radiological Protection, 1995) to evaluate the deposition efficiencies and fluxes of inhaled PBDEs in the 12963 Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | with the results of the previous measurements in Shanghai 108−367 pg m −3(Yang et al., 2013) and 104 ± 54 pg m −3(Yu et al., 2011), but was much lower than in Beijing 760 pg m −3(Yang et al., 2013) and in waste recycling zones of Qingyuan close to Guangzhou 3260 pg m −3(Tian et al., 2011).In addition, this results were compared with those reported in aerosol samples from Ontario 88−1250 pg m −3 (Gouin et al., 2012), Chicago 100 ± 35 pg m −3 (Hoh and Hites, 2005) and three different stations in western Europe 0.22−37 pg m −3 (Lee et al., 2004), as well as other Asian cities such as Osaka (9.9−22.3pg m −3 ) (Kakimoto et al., 2014), Busan (5.3−16 pg m −3 ), (Rudichet al., 2007)  and Singapore (7.5 pg m −3 )(Shen et al., 2013).These direct comparison of PBDEs concentrations between various urban environments should be done with caution.Because of the differences existing within any urban environment, PBDE levels could be significantly affected by the location of the sampling site and its proximity to emission sources.Moreover, sampling methodology is a critical parameter affecting the comparison between the observed concentrations of PBDEs in different sites.In most published studies, collection of particulate PBDEs has been performed by using different sampling, pretreatment and instrumental analysis system devices and in some cases underestimation of PBDE concentrations might have occurred because more volatile species were mainly in the gas phase and easy to be lost during membrane sampling or storage period.Seasonal variations were distinct at this urban sites, with significantly higher concentrations measured during winter (104−141 pg m −3 ) and lower concentrations measured Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | . A range of 10-70 µg m −3 was assumed for size-specific particle concentration to calculate K p-measured .The measurement periods ranged from September 2008 to August 2009.The average temperature was 18.4 • C, similar to the temperatures in our study (19 • C).Therefore, their data could serve as useful reference for us to compare Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | et al.
Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper | Discussion Paper |ing in respiratory tract.The particle-size distribution of PBDEs were bimodal distribution with one peak at fine size fractions and the second peak at coarse size fractions.The peaks intensities were associated the molecular weights of PBDEs.The possible reasons were expounded by the parameters such as physicochemical properties of PBDEs, meteorological factors and the emission source.Adsorption and absorption mechanisms respectively played the crucial roles in highly and lightly brominated congeners.Most PBDEs were predominantly distributed in small particles, which contributed the majority of PBDEs deposition fluxes in the human respiratory tract.Total deposition fluxes of target Σ13PBDE in the respiratory tract was calculated at 26.8 ng h −1 .
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Figure 1 .Figure 2 .Figure 3 .
Figure 1.Concentration profiles of PBDE homologue groups in urban atmosphere.Solid vertical lines and the horizontal lines represent median and mean values, respectively.Box plots represent 25th-75th percentiles, whereas whiskers indicate 5th and 95th percentile.The solid spheres crosses represent maximum and minimum values.

Figure 4 .Figure 5 .Figure 6 .Figure 8 .
Figure 4. Pearson correlation between GMD and log p • L of all particulate PBDEs (left) as well as between MFCP and log p • L (right) at 292 K.

Table 3 .
Seasonal concentrations of size-resolved particulate PBDEs in urban Shanghai (pg m −3 ).