Environmental Pollution 205 (2015) 70e77
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Risk of human exposure to polycyclic aromatic hydrocarbons: A case study in Beijing, China Yanxin Yu a, Qi Li a, Hui Wang a, Bin Wang b, *, Xilong Wang c, Aiguo Ren b, Shu Tao c a
College of Water Science, Beijing Normal University, Beijing 100875, PR China Institute of Reproductive & Child Health/Ministry of Health Key Laboratory of Reproductive Health, School of Public Health, Peking University, Beijing 100191, PR China c Laboratory for Earth Surface Processes, College of Urban and Environmental Sciences, Peking University, Beijing 100871, PR China b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 March 2015 Received in revised form 15 May 2015 Accepted 17 May 2015 Available online xxx
Polycyclic aromatic hydrocarbons (PAHs) can cause adverse effects on human health. The relative contributions of their two major intake routes (diet and inhalation) to population PAH exposure are still unclear. We modeled the contributions of diet and inhalation to the overall PAH exposure of the population of Beijing in China, and assessed their human incremental lifetime cancer risks (ILCR) using a Mont Carlo simulation approach. The results showed that diet accounted for about 85% of low-molecularweight PAH (L-PAH) exposure, while inhalation accounted for approximately 57% of high-molecularweight PAH (H-PAH) exposure of the Beijing population. Meat and cereals were the main contributors to dietary PAH exposure. Both gaseous- and particulate-phase PAHs contributed to L-PAH exposure through inhalation, whereas exposure to H-PAHs was mostly from the particulate-phase. To reduce the cancer incidence of the Beijing population, more attention should be given to inhaled particulate-phase PAHs with considerable carcinogenic potential. © 2015 Elsevier Ltd. All rights reserved.
Keywords: Polycyclic aromatic hydrocarbon Diet Inhalation ILCR Cancer risk
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) emitted from the incomplete combustion of fossil fuels or biomass have attracted widespread public concern because of their adverse effects on human health, including carcinogenicity, teratogenicity, and € m et al., 2002). PAHs are ingested into the mutagenicity (Bostro human body mainly through diet and inhalation (ACGIH, 2005). The relative contributions of the two routes to the total level of PAH exposure in the general population are crucial for PAH exposure assessment, especially in China with its high PAH emissions (Zhang et al., 2007, 2009). Unfortunately, the studies conducted to date are limited and have yielded somewhat inconsistent conclusions. Studies conducted in some East Asian regions have revealed that dietary exposure to PAHs contributes more to the overall exposure level of the local population; e.g., for the sum of the 16 EPA prioritycontrolled PAHs (SPAH16) in Tianjin, China (Li et al., 2005), benzo[a] pyrene equivalent PAH (BAPeq) in Taiyuan, China (Xia et al., 2010, 2013), and pyrene, benzo[b]fluoranthene, and benzo[a]pyrene in
* Corresponding author. E-mail address: [email protected] (B. Wang). http://dx.doi.org/10.1016/j.envpol.2015.05.022 0269-7491/© 2015 Elsevier Ltd. All rights reserved.
Tokyo, Japan (Suzuki and Yoshinaga, 2007). However, in the United States, the primary routes of exposure to low-molecular-weight PAHs (L-PAHs, generally PAHs with 4 benzene rings), including naphthalene, fluorene and pyrene, were inhalation, whereas BAP exposure was predominantly from food intake (Shin et al., 2013). This suggests that the major PAH exposure route varies among populations in different areas and according to PAH molecular weight. Previous studies have focused on total PAHs, BAPeq, or a limited range of PAHs to investigate the relative contributions of various exposure routes to the overall exposure level of the general population. A critical knowledge gap remains with respect to the contribution of the 16 individual U.S. Environmental Protection Agency (EPA) priority-controlled PAHs. The bio-accessibilities of different PAHs greatly depend on their physicochemical properties. For example, fine particulate matter with a relatively high content of high-molecular-weight PAHs (HPAHs, generally PAHs with >4 benzene rings) can penetrate deep into the lungs when inhaled, resulting in greater bio-accessibility than L-PAHs (Ohura et al., 2005). Therefore, it was proposed that PAHs in the particulate phase might pose a greater adverse health effect on the human body than PAHs in the gaseous phase (Li et al., 2005; Zhang et al., 2009). Regarding dietary exposure, the difference in PAH bio-accessibility among food types is becoming a
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
concern. It has been reported that the PAH concentrations vary among food types and according to PAH molecular weight (Xia et al., 2010; Yu et al., 2011). For these reasons, a systematic and detailed survey of the relative contributions of diet and inhalation is necessary. Beijing is the capital of China, with the highest population (about 21 million in 2013) and vehicle volumes (about 5.4 million in 2013) in northern China. Both the annual energy consumption and PAH emission density in this area account for a great proportion of the national total (Zhang et al., 2007). The annual-average concentrations of the sum of the 15 EPA priority-controlled PAHs (SPAH15), excluding naphthalene, in Beijing urban air is reported to be about 206 ng m3, with a BAP concentration of about 6 ng m3, which is considerably higher than the national standard (1 ng/m3) (Li et al. 2014). Yu et al. (2011) reported that the median concentrations of SPAH15 in human milk, placenta, and umbilical cord blood for the Beijing population were 278, 819, and 1372 ng g1 of fat, respectively, which were higher by almost an order of magnitude than corresponding levels in Japan and the United States, which may be caused by the higher PAH concentrations of various food types in the local area (Yu et al., 2011). An epidemiological investigation showed that there were 40,307 new cases of malignant tumors in Beijing in 2012 (Beijing Municipal Government, 2014), which was twice the level of 10 years ago. Among them, lung cancer was ranked first, followed by colorectal, liver, stomach, and prostate cancer for males. The ranking order for females was thyroid carcinoma, followed by breast, lung, colorectal, and uterine cancer. Animal experiments have shown that the position of a tumor is associated with the route of PAH exposure (Hecht, 1999; Latif et al., 2010). In the Beijing population, we found that L-PAHs could more readily penetrate the barrier between placenta and umbilical blood than H-PAHs (Yu et al., 2011). Our recent study showed that PAH concentrations in maternal serum had a strong association with the increased risks of fetal neural tube defects, but no relationships between human serum PAH concentrations and indoor air pollution were found (Wang et al., 2015). To evaluate the overall PAH exposure, the route from food consumption must be taken into consideration. Therefore, the relative contributions of diet and inhalation routes to the overall PAH exposure level of the Beijing population are important, and information is urgently required for environmental scientists, policy makers, and local residents of Beijing. The aims of this study were to investigate: 1) the exposure levels of the Beijing population to 15 individual U.S. EPA prioritycontrolled PAHs through the diet and inhalation routes; 2) the main contributors to dietary and inhaled exposure to PAHs in the population; and 3) the potential cancer risk for the Beijing population caused by PAH exposure. 2. Materials and methods 2.1. Target population and PAHs of concern The target population was residents of Beijing, with ages ranging from 1 to 72 years. The population was divided into two gender groups (male and female). Each group was further divided into four subgroups by age: children (1e6 years old), adolescents (7e18 years old), adults (19e65 years old), and seniors (66e72 years old). The following 15 U. S. EPA priority-controlled PAHs were selected: acenaphthylene (ACY), acenaphthene (ACE), fluorene (FLO), phenanthrene (PHE), anthracene (ANT), fluoranthene (FLA), pyrene (PYR), benzo[a]anthracene (BAA), chrysene (CHR), benzo[b] fluoranthene (BBF), benzo[k]fluoranthene (BKF), benzo[a]pyrene (BAP), indeno[1,2,3-cd]pyrene (IcdP), dibenzo[a,h]anthrancene (DahA), and benzo[g,h,i]perylene (BghiP). Naphthalene was
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excluded because of its higher volatility and poor quantification. Seven of the PAHs were classed as L-PAHs (ACY, ACE, FLO, PHE, ANT, FLA, and PYR) and eight as H-PAHs (BAA, CHR, BBF, BKF, BAP, IcdP, DahA, and BghiP). 2.2. Dietary exposure estimates The residue levels in seven food categoriesdfruits, vegetables, cereals, fish, meat, eggs, and milkdof all selected PAHs were reported by us previously (Yu et al., 2011). To the best of our knowledge, the reported PAH concentrations of the seven food categories in Beijing are the only comprehensively published data. The amounts of the various food categories consumed and body weight of all subgroups were obtained from the Chinese national health and nutrition survey report (Zhai and Yang, 2006), and the details are listed in Table S1 in the Supplementary Information. The data for the concentration of PAHs in foodstuffs were tested to determine if they followed a logarithmic normal distribution, while food consumption and body weight were found to approximately follow a normal distribution. The dietary exposure level (ED, ng person1 day1) of PAHs is the sum of exposures from intake of the seven food categories as follows:
ED ¼
X
Ci FIi
(1)
where Ci and FIi are the PAH concentration (ng g1) and intake rate (g day 1) of food category (i) according to age group in Beijing. The body-weight (BW) adjusted dietary exposure to PAHs (ng kg1 day1) was calculated by dividing ED by body weight (ED/BW). It should be noted that the differences in the bioaccessibility of PAHs in various food categories were neglected in our study because of the limited data available. 2.3. Inhaled exposure estimates The inhaled exposure level (EI, ng person1 day1) was calculated as:
EI ¼ CgPAH þ CpPAH BR
(2)
where CgPAH (ng m3) and CpPAH (ng m3) are the concentrations of gaseous-phase (gPAHs) and particulate-phase (pPAHs) PAHs in Beijing, respectively, as reported in our previous study (Liu et al., 2007a, b) and their detailed description was provided in the Supplementary Information. BR (m3 day1) is the breathing rate of various age groups in Beijing; i.e., 9.3, 15, 16.5, 13 (m3 day1) for children, adolescents, adults, and senior males, respectively, and 8.6, 12, 11, 9.9 (m3 day1) for the corresponding subgroups in the female group, respectively (Wang et al., 2010, 2009). The BR variation was assumed to be 10%. The body weight-adjusted inhaled exposure dose of PAHs (ng kg1 day1) was calculated by dividing EI by body weight (EI/BW). It was assumed that there was no difference in the bioaccessibility of the gaseous- and particulatephase PAHs and the same equation was used to assess their PAH exposure levels, therefore, these results should be interpolated with care. 2.4. Carcinogenic risk assessment The incremental lifetime cancer risk (ILCR) was used to express the carcinogenic risk caused by PAH exposure, as follows:
ILCR ¼ ðE SF ED CF EFÞ=ðBW ATÞ
(3)
where E (ng person1 day1) is the exposure level of BAPeq; SF is the
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Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
Table 1 The estimated median dietary exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Dietary exposure 1.38 104 7.32 102 1.44 104 1.29 102 Dietary exposure 5.47 102 2.92 101 5.74 102 5.21 100
Female Adolescents
Adults
Seniors ¡1
Girls
Adolescents
Adults
Seniors
¡1
without body-weight adjustment (ng person day 1.82 104 1.75 104 1.62 104 1.04 103 1.03 103 9.03 102 1.92 104 1.83 104 1.70 104 1.81 102 1.81 102 1.66 102 with body-weight adjustment (ng kg¡1 day¡1) 3.44 102 2.48 102 2.31 102 1.96 101 1.47 101 1.30 101 3.64 102 2.62 102 2.45 102 3.43 100 2.58 100 2.39 100
) 1.29 6.80 1.37 1.20
104 102 104 102
1.59 8.83 1.67 1.53
104 102 104 102
1.50 8.66 1.58 1.53
104 102 104 102
1.41 7.90 1.47 1.42
104 102 104 102
5.39 2.83 5.68 5.00
102 101 102 100
3.27 1.82 3.45 3.16
102 101 102 100
2.47 1.45 2.60 2.53
102 101 102 100
2.31 1.30 2.41 2.34
102 101 102 100
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
cancer slope factor of BAP, with a geometric mean of 7.27 (mg kg1 d1)1 and a geometric standard deviation of 1.53 for dietary exposure (Xia et al., 2010), and 3.14 (mg kg1 d1)1 and 1.8 for inhalation exposure (Chen and Liao, 2006), respectively; EF (day year1) is the exposure frequency (365), ED (year) is the exposure duration (children, 6; adolescents, 12; adults, 47; seniors, 7), CF (mg ng1) is a conversion factor (i.e. 106), and AT (days) is the lifespan of carcinogens (i.e., 25,550 days). The standard deviations of E and BW and the geometric standard deviation of SF were used in a Monte Carlo simulation. Other parameters were set to be constant. 2.5. Data analysis SPSS 13.0 was used for the statistical analysis, and a significance level of 0.05 was applied. A Monte Carlo simulation (10,000 runs) was conducted to estimate the exposure of the Beijing population to the selected PAHs and to assess their incremental lifetime cancer risk, with consideration of the variations of all the factors included in the calculation. 3. Results and discussion 3.1. Dietary exposure to PAHs Table 1 summarizes the median EDs of the sum of L-PAHs (SLPAHs), sum of H-PAHs (SH-PAHs), sum of all PAHs (SPAH15), and benzo[a]pyrene equivalent PAHs (BAPeq), with and without bodyweight adjustment, among the eight subgroups [i.e., boys (BOY), male adolescents (MAO), male adults (MAU), male seniors (MSE), girls (GIR), female adolescents (FAO), female adults (FAU), and female seniors (FSE)] in the Beijing population. The detailed statistical results for the calculated daily EDs of the 15 individual PAHs, SL-PAHs, SH-PAHs, SPAH15, and BAPeq for the eight subgroups are presented in Table S2-S3. The overall order of EDs before bodyweight adjustment (adolescents > adults > seniors > children) was different from that after body-weight adjustment (children > adolescents > adults > seniors) for any individual PAH (Table S3). For example, the adolescents [1.92 104 (MAO) and 1.67 104 (FAO) ng person1 day1] and children [1.44 104 (MSE) and 1.37 104 (FSE) ng person1 day1] had the highest and lowest median EDs of the SPAH15 among the four age groups in the male and female groups without body-weight adjustment, respectively, while children [5.74 102 (BOY) and 5.68 102 (GIR) ng kg1 day1] were ranked first after body-weight adjustment. The reason for the highest EDs in Beijing being found in adolescents is that this group consumed a larger amount of foodstuffs with higher PAH levels than did the other age groups. However, when bodyweight adjustment was considered, children had the highest EDs for all selected PAHs among the four age groups because they had the lowest body weight. This difference suggests that body fat has
an important role in diluting PAHs in the body. This is consistent with our previous study of dichlorodiphenyltrichloroethanes (DDTs), in which the levels of DDTs in the body appeared to be diluted by body fat to some degree (Tao et al., 2008). Therefore, children should attract more public concern because they experience the highest risk of dietary PAH exposure. With or without body-weight adjustment, males had higher EDs to SPAH15, BAPeq, SL-PAHs, and SH-PAHs than females (Table 1). This is mainly because males consume more foodstuffs than females, which is consistent with previous reports that PAH intake by et al., 2003; Martí-Cid males was higher than that of females (Falco et al., 2008; Xia et al., 2010). Daily EDs of SL-PAHs [percentile range from 10 to 90% (PR): 3.00 104e8.15 104 ng person1 day1 or 1.73 102e6.91 102 ng kg1 day1] were significantly higher than those of SH-PAHs (PR: 3.22 102e2.80 103 ng person1 day1 or 6.85 100e6.18 101 ng kg1 day1) with or without body-weight adjustment for the whole Beijing population. The highest ED of the individual PAHs was found for PHE, with the PR being 2.31 103e5.90 104 ng person1 day1 or 1 3 1 1 4.98 10 e1.32 10 ng kg day , which was about three orders of magnitude higher than the lowest, DahA, which had a PR of 2.18 100e2.12 101 ng person1 day1 or 4.69 102 4.73 101 ng kg1 day1. The order of the EDs for the 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BAA > BBF > BKF > BAP > IcdP > BghiP > DahA. Detailed information for the individual PAHs is presented in Table S3. The relative contributions of the seven food categories to the EDs of SPAH15, BAPeq, SL-PAHs, and SH-PAHs for the whole Beijing population are shown in Fig. 1. A similar pattern in the contribution of the seven food categories to EDs for SPAH15 and SL-PAHs was observed, with the order being meat (34.9e43.4%) > cereals (21.2e26.5%) > fish (11.7e13.9%) > milk (8.3e16.9%) > eggs (6.2e7.8%) > vegetables (2.6e3.8%) > fruits (1.5e1.7%). This was slightly different from the order for BAPeq [cereals (49.0e54.4%) > meat (16.1e21.1%) > fish (10.1e10.4%), vegetables (8.2e10.3%) > milk (3.3e7.3%), eggs (3.8e5.1%) > fruits (0.6e0.7%)], and SH-PAHs [meat (32.7e40.8%) > cereals (20.5e34.7%) > vegetables (8.7e11.4%), fish (8.2e9.1%) > eggs (6.1e8.1%) > milk (2.6e5.4%) > fruits (0.7e0.9%)]. This suggests that cereals and vegetables contribute more to the EDs of H-PAHs than L-PAHs. An important reason for the observed difference is that the residue levels of L-PAHs and H-PAHs vary according to food type. The median concentration of SL-PAHs in the seven food types, from high to low, was meat (34.4 ng g1), fish (34.1 ng g1), eggs (13.8 ng g1), cereals (10.4 ng g1), milk (7.80 ng g1), fruits (6.77 ng g1), and vegetables (3.31 ng g1), while that of SH-PAHs was fish (2.76 ng g1), meat (1.95 ng g1), cereals (0.942 ng g1), eggs (0.648 ng g1), vegetables (0.526 ng g1), milk (0.117 ng g1), and fruits (0.0735 ng g1) (Yu et al., 2011).
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
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Fig. 1. Relative contributions of the seven food categories to the dietary exposure levels of the sum of low-molecular weight PAHs (SL-PAHs) (A), sum of high-molecular weight PAHs (SH-PAHs) (B), sum of 15 PAHs (SPAH15) (C), and benzo[a]pyrene equivalent PAHs (BAPeq) (D) for Beijing population.
The intake dose of BAPeq of the Beijing population was three-to five-fold lower than that of the Taiyuan population and two orders of magnitude lower than that of the Tianjin population, because the concentrations of BAPeq in various food categories from Beijing were significantly lower than those from Taiyuan and Tianjin; this was particularly true for fruits, fish, meat, and milk (Li et al., 2005; Li, 2007; Xia et al., 2010) (Table S4). However, the dietary PAH exposure level of the Beijing population is higher than that reported from developed countries. For example, the dietary intake dose of the 16 EPA priority-controlled PAHs through fish consumption for the general Korean population was reported to be 15.3 ng kg1 d1 (Moon et al., 2010), which is lower than the ED of SPAH15 via fish consumption by residents of Beijing (ranging from 22.2 to 52.8 ng kg1 d1) (Yu et al., 2011). In addition, the Beijing population ingested more BAPeq than did the residents of six cities in Korea (Yoon et al., 2007) and a study group in Catalonia, Spain (Martí-Cid et al., 2008). This is not only because the Beijing population consumed food types with a higher BAPeq concentration but also because they ingest larger amounts of various foodstuffs than the population of Korea (Table S4). 3.2. Inhalation exposure to PAHs Table 2 summarizes the median EIs of SL-PAHs, SH-PAHs, SPAH15, and BAPeq among the eight subgroups of the Beijing population. The detailed statistical results of the daily EIs of PAHs (15 individual PAHs, SL-PAHs, SH-PAHs, SPAH15, and BAPeq, with and without body-weight adjustment, for the eight subgroups are
provided in Table S5eS6. The overall orders of the age groups for EIs without body-weight adjustment for males (adults > adolescents > seniors > boys) were different from females (adolescents > adults > seniors > girls). For the EIs with bodyweight adjustment, the orders for the male and female groups were identical; i.e., (children > adolescents > adults > seniors). This decreasing trend of EIs with an increase in age was consistent with that of dietary exposure (Table 1). Males had higher EIs of SPAH15, BAPeq, SL-PAHs, and SH-PAHs than females with or without bodyweight adjustment, likely because males have a higher respiration rate than females (Wang et al., 2009, 2010). Daily EIs of SL-PAHs (PR: 2.03 103e3.91 103 ng person1 day1 or 3.87 101e8.84 101 ng kg1 day1) were significantly higher than those of SH-PAHs (PR: 8.53 102e1.64 103 ng person1 day1 or 1.60 101e3.68 101 ng kg1 day1) for the whole population. The highest EI of the PAHs was found for PHE, with the PR being 5.32 102e1.02 103 ng person1 day1 or 1.01 101e2.31 101 ng kg1 day1, which was about one order of magnitude higher than the lowest value, for ANT, which had a PR of 4.27 101e8.19 101 ng person1 day1 or 8.16 101 1.88 100 ng kg1 day1. The order of EIs for the 15 individual PAHs was PHE > FLA > PYR > FLO > ACY > BBF > CHR > BAP > BKF > BAA > ACE > BghiP > IcdP > DahA > ANT, which differed from that of the EDs. The EIs of the SL-PAHs were about 2.4fold higher than those of the SH-PAHs, although the differences were smaller than those for the EDs. The detailed information for the individual PAHs is presented in Table S6.
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Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
Table 2 The estimated median inhaled exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Inhaled exposure 2.20 103 9.23 102 3.33 103 4.52 102 Inhaled exposure 8.84 101 3.68 101 1.34 102 1.81 101
Female Adolescents
Adults
Seniors ¡1
Girls
Adolescents
Adults
Seniors
¡1
without body-weight adjustment (ng person day 3.55 103 3.91 103 3.08 103 1.49 103 1.64 103 1.29 103 5.36 103 5.90 103 4.65 103 7.30 102 8.01 102 6.32 102 with body-weight adjustment (ng kg¡1 day¡1) 6.74 101 5.59 101 4.38 101 2.80 101 2.32 101 1.83 101 1.02 102 8.34 101 6.62 101 1.39 101 1.14 101 9.09 100
) 2.03 8.53 3.07 4.18
103 102 103 102
2.84 1.19 4.29 5.83
103 103 103 102
2.61 1.09 3.93 5.36
103 103 103 102
2.34 9.81 3.54 4.81
103 102 103 102
8.44 3.53 1.27 1.73
101 101 102 101
5.88 2.46 8.87 1.21
101 101 101 101
4.30 1.78 6.40 8.86
101 101 101 100
3.87 1.60 5.75 7.93
101 101 101 100
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
The relative contributions of gPAHs and pPAHs to the inhaled exposure levels of the male and female groups were identical. For the whole population, the inhaled dose of gPAHs accounted for 75.5% (SL-PAHs), 0.9% (SH-PAHs), 49.9% (SPAH15), and 0.4% (BAPeq) of the total inhaled exposure, while pPAHs accounted for 24.6% (SLPAHs), 99.1% (SH-PAHs), 50.1% (SPAH15), and 99.6% (BAPeq). This suggests that gPAHs were the major contributor to the inhaled LPAHs but made a negligible contribution to inhaled H-PAHs compared to pPAHs. This is because L-PAHs were the predominant compounds in the vapor phase, while H-PAHs were dominant in the particulate phase (Liu et al., 2007a, b). Out results are higher than the inhaled BAPeq exposure reported in Tianjin, China [i.e., 322 (for children) and 519 (for adults) ng person1 day1 (sampled in 2005) (Bai et al., 2009). Because the atmospheric concentrations of BAP and DahA, which have higher toxicity equivalence factors in Beijing (Liu et al., 2007a, b) were two-to three-fold higher than those in Taiyuan (Xia et al., 2013), the inhaled doses of BAPeq by the Beijing population ranged from twoto four-fold higher compared to the Taiyuan population. The daily inhalation exposure levels of the population in Taiwan-Taichung to BAPeq sampled in 2002e2003 were 252 (children), 1590 (adolescents) and 1628 (adults) ng person1 day1, respectively (Chen and Liao, 2006), which were higher than those in Beijing [i.e., 452 (children), 730 (adolescents), and 801 (adults) ng person1 day1]. This is mainly because the atmospheric BAPeq level (59.4 ± 37.6 ng m3) in Beijing (Liu et al., 2007a, b) was substantially lower than that in Taiwan-Taichung (60.3 mg m3) (Fang et al., 2004a; Fang et al., 2004b; Tsai et al., 2004). Compared to the low levels of inhaled BAPeq (5.44 ng person1 day1) in Japan, which is due to its lower atmospheric BAPeq concentration (0.360 ng m3 in Japan in 2003) (Suzuki and Yoshinaga, 2007), the Beijing population experiences greater inhalation exposure to PAHs.
3.3. Overall exposure to PAHs Table 3 summarizes the median total exposure levels (ETs) of SL-PAHs, SH-PAHs, SPAH15, and BAPeq among the eight subgroups of the Beijing population. The detailed statistical results of the daily ETs (sum of EI and ED) of PAHs (15 individual PAHs, SL-PAHs, SHPAHs, SPAH15, and BAPeq for the eight subgroups are listed in Tables S7-S8. In males without body-weight adjustment, adolescents had a slightly higher ET of SL-PAHs and a lower ET of SHPAHs than adults. However, their ETs of SPAH15 were almost identical, and both were higher than the values of the other two groups. In females without body-weight adjustment, the orders of the four age groups for SL-PAHs, SH-PAHs, SPAH15, and BAPeq were consistent (i.e., adolescents > adults > seniors > girls). After bodyweight adjustment, the ET order for both the male and female groups was consistent with that for the EI and ED; i.e., (children > adolescents > adults > seniors). Overall, male groups had higher ETs than female groups for all individual PAHs (Table S8). The daily ETs of SL-PAHs (PR: 1.57 104e2.31 104 ng person1 day1 or 2.82 102e6.67 102 ng kg1 day1) were significantly higher than those of SH-PAHs (PR: 1.77 103e3.06 103 ng person1 day1 or 3.37 1 1 1 1 10 e7.63 10 ng kg day ), with or without body-weight adjustment, for the whole population. PHE had the highest ETs among all of the individual PAHs, with the PR being 3.15 103e6.00 104 ng person1 day1 or 6.69 101e1.34 103 ng/kg day1, which was two orders of magnitude higher than the lowest ETs for DahA (PR: 3.13 101e1.77 102 ng person1 day1 or 6.48 101 3.88 10 0 ng kg1 day1). The order of the ETs for the 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BBF
Table 3 The estimated median total exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Inhaled exposure 1.67 104 1.91 103 1.86 104 6.31 102 Inhaled exposure 6.67 102 7.63 101 7.45 102 2.53 101
Female Adolescents
Adults
Seniors
without body-weight adjustment (ng person¡1 day¡1) 2.31 104 2.27 104 2.02 104 2.93 103 3.06 103 2.55 103 2.58 104 2.58 104 2.28 104 9.90 102 1.06 103 8.68 102 with body-weight adjustment (ng kg¡1 day¡1) 4.35 102 3.24 102 2.89 102 5.51 101 4.35 101 3.63 101 4.88 102 3.67 102 3.24 102 1.87 101 1.51 101 1.24 101
Girls
Adolescents
Adults
Seniors
1.57 1.77 1.74 5.88
104 103 104 102
1.98 2.39 2.23 7.98
104 103 104 102
1.86 2.26 2.08 7.53
104 103 104 102
1.72 2.06 1.92 6.81
104 103 104 102
6.54 7.32 7.27 2.45
102 101 102 101
4.08 4.92 4.57 1.65
102 101 102 101
3.06 3.74 3.40 1.24
102 101 102 101
2.82 3.37 3.14 1.12
102 101 102 101
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
> BAA > BKF > BAP > BghiP > IcdP > DahA. This was almost identical to the ranking order for the dietary intake of individual PAHs. The relative contributions of EI and ED to the ET of the 15 individual PAHs among the four age groups were similar (Table S9). In the male and female subgroups, the contributions of EI and ED to the ET of the 15 individual PAHs exhibited minor differences (Fig. 2). This suggested that the intake of L-PAHs occurred mostly by dietary exposure, whereas inhalation exposure was more important for HPAHs. For the whole Beijing population, diet accounted for about 84.7% of the SL-PAH intake and 42.6% of the SH-PAH intake, while inhalation accounted for about 15.3 and 57.4%, respectively. These results were similar to the 75.0% contribution of dietary exposure to SPAH16 for the Tianjin population (Li et al., 2005). Different patterns of contribution have been reported in developed countries. More than 90.0% of PYR, BBF, and BAP (as H-PAHs) were ingested in the diet of non-smoking university students in Japan (Suzuki and Yoshinaga, 2007). In the U.S. population, L-PAHs (e.g., NAP, FLO, PHE, and PYR) were ingested mainly by inhalation (more than 97%), whereas BAP (as an important H-PAH) was ingested in the diet (more than 95%) (Shin et al., 2013). It was reported that about 98% of PAH exposure was in the diet route in Spain (Linares et al., 2010). These differences cannot be easily explained by our data, and therefore further in-depth studies are necessary. 3.4. Exposure risk assessment The probability distribution patterns of the ILCR induced by the ETs of BAPeq for the male and female groups of the Beijing
75
population were similar (Fig. 3). The ILCR of the ET for the whole Beijing population ranged from 1.0 107 to 1.0 103, and most of the risk was below the serious or priority risk level (104), as regulated by the U.S. EPA. The percentages of the population with an ICLR due to an ET above the serious acceptable risk level were 0.5% (BOY), 1.3% (MAO), 14.0% (MAU), 0.0% (MSE), 0.3% (GIR), 0.8% (FAO), 10.5% (FAU), and 0.0% (FSE), respectively. Adults, especially males, had the highest cancer risk, whereas other age groups had a low risk (90%. This suggests that this population had a potential cancer risk higher than the acceptable level. The order of the median ILCR for the four age subgroups was adults (3.37 105) > adolescents (1.10 105) > children (7.97 106) > seniors (4.38 106). This order is different from that of the ET (children > adolescents > adults > seniors). This is mainly because adults have the highest exposure duration (47 years), followed by adolescents (12 years). The detailed statistical results for the ILCRs induced by the ED and EI of BAPeq are presented in Table S10. The percentages of the population with an ICLR from both ED and EI above the acceptable risk level ranged from ~80% in seniors to ~100% in adults. However, EI contributed more to the total ILCR than ED because the more highly toxic H-PAHs were ingested by inhalation (Fig. 3). The cancer risk resulting from the ED of PAHs for the Beijing population in this study was lower than that of the Taiyuan population (20.5e39.4% above the serious cancer risk level) in China (Xia et al., 2010), while the two populations had a comparable risk from EI (Xia et al., 2013). Compared to the dietary ILCR (2.30 105) as reported in Korea (Moon et al., 2010) and Egypt (1.31
Fig. 2. Relative contributions of diet and inhalation to the total exposure levels of the 15 individual polycyclic aromatic hydrocarbons (PAHs) for the male (A) and female (B) groups in Beijing population.
Fig. 3. Cumulative probability of incremental lifetime cancer risk (ICLR) of total exposure doses to benzo[a]pyrene equivalent PAHs (BAPequ) for male (left panel) and female (right panel) subgroups in Beijing population.
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Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
105e8.92 105) (Khairy and Lohmann, 2013), the Beijing population has a slightly lower cancer risk in terms of the ILCR (PR: 1.81 106e1.78 105). The cancer risk of inhalation exposure to PAHs for the Beijing population was lower than that reported in Taiwan (Chen and Liao, 2006) and Egypt (Khairy and Lohmann, 2013). However, the ILCR caused by PAHs for the Beijing population was considerably higher than that reported in the Spanish population (109e107) (Linares et al., 2010). This study had several limitations, which should be considered when interpreting the findings. First, both the air and food samples were collected during 2005e2006. The concentrations of PAHs were likely to be different from previous studies because great effort were made by the Beijing government to improve the local air quality for the Olympic Games in 2008. Second, variances in the bioavailability of PAHs among food categories and inhaled air were not considered in the modeling; therefore, the circulating PAH levels were not considered. However, our study also had several strengths. First, the sampling periods of the food and air samples from Beijing were matched, which enabled investigation of the relative contributions of the PAH exposures from the diet and inhalation. Second, selection bias was minimal because seven food categories (i.e., fruits, vegetables, cereals, fish, meat, eggs, and milk) and inhaled gaseous and particulate phases were considered systematically, unlike other studies. Third, the relative contributions of the 15 individual EPA priority-controlled PAHs from the diet and inhalation routes were evaluated. Therefore, our study may present the reference to evaluate the improvement of air pollution control in the following years. 4. Conclusions The order of the overall exposure of the Beijing population to 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BBF > BAA > BKF > BAP > BghiP > IcdP > DahA. With body-weight adjustment, the orders of the PAH exposure levels by the inhalation and dietary routes among the four age groups were identical (i.e., children > adolescents > adults > seniors). Males had higher exposure levels to all individual PAHs than females. Diet mainly accounted for the exposure to L-PAHs (about 84.7% for SLPAHs), while exposure to H-PAHs was mainly from inhalation (about 57.4% for SH-PAHs). Meat and cereals were the main sources of dietary exposure to PAHs. Both gaseous- and particulate-phase PAHs contributed to the inhalation of L-PAHs, whereas inhaled HPAH or BAPeq exposure was almost exclusively due to particulatephase PAHs. The majority of the ILCRs from overall daily exposure to BAPeq for the four subgroups of the Beijing population were below the serious risk level. However, almost 90% were higher than the acceptable level, which suggested that the potential cancer risk of the Beijing population should attract wide public concern. Conflicts of interest The authors declare they have no actual or potential competing financial interests. Role of the funding sources The funding agencies have no role in study design, implementation, data analysis, and interpretation. Acknowledgment This research was supported by grants from the National Natural Science Foundation of China (No. 41371466, 41401583, 41390240).
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Validation of dietary intake of dichlorodiphenyltrichloroethane and metabolites in two populations from Beijing and Shenyang, China based on the residuals in human milk. Environ. Sci. Technol. 42, 7709e7714. Tsai, P.J., Shih, T.S., Chen, H.L., Lee, W.J., Lai, C.H., Liou, S.H., 2004. Assessing and predicting the exposures polycyclic aromatic hydrocarbons (PAHs) and their carcinogenic potencies from vehicle engine exhausts to highway toll station workers. Atmos. Environ. 38, 333e343. Wang, B., Jin, L., Ren, A.G., Yuan, Y., Liu, J.F., Li, Z.W., Zhang, L., Yi, D.Q., Wang, L.L., Zhang, Y.L., Wang, X.L., Tao, S., Finnell, R.H., 2015. Levels of polycyclic aromatic
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Environmental Pollution journal homepage: www.elsevier.com/locate/envpol
Risk of human exposure to polycyclic aromatic hydrocarbons: A case study in Beijing, China Yanxin Yu a, Qi Li a, Hui Wang a, Bin Wang b, *, Xilong Wang c, Aiguo Ren b, Shu Tao c a
College of Water Science, Beijing Normal University, Beijing 100875, PR China Institute of Reproductive & Child Health/Ministry of Health Key Laboratory of Reproductive Health, School of Public Health, Peking University, Beijing 100191, PR China c Laboratory for Earth Surface Processes, College of Urban and Environmental Sciences, Peking University, Beijing 100871, PR China b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 6 March 2015 Received in revised form 15 May 2015 Accepted 17 May 2015 Available online xxx
Polycyclic aromatic hydrocarbons (PAHs) can cause adverse effects on human health. The relative contributions of their two major intake routes (diet and inhalation) to population PAH exposure are still unclear. We modeled the contributions of diet and inhalation to the overall PAH exposure of the population of Beijing in China, and assessed their human incremental lifetime cancer risks (ILCR) using a Mont Carlo simulation approach. The results showed that diet accounted for about 85% of low-molecularweight PAH (L-PAH) exposure, while inhalation accounted for approximately 57% of high-molecularweight PAH (H-PAH) exposure of the Beijing population. Meat and cereals were the main contributors to dietary PAH exposure. Both gaseous- and particulate-phase PAHs contributed to L-PAH exposure through inhalation, whereas exposure to H-PAHs was mostly from the particulate-phase. To reduce the cancer incidence of the Beijing population, more attention should be given to inhaled particulate-phase PAHs with considerable carcinogenic potential. © 2015 Elsevier Ltd. All rights reserved.
Keywords: Polycyclic aromatic hydrocarbon Diet Inhalation ILCR Cancer risk
1. Introduction Polycyclic aromatic hydrocarbons (PAHs) emitted from the incomplete combustion of fossil fuels or biomass have attracted widespread public concern because of their adverse effects on human health, including carcinogenicity, teratogenicity, and € m et al., 2002). PAHs are ingested into the mutagenicity (Bostro human body mainly through diet and inhalation (ACGIH, 2005). The relative contributions of the two routes to the total level of PAH exposure in the general population are crucial for PAH exposure assessment, especially in China with its high PAH emissions (Zhang et al., 2007, 2009). Unfortunately, the studies conducted to date are limited and have yielded somewhat inconsistent conclusions. Studies conducted in some East Asian regions have revealed that dietary exposure to PAHs contributes more to the overall exposure level of the local population; e.g., for the sum of the 16 EPA prioritycontrolled PAHs (SPAH16) in Tianjin, China (Li et al., 2005), benzo[a] pyrene equivalent PAH (BAPeq) in Taiyuan, China (Xia et al., 2010, 2013), and pyrene, benzo[b]fluoranthene, and benzo[a]pyrene in
* Corresponding author. E-mail address: [email protected] (B. Wang). http://dx.doi.org/10.1016/j.envpol.2015.05.022 0269-7491/© 2015 Elsevier Ltd. All rights reserved.
Tokyo, Japan (Suzuki and Yoshinaga, 2007). However, in the United States, the primary routes of exposure to low-molecular-weight PAHs (L-PAHs, generally PAHs with 4 benzene rings), including naphthalene, fluorene and pyrene, were inhalation, whereas BAP exposure was predominantly from food intake (Shin et al., 2013). This suggests that the major PAH exposure route varies among populations in different areas and according to PAH molecular weight. Previous studies have focused on total PAHs, BAPeq, or a limited range of PAHs to investigate the relative contributions of various exposure routes to the overall exposure level of the general population. A critical knowledge gap remains with respect to the contribution of the 16 individual U.S. Environmental Protection Agency (EPA) priority-controlled PAHs. The bio-accessibilities of different PAHs greatly depend on their physicochemical properties. For example, fine particulate matter with a relatively high content of high-molecular-weight PAHs (HPAHs, generally PAHs with >4 benzene rings) can penetrate deep into the lungs when inhaled, resulting in greater bio-accessibility than L-PAHs (Ohura et al., 2005). Therefore, it was proposed that PAHs in the particulate phase might pose a greater adverse health effect on the human body than PAHs in the gaseous phase (Li et al., 2005; Zhang et al., 2009). Regarding dietary exposure, the difference in PAH bio-accessibility among food types is becoming a
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
concern. It has been reported that the PAH concentrations vary among food types and according to PAH molecular weight (Xia et al., 2010; Yu et al., 2011). For these reasons, a systematic and detailed survey of the relative contributions of diet and inhalation is necessary. Beijing is the capital of China, with the highest population (about 21 million in 2013) and vehicle volumes (about 5.4 million in 2013) in northern China. Both the annual energy consumption and PAH emission density in this area account for a great proportion of the national total (Zhang et al., 2007). The annual-average concentrations of the sum of the 15 EPA priority-controlled PAHs (SPAH15), excluding naphthalene, in Beijing urban air is reported to be about 206 ng m3, with a BAP concentration of about 6 ng m3, which is considerably higher than the national standard (1 ng/m3) (Li et al. 2014). Yu et al. (2011) reported that the median concentrations of SPAH15 in human milk, placenta, and umbilical cord blood for the Beijing population were 278, 819, and 1372 ng g1 of fat, respectively, which were higher by almost an order of magnitude than corresponding levels in Japan and the United States, which may be caused by the higher PAH concentrations of various food types in the local area (Yu et al., 2011). An epidemiological investigation showed that there were 40,307 new cases of malignant tumors in Beijing in 2012 (Beijing Municipal Government, 2014), which was twice the level of 10 years ago. Among them, lung cancer was ranked first, followed by colorectal, liver, stomach, and prostate cancer for males. The ranking order for females was thyroid carcinoma, followed by breast, lung, colorectal, and uterine cancer. Animal experiments have shown that the position of a tumor is associated with the route of PAH exposure (Hecht, 1999; Latif et al., 2010). In the Beijing population, we found that L-PAHs could more readily penetrate the barrier between placenta and umbilical blood than H-PAHs (Yu et al., 2011). Our recent study showed that PAH concentrations in maternal serum had a strong association with the increased risks of fetal neural tube defects, but no relationships between human serum PAH concentrations and indoor air pollution were found (Wang et al., 2015). To evaluate the overall PAH exposure, the route from food consumption must be taken into consideration. Therefore, the relative contributions of diet and inhalation routes to the overall PAH exposure level of the Beijing population are important, and information is urgently required for environmental scientists, policy makers, and local residents of Beijing. The aims of this study were to investigate: 1) the exposure levels of the Beijing population to 15 individual U.S. EPA prioritycontrolled PAHs through the diet and inhalation routes; 2) the main contributors to dietary and inhaled exposure to PAHs in the population; and 3) the potential cancer risk for the Beijing population caused by PAH exposure. 2. Materials and methods 2.1. Target population and PAHs of concern The target population was residents of Beijing, with ages ranging from 1 to 72 years. The population was divided into two gender groups (male and female). Each group was further divided into four subgroups by age: children (1e6 years old), adolescents (7e18 years old), adults (19e65 years old), and seniors (66e72 years old). The following 15 U. S. EPA priority-controlled PAHs were selected: acenaphthylene (ACY), acenaphthene (ACE), fluorene (FLO), phenanthrene (PHE), anthracene (ANT), fluoranthene (FLA), pyrene (PYR), benzo[a]anthracene (BAA), chrysene (CHR), benzo[b] fluoranthene (BBF), benzo[k]fluoranthene (BKF), benzo[a]pyrene (BAP), indeno[1,2,3-cd]pyrene (IcdP), dibenzo[a,h]anthrancene (DahA), and benzo[g,h,i]perylene (BghiP). Naphthalene was
71
excluded because of its higher volatility and poor quantification. Seven of the PAHs were classed as L-PAHs (ACY, ACE, FLO, PHE, ANT, FLA, and PYR) and eight as H-PAHs (BAA, CHR, BBF, BKF, BAP, IcdP, DahA, and BghiP). 2.2. Dietary exposure estimates The residue levels in seven food categoriesdfruits, vegetables, cereals, fish, meat, eggs, and milkdof all selected PAHs were reported by us previously (Yu et al., 2011). To the best of our knowledge, the reported PAH concentrations of the seven food categories in Beijing are the only comprehensively published data. The amounts of the various food categories consumed and body weight of all subgroups were obtained from the Chinese national health and nutrition survey report (Zhai and Yang, 2006), and the details are listed in Table S1 in the Supplementary Information. The data for the concentration of PAHs in foodstuffs were tested to determine if they followed a logarithmic normal distribution, while food consumption and body weight were found to approximately follow a normal distribution. The dietary exposure level (ED, ng person1 day1) of PAHs is the sum of exposures from intake of the seven food categories as follows:
ED ¼
X
Ci FIi
(1)
where Ci and FIi are the PAH concentration (ng g1) and intake rate (g day 1) of food category (i) according to age group in Beijing. The body-weight (BW) adjusted dietary exposure to PAHs (ng kg1 day1) was calculated by dividing ED by body weight (ED/BW). It should be noted that the differences in the bioaccessibility of PAHs in various food categories were neglected in our study because of the limited data available. 2.3. Inhaled exposure estimates The inhaled exposure level (EI, ng person1 day1) was calculated as:
EI ¼ CgPAH þ CpPAH BR
(2)
where CgPAH (ng m3) and CpPAH (ng m3) are the concentrations of gaseous-phase (gPAHs) and particulate-phase (pPAHs) PAHs in Beijing, respectively, as reported in our previous study (Liu et al., 2007a, b) and their detailed description was provided in the Supplementary Information. BR (m3 day1) is the breathing rate of various age groups in Beijing; i.e., 9.3, 15, 16.5, 13 (m3 day1) for children, adolescents, adults, and senior males, respectively, and 8.6, 12, 11, 9.9 (m3 day1) for the corresponding subgroups in the female group, respectively (Wang et al., 2010, 2009). The BR variation was assumed to be 10%. The body weight-adjusted inhaled exposure dose of PAHs (ng kg1 day1) was calculated by dividing EI by body weight (EI/BW). It was assumed that there was no difference in the bioaccessibility of the gaseous- and particulatephase PAHs and the same equation was used to assess their PAH exposure levels, therefore, these results should be interpolated with care. 2.4. Carcinogenic risk assessment The incremental lifetime cancer risk (ILCR) was used to express the carcinogenic risk caused by PAH exposure, as follows:
ILCR ¼ ðE SF ED CF EFÞ=ðBW ATÞ
(3)
where E (ng person1 day1) is the exposure level of BAPeq; SF is the
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Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
Table 1 The estimated median dietary exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Dietary exposure 1.38 104 7.32 102 1.44 104 1.29 102 Dietary exposure 5.47 102 2.92 101 5.74 102 5.21 100
Female Adolescents
Adults
Seniors ¡1
Girls
Adolescents
Adults
Seniors
¡1
without body-weight adjustment (ng person day 1.82 104 1.75 104 1.62 104 1.04 103 1.03 103 9.03 102 1.92 104 1.83 104 1.70 104 1.81 102 1.81 102 1.66 102 with body-weight adjustment (ng kg¡1 day¡1) 3.44 102 2.48 102 2.31 102 1.96 101 1.47 101 1.30 101 3.64 102 2.62 102 2.45 102 3.43 100 2.58 100 2.39 100
) 1.29 6.80 1.37 1.20
104 102 104 102
1.59 8.83 1.67 1.53
104 102 104 102
1.50 8.66 1.58 1.53
104 102 104 102
1.41 7.90 1.47 1.42
104 102 104 102
5.39 2.83 5.68 5.00
102 101 102 100
3.27 1.82 3.45 3.16
102 101 102 100
2.47 1.45 2.60 2.53
102 101 102 100
2.31 1.30 2.41 2.34
102 101 102 100
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
cancer slope factor of BAP, with a geometric mean of 7.27 (mg kg1 d1)1 and a geometric standard deviation of 1.53 for dietary exposure (Xia et al., 2010), and 3.14 (mg kg1 d1)1 and 1.8 for inhalation exposure (Chen and Liao, 2006), respectively; EF (day year1) is the exposure frequency (365), ED (year) is the exposure duration (children, 6; adolescents, 12; adults, 47; seniors, 7), CF (mg ng1) is a conversion factor (i.e. 106), and AT (days) is the lifespan of carcinogens (i.e., 25,550 days). The standard deviations of E and BW and the geometric standard deviation of SF were used in a Monte Carlo simulation. Other parameters were set to be constant. 2.5. Data analysis SPSS 13.0 was used for the statistical analysis, and a significance level of 0.05 was applied. A Monte Carlo simulation (10,000 runs) was conducted to estimate the exposure of the Beijing population to the selected PAHs and to assess their incremental lifetime cancer risk, with consideration of the variations of all the factors included in the calculation. 3. Results and discussion 3.1. Dietary exposure to PAHs Table 1 summarizes the median EDs of the sum of L-PAHs (SLPAHs), sum of H-PAHs (SH-PAHs), sum of all PAHs (SPAH15), and benzo[a]pyrene equivalent PAHs (BAPeq), with and without bodyweight adjustment, among the eight subgroups [i.e., boys (BOY), male adolescents (MAO), male adults (MAU), male seniors (MSE), girls (GIR), female adolescents (FAO), female adults (FAU), and female seniors (FSE)] in the Beijing population. The detailed statistical results for the calculated daily EDs of the 15 individual PAHs, SL-PAHs, SH-PAHs, SPAH15, and BAPeq for the eight subgroups are presented in Table S2-S3. The overall order of EDs before bodyweight adjustment (adolescents > adults > seniors > children) was different from that after body-weight adjustment (children > adolescents > adults > seniors) for any individual PAH (Table S3). For example, the adolescents [1.92 104 (MAO) and 1.67 104 (FAO) ng person1 day1] and children [1.44 104 (MSE) and 1.37 104 (FSE) ng person1 day1] had the highest and lowest median EDs of the SPAH15 among the four age groups in the male and female groups without body-weight adjustment, respectively, while children [5.74 102 (BOY) and 5.68 102 (GIR) ng kg1 day1] were ranked first after body-weight adjustment. The reason for the highest EDs in Beijing being found in adolescents is that this group consumed a larger amount of foodstuffs with higher PAH levels than did the other age groups. However, when bodyweight adjustment was considered, children had the highest EDs for all selected PAHs among the four age groups because they had the lowest body weight. This difference suggests that body fat has
an important role in diluting PAHs in the body. This is consistent with our previous study of dichlorodiphenyltrichloroethanes (DDTs), in which the levels of DDTs in the body appeared to be diluted by body fat to some degree (Tao et al., 2008). Therefore, children should attract more public concern because they experience the highest risk of dietary PAH exposure. With or without body-weight adjustment, males had higher EDs to SPAH15, BAPeq, SL-PAHs, and SH-PAHs than females (Table 1). This is mainly because males consume more foodstuffs than females, which is consistent with previous reports that PAH intake by et al., 2003; Martí-Cid males was higher than that of females (Falco et al., 2008; Xia et al., 2010). Daily EDs of SL-PAHs [percentile range from 10 to 90% (PR): 3.00 104e8.15 104 ng person1 day1 or 1.73 102e6.91 102 ng kg1 day1] were significantly higher than those of SH-PAHs (PR: 3.22 102e2.80 103 ng person1 day1 or 6.85 100e6.18 101 ng kg1 day1) with or without body-weight adjustment for the whole Beijing population. The highest ED of the individual PAHs was found for PHE, with the PR being 2.31 103e5.90 104 ng person1 day1 or 1 3 1 1 4.98 10 e1.32 10 ng kg day , which was about three orders of magnitude higher than the lowest, DahA, which had a PR of 2.18 100e2.12 101 ng person1 day1 or 4.69 102 4.73 101 ng kg1 day1. The order of the EDs for the 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BAA > BBF > BKF > BAP > IcdP > BghiP > DahA. Detailed information for the individual PAHs is presented in Table S3. The relative contributions of the seven food categories to the EDs of SPAH15, BAPeq, SL-PAHs, and SH-PAHs for the whole Beijing population are shown in Fig. 1. A similar pattern in the contribution of the seven food categories to EDs for SPAH15 and SL-PAHs was observed, with the order being meat (34.9e43.4%) > cereals (21.2e26.5%) > fish (11.7e13.9%) > milk (8.3e16.9%) > eggs (6.2e7.8%) > vegetables (2.6e3.8%) > fruits (1.5e1.7%). This was slightly different from the order for BAPeq [cereals (49.0e54.4%) > meat (16.1e21.1%) > fish (10.1e10.4%), vegetables (8.2e10.3%) > milk (3.3e7.3%), eggs (3.8e5.1%) > fruits (0.6e0.7%)], and SH-PAHs [meat (32.7e40.8%) > cereals (20.5e34.7%) > vegetables (8.7e11.4%), fish (8.2e9.1%) > eggs (6.1e8.1%) > milk (2.6e5.4%) > fruits (0.7e0.9%)]. This suggests that cereals and vegetables contribute more to the EDs of H-PAHs than L-PAHs. An important reason for the observed difference is that the residue levels of L-PAHs and H-PAHs vary according to food type. The median concentration of SL-PAHs in the seven food types, from high to low, was meat (34.4 ng g1), fish (34.1 ng g1), eggs (13.8 ng g1), cereals (10.4 ng g1), milk (7.80 ng g1), fruits (6.77 ng g1), and vegetables (3.31 ng g1), while that of SH-PAHs was fish (2.76 ng g1), meat (1.95 ng g1), cereals (0.942 ng g1), eggs (0.648 ng g1), vegetables (0.526 ng g1), milk (0.117 ng g1), and fruits (0.0735 ng g1) (Yu et al., 2011).
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
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Fig. 1. Relative contributions of the seven food categories to the dietary exposure levels of the sum of low-molecular weight PAHs (SL-PAHs) (A), sum of high-molecular weight PAHs (SH-PAHs) (B), sum of 15 PAHs (SPAH15) (C), and benzo[a]pyrene equivalent PAHs (BAPeq) (D) for Beijing population.
The intake dose of BAPeq of the Beijing population was three-to five-fold lower than that of the Taiyuan population and two orders of magnitude lower than that of the Tianjin population, because the concentrations of BAPeq in various food categories from Beijing were significantly lower than those from Taiyuan and Tianjin; this was particularly true for fruits, fish, meat, and milk (Li et al., 2005; Li, 2007; Xia et al., 2010) (Table S4). However, the dietary PAH exposure level of the Beijing population is higher than that reported from developed countries. For example, the dietary intake dose of the 16 EPA priority-controlled PAHs through fish consumption for the general Korean population was reported to be 15.3 ng kg1 d1 (Moon et al., 2010), which is lower than the ED of SPAH15 via fish consumption by residents of Beijing (ranging from 22.2 to 52.8 ng kg1 d1) (Yu et al., 2011). In addition, the Beijing population ingested more BAPeq than did the residents of six cities in Korea (Yoon et al., 2007) and a study group in Catalonia, Spain (Martí-Cid et al., 2008). This is not only because the Beijing population consumed food types with a higher BAPeq concentration but also because they ingest larger amounts of various foodstuffs than the population of Korea (Table S4). 3.2. Inhalation exposure to PAHs Table 2 summarizes the median EIs of SL-PAHs, SH-PAHs, SPAH15, and BAPeq among the eight subgroups of the Beijing population. The detailed statistical results of the daily EIs of PAHs (15 individual PAHs, SL-PAHs, SH-PAHs, SPAH15, and BAPeq, with and without body-weight adjustment, for the eight subgroups are
provided in Table S5eS6. The overall orders of the age groups for EIs without body-weight adjustment for males (adults > adolescents > seniors > boys) were different from females (adolescents > adults > seniors > girls). For the EIs with bodyweight adjustment, the orders for the male and female groups were identical; i.e., (children > adolescents > adults > seniors). This decreasing trend of EIs with an increase in age was consistent with that of dietary exposure (Table 1). Males had higher EIs of SPAH15, BAPeq, SL-PAHs, and SH-PAHs than females with or without bodyweight adjustment, likely because males have a higher respiration rate than females (Wang et al., 2009, 2010). Daily EIs of SL-PAHs (PR: 2.03 103e3.91 103 ng person1 day1 or 3.87 101e8.84 101 ng kg1 day1) were significantly higher than those of SH-PAHs (PR: 8.53 102e1.64 103 ng person1 day1 or 1.60 101e3.68 101 ng kg1 day1) for the whole population. The highest EI of the PAHs was found for PHE, with the PR being 5.32 102e1.02 103 ng person1 day1 or 1.01 101e2.31 101 ng kg1 day1, which was about one order of magnitude higher than the lowest value, for ANT, which had a PR of 4.27 101e8.19 101 ng person1 day1 or 8.16 101 1.88 100 ng kg1 day1. The order of EIs for the 15 individual PAHs was PHE > FLA > PYR > FLO > ACY > BBF > CHR > BAP > BKF > BAA > ACE > BghiP > IcdP > DahA > ANT, which differed from that of the EDs. The EIs of the SL-PAHs were about 2.4fold higher than those of the SH-PAHs, although the differences were smaller than those for the EDs. The detailed information for the individual PAHs is presented in Table S6.
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Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
Table 2 The estimated median inhaled exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Inhaled exposure 2.20 103 9.23 102 3.33 103 4.52 102 Inhaled exposure 8.84 101 3.68 101 1.34 102 1.81 101
Female Adolescents
Adults
Seniors ¡1
Girls
Adolescents
Adults
Seniors
¡1
without body-weight adjustment (ng person day 3.55 103 3.91 103 3.08 103 1.49 103 1.64 103 1.29 103 5.36 103 5.90 103 4.65 103 7.30 102 8.01 102 6.32 102 with body-weight adjustment (ng kg¡1 day¡1) 6.74 101 5.59 101 4.38 101 2.80 101 2.32 101 1.83 101 1.02 102 8.34 101 6.62 101 1.39 101 1.14 101 9.09 100
) 2.03 8.53 3.07 4.18
103 102 103 102
2.84 1.19 4.29 5.83
103 103 103 102
2.61 1.09 3.93 5.36
103 103 103 102
2.34 9.81 3.54 4.81
103 102 103 102
8.44 3.53 1.27 1.73
101 101 102 101
5.88 2.46 8.87 1.21
101 101 101 101
4.30 1.78 6.40 8.86
101 101 101 100
3.87 1.60 5.75 7.93
101 101 101 100
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
The relative contributions of gPAHs and pPAHs to the inhaled exposure levels of the male and female groups were identical. For the whole population, the inhaled dose of gPAHs accounted for 75.5% (SL-PAHs), 0.9% (SH-PAHs), 49.9% (SPAH15), and 0.4% (BAPeq) of the total inhaled exposure, while pPAHs accounted for 24.6% (SLPAHs), 99.1% (SH-PAHs), 50.1% (SPAH15), and 99.6% (BAPeq). This suggests that gPAHs were the major contributor to the inhaled LPAHs but made a negligible contribution to inhaled H-PAHs compared to pPAHs. This is because L-PAHs were the predominant compounds in the vapor phase, while H-PAHs were dominant in the particulate phase (Liu et al., 2007a, b). Out results are higher than the inhaled BAPeq exposure reported in Tianjin, China [i.e., 322 (for children) and 519 (for adults) ng person1 day1 (sampled in 2005) (Bai et al., 2009). Because the atmospheric concentrations of BAP and DahA, which have higher toxicity equivalence factors in Beijing (Liu et al., 2007a, b) were two-to three-fold higher than those in Taiyuan (Xia et al., 2013), the inhaled doses of BAPeq by the Beijing population ranged from twoto four-fold higher compared to the Taiyuan population. The daily inhalation exposure levels of the population in Taiwan-Taichung to BAPeq sampled in 2002e2003 were 252 (children), 1590 (adolescents) and 1628 (adults) ng person1 day1, respectively (Chen and Liao, 2006), which were higher than those in Beijing [i.e., 452 (children), 730 (adolescents), and 801 (adults) ng person1 day1]. This is mainly because the atmospheric BAPeq level (59.4 ± 37.6 ng m3) in Beijing (Liu et al., 2007a, b) was substantially lower than that in Taiwan-Taichung (60.3 mg m3) (Fang et al., 2004a; Fang et al., 2004b; Tsai et al., 2004). Compared to the low levels of inhaled BAPeq (5.44 ng person1 day1) in Japan, which is due to its lower atmospheric BAPeq concentration (0.360 ng m3 in Japan in 2003) (Suzuki and Yoshinaga, 2007), the Beijing population experiences greater inhalation exposure to PAHs.
3.3. Overall exposure to PAHs Table 3 summarizes the median total exposure levels (ETs) of SL-PAHs, SH-PAHs, SPAH15, and BAPeq among the eight subgroups of the Beijing population. The detailed statistical results of the daily ETs (sum of EI and ED) of PAHs (15 individual PAHs, SL-PAHs, SHPAHs, SPAH15, and BAPeq for the eight subgroups are listed in Tables S7-S8. In males without body-weight adjustment, adolescents had a slightly higher ET of SL-PAHs and a lower ET of SHPAHs than adults. However, their ETs of SPAH15 were almost identical, and both were higher than the values of the other two groups. In females without body-weight adjustment, the orders of the four age groups for SL-PAHs, SH-PAHs, SPAH15, and BAPeq were consistent (i.e., adolescents > adults > seniors > girls). After bodyweight adjustment, the ET order for both the male and female groups was consistent with that for the EI and ED; i.e., (children > adolescents > adults > seniors). Overall, male groups had higher ETs than female groups for all individual PAHs (Table S8). The daily ETs of SL-PAHs (PR: 1.57 104e2.31 104 ng person1 day1 or 2.82 102e6.67 102 ng kg1 day1) were significantly higher than those of SH-PAHs (PR: 1.77 103e3.06 103 ng person1 day1 or 3.37 1 1 1 1 10 e7.63 10 ng kg day ), with or without body-weight adjustment, for the whole population. PHE had the highest ETs among all of the individual PAHs, with the PR being 3.15 103e6.00 104 ng person1 day1 or 6.69 101e1.34 103 ng/kg day1, which was two orders of magnitude higher than the lowest ETs for DahA (PR: 3.13 101e1.77 102 ng person1 day1 or 6.48 101 3.88 10 0 ng kg1 day1). The order of the ETs for the 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BBF
Table 3 The estimated median total exposure levels of polycyclic aromatic hydrocarbons (PAHs) for the eight subgroups of Beijing population. PAHsa
Male Boys
SL-PAHs SH-PAHs SPAH15 BAPeq SL-PAHs SH-PAHs SPAH15 BAPeq a
Inhaled exposure 1.67 104 1.91 103 1.86 104 6.31 102 Inhaled exposure 6.67 102 7.63 101 7.45 102 2.53 101
Female Adolescents
Adults
Seniors
without body-weight adjustment (ng person¡1 day¡1) 2.31 104 2.27 104 2.02 104 2.93 103 3.06 103 2.55 103 2.58 104 2.58 104 2.28 104 9.90 102 1.06 103 8.68 102 with body-weight adjustment (ng kg¡1 day¡1) 4.35 102 3.24 102 2.89 102 5.51 101 4.35 101 3.63 101 4.88 102 3.67 102 3.24 102 1.87 101 1.51 101 1.24 101
Girls
Adolescents
Adults
Seniors
1.57 1.77 1.74 5.88
104 103 104 102
1.98 2.39 2.23 7.98
104 103 104 102
1.86 2.26 2.08 7.53
104 103 104 102
1.72 2.06 1.92 6.81
104 103 104 102
6.54 7.32 7.27 2.45
102 101 102 101
4.08 4.92 4.57 1.65
102 101 102 101
3.06 3.74 3.40 1.24
102 101 102 101
2.82 3.37 3.14 1.12
102 101 102 101
SL-PAHs: sum of low molecular-weight PAHs; SH-PAHs: sum of high-molecular weight PAHs, SPAH15: sum of all PAHs, and BAPeq: benzo[a]pyrene equivalent PAH.
Y. Yu et al. / Environmental Pollution 205 (2015) 70e77
> BAA > BKF > BAP > BghiP > IcdP > DahA. This was almost identical to the ranking order for the dietary intake of individual PAHs. The relative contributions of EI and ED to the ET of the 15 individual PAHs among the four age groups were similar (Table S9). In the male and female subgroups, the contributions of EI and ED to the ET of the 15 individual PAHs exhibited minor differences (Fig. 2). This suggested that the intake of L-PAHs occurred mostly by dietary exposure, whereas inhalation exposure was more important for HPAHs. For the whole Beijing population, diet accounted for about 84.7% of the SL-PAH intake and 42.6% of the SH-PAH intake, while inhalation accounted for about 15.3 and 57.4%, respectively. These results were similar to the 75.0% contribution of dietary exposure to SPAH16 for the Tianjin population (Li et al., 2005). Different patterns of contribution have been reported in developed countries. More than 90.0% of PYR, BBF, and BAP (as H-PAHs) were ingested in the diet of non-smoking university students in Japan (Suzuki and Yoshinaga, 2007). In the U.S. population, L-PAHs (e.g., NAP, FLO, PHE, and PYR) were ingested mainly by inhalation (more than 97%), whereas BAP (as an important H-PAH) was ingested in the diet (more than 95%) (Shin et al., 2013). It was reported that about 98% of PAH exposure was in the diet route in Spain (Linares et al., 2010). These differences cannot be easily explained by our data, and therefore further in-depth studies are necessary. 3.4. Exposure risk assessment The probability distribution patterns of the ILCR induced by the ETs of BAPeq for the male and female groups of the Beijing
75
population were similar (Fig. 3). The ILCR of the ET for the whole Beijing population ranged from 1.0 107 to 1.0 103, and most of the risk was below the serious or priority risk level (104), as regulated by the U.S. EPA. The percentages of the population with an ICLR due to an ET above the serious acceptable risk level were 0.5% (BOY), 1.3% (MAO), 14.0% (MAU), 0.0% (MSE), 0.3% (GIR), 0.8% (FAO), 10.5% (FAU), and 0.0% (FSE), respectively. Adults, especially males, had the highest cancer risk, whereas other age groups had a low risk (90%. This suggests that this population had a potential cancer risk higher than the acceptable level. The order of the median ILCR for the four age subgroups was adults (3.37 105) > adolescents (1.10 105) > children (7.97 106) > seniors (4.38 106). This order is different from that of the ET (children > adolescents > adults > seniors). This is mainly because adults have the highest exposure duration (47 years), followed by adolescents (12 years). The detailed statistical results for the ILCRs induced by the ED and EI of BAPeq are presented in Table S10. The percentages of the population with an ICLR from both ED and EI above the acceptable risk level ranged from ~80% in seniors to ~100% in adults. However, EI contributed more to the total ILCR than ED because the more highly toxic H-PAHs were ingested by inhalation (Fig. 3). The cancer risk resulting from the ED of PAHs for the Beijing population in this study was lower than that of the Taiyuan population (20.5e39.4% above the serious cancer risk level) in China (Xia et al., 2010), while the two populations had a comparable risk from EI (Xia et al., 2013). Compared to the dietary ILCR (2.30 105) as reported in Korea (Moon et al., 2010) and Egypt (1.31
Fig. 2. Relative contributions of diet and inhalation to the total exposure levels of the 15 individual polycyclic aromatic hydrocarbons (PAHs) for the male (A) and female (B) groups in Beijing population.
Fig. 3. Cumulative probability of incremental lifetime cancer risk (ICLR) of total exposure doses to benzo[a]pyrene equivalent PAHs (BAPequ) for male (left panel) and female (right panel) subgroups in Beijing population.
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105e8.92 105) (Khairy and Lohmann, 2013), the Beijing population has a slightly lower cancer risk in terms of the ILCR (PR: 1.81 106e1.78 105). The cancer risk of inhalation exposure to PAHs for the Beijing population was lower than that reported in Taiwan (Chen and Liao, 2006) and Egypt (Khairy and Lohmann, 2013). However, the ILCR caused by PAHs for the Beijing population was considerably higher than that reported in the Spanish population (109e107) (Linares et al., 2010). This study had several limitations, which should be considered when interpreting the findings. First, both the air and food samples were collected during 2005e2006. The concentrations of PAHs were likely to be different from previous studies because great effort were made by the Beijing government to improve the local air quality for the Olympic Games in 2008. Second, variances in the bioavailability of PAHs among food categories and inhaled air were not considered in the modeling; therefore, the circulating PAH levels were not considered. However, our study also had several strengths. First, the sampling periods of the food and air samples from Beijing were matched, which enabled investigation of the relative contributions of the PAH exposures from the diet and inhalation. Second, selection bias was minimal because seven food categories (i.e., fruits, vegetables, cereals, fish, meat, eggs, and milk) and inhaled gaseous and particulate phases were considered systematically, unlike other studies. Third, the relative contributions of the 15 individual EPA priority-controlled PAHs from the diet and inhalation routes were evaluated. Therefore, our study may present the reference to evaluate the improvement of air pollution control in the following years. 4. Conclusions The order of the overall exposure of the Beijing population to 15 individual PAHs was PHE > FLO > FLA > PYR > ACE > ACY > ANT > CHR > BBF > BAA > BKF > BAP > BghiP > IcdP > DahA. With body-weight adjustment, the orders of the PAH exposure levels by the inhalation and dietary routes among the four age groups were identical (i.e., children > adolescents > adults > seniors). Males had higher exposure levels to all individual PAHs than females. Diet mainly accounted for the exposure to L-PAHs (about 84.7% for SLPAHs), while exposure to H-PAHs was mainly from inhalation (about 57.4% for SH-PAHs). Meat and cereals were the main sources of dietary exposure to PAHs. Both gaseous- and particulate-phase PAHs contributed to the inhalation of L-PAHs, whereas inhaled HPAH or BAPeq exposure was almost exclusively due to particulatephase PAHs. The majority of the ILCRs from overall daily exposure to BAPeq for the four subgroups of the Beijing population were below the serious risk level. However, almost 90% were higher than the acceptable level, which suggested that the potential cancer risk of the Beijing population should attract wide public concern. Conflicts of interest The authors declare they have no actual or potential competing financial interests. Role of the funding sources The funding agencies have no role in study design, implementation, data analysis, and interpretation. Acknowledgment This research was supported by grants from the National Natural Science Foundation of China (No. 41371466, 41401583, 41390240).
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