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Article

Main Emission Sources and Health Risks of Polycyclic Aromatic Hydrocarbons and Nitro-Polycyclic Aromatic Hydrocarbons at Three Typical Sites in Hanoi

by
Hao Zhang
1,
Chau-Thuy Pham
2,*,
Bin Chen
3,4,5,*,
Xuan Zhang
1,
Yan Wang
1,
Pengchu Bai
1,
Lulu Zhang
6,7,
Seiya Nagao
7,
Akira Toriba
8,
Trung-Dung Nghiem
9 and
Ning Tang
7,10
1
Graduate School of Medical Sciences, Kanazawa University, Kakuma-machi, Kanazawa 920-1192, Japan
2
University of Engineering and Technology, Vietnam National University, Hanoi, 144 Xuan Thuy, Cau Giay, Hanoi 131001, Vietnam
3
Collaborative Innovation Center on Forecast and Evaluation of Meteorological Disasters, Nanjing University of Information Science and Technology, Nanjing 210044, China
4
Key Laboratory of Cloud-Precipitation Physics and Severe Storms, Institute of Atmospheric Physics, Chinese Academy of Sciences, Beijing 100029, China
5
Institute of Carbon Neutrality, Qilu Zhongke, Jinan 250100, China
6
School of Civil Engineering, Architecture and Environment, Hubei University of Technology, Wuhan 430068, China
7
Institute of Nature and Environmental Technology, Kanazawa University, Kakuma-machi, Kanazawa 920-1192, Japan
8
Graduate School of Biomedical Sciences, Nagasaki University, 1-14 Bunkyomachi, Nagasaki 852-8521, Japan
9
School of Environmental Science and Technology, Hanoi University of Science and Technology, No 1 Dai Co Viet, Hanoi 112400, Vietnam
10
Institute of Medical, Pharmaceutical and Health Sciences, Kanazawa University, Kakuma-machi, Kanazawa 920-1192, Japan
*
Authors to whom correspondence should be addressed.
Atmosphere 2023, 14(5), 782; https://doi.org/10.3390/atmos14050782
Submission received: 10 April 2023 / Revised: 22 April 2023 / Accepted: 24 April 2023 / Published: 26 April 2023
(This article belongs to the Special Issue Effects of Natural and Anthropogenic Factors on Climate and Environment (2nd Edition))
Browse Figures
Versions Notes

Abstract

:
Particulate matter-bound polycyclic aromatic hydrocarbons (PAHs) and nitro-PAHs (NPAHs) were first systematically studied in downtown (XT), suburban (GL) and rural (DA) sites in winter and summer in Hanoi, Vietnam, from 2019 to 2022. The mean concentrations of PAHs and NPAHs ranged from 0.76 ng m−3 to 50.2 ng m−3 and 6.07 pg m−3 to 1.95 ng m−3, respectively. The concentrations of PAHs and NPAHs in winter were higher than in summer, except for NPAHs in XT. We found the benzo[a]pyrene (BaP)/benzo[ghi]perylene (BgPe) ratio could effectively identify biomass burning in this study, in which a higher [BaP]/[BgPe] value indicates a greater effect of biomass burning on PAHs and NPAHs. The results indicated that atmospheric PAHs and NPAHs were mainly affected by motor vehicles (especially the unique motorcycles in Southeast Asia) in the summer in Hanoi. In winter, all sites were affected by the burning of rice straw to varying degrees, especially DA. The incremental lifetime cancer risk (ILCR) in Hanoi was first determined through ingestion, inhalation and dermal absorption. The results showed that residents in Hanoi faced high health risks, while females experienced higher health risks than males. The ingestion and dermal pathways indicated higher exposure risks than the usually considered inhalation pathway.

1. Introduction

Among all air pollutants, polycyclic aromatic hydrocarbons (PAHs) and nitro-PAHs (NPAHs) are widely noted for their carcinogenicity and mutagenicity, and are associated with oxidative stress and DNA damage [1,2,3,4]. PAHs are formed through the incomplete combustion of organic materials. For example, the incomplete combustion of petroleum, biomass and coal for industry, vehicle (cars and motorcycles) emissions, residential heating and cooking in daily life are the main anthropogenic sources of atmospheric PAHs [5,6]. PAHs can also be formed through natural sources, for example, forest fires, volcanic activity and living vegetation [7]. Most NPAHs, for example, 6-nitrobenzo[a]pyrene and 1-nitropyrene, are formed during the incomplete combustion of fossil fuel and atmospheric reactions, but some NPAHs (e.g., 2-nitropyrene) can only be formed through atmospheric reactions [8,9].
The inhalation of urban air and tobacco smoke, ingestion of cooked food and drinking water, and dermal absorption of the particles that adhere to exposed skin are considered the main exposure pathways in humans [10]. The inhalation exposure pathway is utilized for those PAHs and NPAHs inhaled by humans through breathing the exhaust gases emitted by the combustion of fossil fuel/biomass [11]. The ingestion exposure pathway is used for the adsorption and formation of those PAHs and NPAHs created during the production, transportation and storage of the foods [12,13]. Different from the inhalation and ingestion pathways, dermal exposure is the main pathway when skin is exposed directly to the consumer products that have high levels of PAHs and NPAHs. Skin exposure can be at both the internal boundary and external boundary during the health risk assessment [14]. Previous researches have shown the close relationship between cancer, heritable (paternal germ-line) mutations and human reproduction and exposure to PAHs [15,16]. The metabolites of certain PAHs and NPAHs, such as benzo[a]pyrene and 6-nitrochrysene, which have been reported by the International Agency for Research on Cancer as Group 1 (carcinogenic to humans) and Group 2A (probably carcinogenic to humans) substances, respectively, might show harmful biological activities to humans [17,18,19]. Moreover, some NPAHs (with direct-acting mutagenicity) that can cause a higher toxicity than their parent PAHs have been increasingly considered [20]. In addition, the study of Shen et al. showed approximately 1372, 709 and 796 cancer cases might occur among one million residents in Shenyang, Chongqing and Changsha, China, due to exposure to PAHs and NPAHs. The study of Yadav et al. showed approximately 102–788 cancer cases might occur among one million residents in Delhi, India, and Thepnuan et al. reported approximately 100 cancer cases might occur among one million residents in Chiang Mai, Thailand [21,22,23]. The lack of detection of PAHs and NPAHs can cause severe underestimations of fine particulate matter (PM2.5)-related health risks since high health-risk PAHs and NPAHs are mainly in particulate matter [24,25]. Thus, the detection of PAHs and NPAHs in PM2.5 is essential, especially in lower- to middle-income countries that are developing rapidly and suffer severe air quality problems [26].
Vietnam is located on the eastern Indochina Peninsula in Southeast Asia and has a population of over 96 million people. In this study, we collected PM2.5 samples at downtown, suburban and rural sites in Hanoi, Vietnam, during the summer and winter from 2019 to 2022. The aim of this research was to (1) determine the characteristics of atmospheric PAHs and NPAHs and the influencing factors of the concentration, composition and emission sources at these different typical sites, and (2) comprehensively evaluate the incremental lifetime cancer risks (ILCRs) of PAHs and NPAHs along three exposure pathways. The results of this study will provide valuable, comprehensive and systematic information for atmospheric conservation not only in Hanoi but also in other regions of Vietnam and Southeast Asian countries with a similar environment.

2. Materials and Methods

2.1. Site Depiction Description

Vietnam has exhibited a relatively high constant growth rate of the gross domestic product (GDP) of approximately 6.5% over the past two decades [27]. The high development speed has aggravated air pollution, with severe health risks to the people of Vietnam [28,29]. Biomass combustion and emissions from motor vehicles, especially motorcycles, are considered to be typical sources of atmospheric PAHs and NPAHs in Vietnam [30,31,32]. Vietnam is the 5th largest rice exporter in global rice production, and approximately 4 × 107 tons of rice is produced in Vietnam each year [33,34]. Rice is harvested two times a year in Vietnam; after each harvest season, there are large amounts of crop residues (e.g., rice straw and rice stubble) that must be treated/processed [35,36,37]. To complete field preparation faster during the two planting periods, straw is normally directly burned in the field [38,39]. Moreover, the registered motorcycles in Vietnam account for over 80% of all registered vehicles, and the number of motorcycles reached more than 6 million units in 2018 [40,41].
In Hanoi, the capital of Vietnam, the air pollution and health risks of PAHs and NPAHs caused by the above reasons have caused much concern [42,43,44,45]. However, to the best of our knowledge, there have only been two investigations so far other than our groups, and these studies were limited to PAHs only. Relevant research in Hanoi was first started in 2005, and roadside particulate and gaseous PAHs were studied [42]. Saha et al. observed the impact of biomass combustion, even in the non-rice straw burning season, through the investigation in 2009–2010 [43]. Our research group first reported concentration and composition of NPAHs and toxic equivalent assessment results of atmospheric PAHs and NPAHs in residential and roadside sites of Hanoi, and we found that the burning of rice straw after the harvest could result in the emission of high levels of PAHs and NPAHs via laboratory and field experiments, and calculated the unit emissions of PAH and NPAH emitted by burning rice straw [32,46,47,48]. Moreover, we simply analyzed the inhalation health risks caused by Hanoi traffic and calculated only PAHs [49,50]. However, the above-mentioned past studies on PAHs and NPAHs in Hanoi lacked an overall analysis. For example, different functional sites in Hanoi include downtown, suburban and rural sites, and in different seasons. In addition, previous scholars conducted simple assessments of the health risks but did not consider ingestion and dermal exposure, the results of which are very likely to underestimate the exposure health risks of PAHs and NPAHs in Hanoi. Therefore, three sites, namely, Xuân Thuỷ (XT), Gia Lâm (GL) and Đông Anh (DA), were selected in Hanoi, Vietnam, as shown in Figure 1. Sampling site XT is situated in the inner part of Hanoi city, far away from rice fields, and represents downtown Hanoi. Sampling site GL is located at the junction of downtown and rural sites, and represents the suburban site of Hanoi. Sampling site DA is situated in the residential site of Hanoi, and is close to rice fields, representing rural Hanoi.

2.2. Sampling

The PM2.5 samples were collected on filters (2500QAT-UP, 25 mm φ and 47 mm φ, Pall Life Sciences, Ann Arbor, MI, USA) by a mini sampler with a Sibata pump (MP-Σ500NII) at a flow rate of 3 L minute−1. The samples were gathered in the winter and summer from 2019 to 2022 in Hanoi, Vietnam. The sampling periods in GL ranged from 24–30 December 2019, 9–15 January 2020 and 15–21 February 2020 (n = 21, called GL-Winter) and from 9–21 June 2020, 7–19 July 2020, and 5–17 August 2020 (n = 21, called GL-Summer). The sampling periods in XT ranged from 20–27 January 2022 (n = 8, called XT-Winter) and 30 June 2022–22 July 2022 (n = 15, called XT-Summer). The sampling periods in DA ranged from 26 November 2021–10 December 2021 (n = 15, called DA-Winter) and 27 June 2022–24 August 2022 (n = 15, called DA-Summer). All filters were replaced every two days in summer and one day in winter. It should be noted that the DA-Winter samples were collected during the rice straw burning period, and the rest of the samples were obtained during the non-rice straw burning season. All sampled filters were stored at −20 °C before analysis.

2.3. PAHs and NPAHs Analysis

Ten PAHs, including fluoranthene (FR), pyrene (Pyr), benz[a]anthracene (BaA), Chrysene (Chr), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[a]pyrene (BaP), benzo[e]pyrene (BeP), benzo[ghi]perylene (BgPe) and indeno [1,2,3-cd]pyrene (IDP) (Supelco Park, Bellefonte, PA, USA), and six NPAHs, including, 2-nitrofluoranthene (2-NFR), 6-nitrochrysene (6-NC), 7-nitrobenz[a]anthracene (7-NbaA), and 6-nitrobenzo[a]pyrene (6-NbaP) (Chiron AS, Trondheim, Norway), 1-, 2-nitropyrenes (NPs) (Aldrich Chemical, Osaka, Japan) were analyzed in this study. Pyrene-d10 (Pyr-d10) and benzo[a]pyrene-d12 (BaP-d12) were used as the internal standards that supplied by Wako Pure Chemicals (Osaka, Japan). All other analytical reagents used in this study were of analytical grade.
Details of the sample analysis process are provided in our previous research and shown in the Supplementary Materials (Text S1) [51,52]. In summary, the sample filters were cut into small pieces and placed into different flasks. Two internal standards (Pyr-d10 and BaP-d12) were then added. Dichloromethane was added to each flask, and the compounds were then ultrasonically extracted, concentrated, diluted and filtered. High-performance liquid chromatography with fluorescence detection (Shimadzu Inc., Kyoto, Japan) was used to analyse the PAHs and NPAHs.
Blank filters were applied in this study to avoid filter contamination during the sampling and transport processes, and the results of the blank filters showed no contamination in the process. The method of recovery was evaluated through the addition of internal standards. The recovery rates were 87% ± 7% for Pyr-d10 and 88% ± 6% for BaP-d12. In addition, blank filters were adopted in three sites to avoid filter contamination during the sampling and transport processes. The same procedures with other sample filters were used to treat the blank filters, including weighing, storage and analysis. The analysis of the blank filters showed no detection of target PAHs and NPAHs, indicating that no contamination occurred during the transportation and storage of the sample filters.

2.4. Health Risk Assessment

Previous reports showed that toxicity equivalency (TEQ) and incremental lifetime cancer risk (ILCR) effectively determine the carcinogenic risk of PAHs and NPAHs. The equations of TEQ are expressed as follows:
TEQi = Ci × TEFi
TEQtotal = ∑TEQi
In Equation (1), the Ci and TEFi are the concentration and reference toxicity equivalence factors, respectively, of each PAH and NPAH (Table S1).
The U.S. EPA standard model was used to calculate the ILCR under different exposure pathways, the equations of which are shown below [53,54]:
ILCRing = (C × CSFing × ∛(BW/70) × IRing × EF × ED)/(BW × AT × 106)
ILCRinh = (C × CSFinh × ∛(BW/70) × IRinh × EF × ED)/(BW × AT × PEF)
ILCRdem = (C × CSFdem × ∛(BW/70) × SA × AF × ABS × EF × ED)/(BW × AT × 106)
In Equations (3)–(5), ILCRing, ILCRinh and ILCRdem are the risk values for ingestion, inhalation and dermal exposure, respectively, C is the concentration of TEQPAH (ng m3), and CSFing, CSFinh and CSFdem are the carcinogenic slope factors for ingestion, inhalation and dermal exposure ((mg kg−1 day−1)−1), respectively [11,55,56]. The values of CSFing, CSFinh and CSFdem capture the cancer-causing ability of BaP at 7.3, 25 and 3.85 ((mg kg−1 day−1)−1), respectively. BW is the average body weight (kg), EF is the annual exposure frequency (day), ED is the duration of exposure (years), SA is the exposed area of skin (cm2), AF is the skin adherence factor (mg cm−2), ABS is the skin absorption factor (day−1), and IRing and IRinh are the intake rates under the ingestion and inhalation exposure routes (mg day−1), respectively [12,57]. The specific values are shown in Table S2.
The sum of all three ILCRs under the different exposure pathways is the total ILCR, which can be calculated as follows:
Total ILCR = ILCRing + ILCRinh + ILCRdem

3. Results

3.1. Concentrations of PAHs and NPAHs

Figure 2 provides the average concentrations of PAHs and NPAHs at the three sites in Hanoi during the sampling periods, the details of which are listed in Tables S3 and S4. The average total concentrations of PAHs and NPAHs at all sites were higher in the winter than in the summer, except for XT-Summer. The results showed that the concentration tendency of PAHs and NPAHs in Hanoi was similar to that in other Asian countries (e.g., China and Japan), in which the main reasons are the differences in sources and meteorological conditions between winter and summer [3,52,58]. However, unlike other Asian countries, the concentrations of PAHs and NPAHs in the downtown site, XT (2.95 ng m−3 and 58.0 pg m−3), were much lower than GL (suburban site, 12.1 ng m−3 and 494 pg m−3, p < 0.01) and DA (rural site, 22.1 ng m−3 and 548 pg m−3, p < 0.01) in the winter. A reduction in traffic and factory manufacturing during the COVID-19 pandemic might be one reason for the relatively low concentration at XT, although the sampling periods were not during the lockdown period. In the summer, the concentrations of PAHs and NPAHs in XT (2.55 ng m−3 and 115 pg m−3) were higher than those in GL (2.20 ng m−3, p > 0.05 and 55.3 pg m−3, p < 0.01) but lower than those in DA (4.38 ng m−3, p < 0.01 and 137 pg m−3, p < 0.05). In addition, the winter–summer ratios of the atmospheric concentrations of PAHs and NPAHs in XT were 1.16 and 0.48, respectively. As shown in Table S3, the increase in PAHs in winter is mainly due to the increase in the particle phase distribution of 4-ring PAHs under the lower temperature condition [24,59]. However, the concentration of NPAHs in the summer was higher than that in the winter, which is due to the increase in the concentration of the secondary formation of NPAHs caused by the strong atmospheric reactivity in the summer and the lack of detection of several NPAH species in the winter (Table S4). This is an important result, and we will identify the phenomenon in our future studies. Nevertheless, compared with XT, the winter–summer ratio of the atmospheric concentrations of PAHs and NPAHs in DA (5.50, 8.93) and GL (5.05, 4.00) were much higher. This result cannot simply be explained by the above reasons. Since the winter sampling of DA was performed in the rice straw combustion period, and the sporadic combustion of RS used as fuel in daily life in the GL site can still emit high concentrations of PAHs and NPAHs [32,47,48], therefore, burning RS in winter may be a reason for the increased concentrations of atmospheric PAHs and NPAHs in DA and GL.

3.2. Prospective Emission Sources of PAHs and NPAHs

The emission sources of PAHs and primary NPAHs can notably control the corresponding composition characteristics [60]. Therefore, to identify the potential sources, diagnostic ratios of PAHs, and primarily NPAHs, such as the ratios of [FR]/([FR] + [Pyr]), [BaA]/([BaA] + [Chr]), [BbF]/([BbF] + [BkF]) and [IDP]/([BgPe] + [IDP]), have typically been used in many studies [6,61,62,63].
Figure 3 showed the values of the PAHs diagnostic ratios at all sampling sites and periods. As shown in Figure 3a, the value of [FR]/([FR] + [Pyr]) in Hanoi was mainly higher than 0.5, which indicates emissions resulting from biomass/coal during winter. The [FR]/([FR] + [Pyr]) value in Hanoi was mainly lower than 0.5, which indicates the notable effect of traffic emissions in summer [64,65,66]. The [BaA]/([BaA] + [Chr]) results showed that the emissions of PAHs and NPAHs resulting from biomass/coal burning cannot be ignored in both summer and winter. The [BbF]/([BbF] + [BkF]) and [IDP]/([IDP] + [BgPe]) results showed that the levels in both summer and winter in Hanoi were affected by traffic emissions (Figure 3b) [6,67,68]. Based on the above results, the atmospheric PAHs and NPAHs of the three different functional sites in Hanoi were comprehensively impacted by biomass combustion and vehicles simultaneously all year. Even in the winter of our selected DA (RS burning period), the proportion of biomass combustion was not fully highlighted. This indicates that these commonly used diagnostic ratios give uncertain results when evaluating the contribution rates of atmospheric PAHs and NPAHs in Hanoi.
Figure 4 shows the ratio of BaP and BgPe in the atmosphere at the three different functional sites in Hanoi in this study and different emission sources in the previous studies. According to ours and other previous studies on the different emission sources of atmospheric PAHs and NPAHs, the [BaP]/[BgPe] ratios are between 0.08 and 0.49 in motorcycle emissions [42,69,70,71,72,73], 0.6 and 1.64 in biomass burning [74,75,76,77,78,79] and 0.53 and 0.90 in automobile emissions [6,80,81,82,83,84,85]. In this study, the ranges of [BaP]/[BgPe] in summer were 0.34–0.47 (GL), 0.23–0.34 (XT) and 0.26–0.42 (DA), which are close to the emission characteristics of motorcycles. In winter, the ranges of [BaP]/[BgPe] were 0.24–0.73 (GL), 0.33–0.51 (XT) and 0.51–1.05 (DA), which are close to the other referenced cities with vehicles as the primary source but lower than the ratio of biomass burning. Our previous studies showed that atmospheric PAHs and NPAHs at the urban stations in Hanoi mainly came from motorcycles, and the contribution rate of vehicles was relatively low [31,46,49]. In addition, as mentioned in the previous section, the DA site is very close to the rice fields, and the samples of DA-Winter were collected during the RS burning period in Hanoi. Therefore, due to the [BaP]/[BgPe] value for motorcycles (median ratio = 0.37) being significantly lower (p < 0.01) than DA-winter (median ratio = 0.81) and higher than the [BaP]/[BgPe] value for biomass burning (median ratio = 0.70), the high [BaP]/[BgPe] ratio of DA-Winter shows that atmospheric PAHs and NPAHs were mainly affected by RS burning. The [BaP]/[BgPe] median values in GL and XT were 0.55 and 0.41, respectively. The [BaP]/[BgPe] median value in GL was significantly higher (p < 0.05) than those for motorcycle exhausts (median ratio = 0.37), and the [BaP]/[BgPe] median value in XT was higher than motorcycle exhausts. This indicated that the other two sites in Hanoi were also impacted by RS burning to varying degrees, although the sampling period was not the main period of RS burning. This may be related to the habit of burning RS as fuel for daily life in some sites of Hanoi in winter [32,47,48,86].
Figure 5 shows the diagnostic ratio distributions of NPAHs and the corresponding sources. 1-NP is mainly formed during the combustion process, whereas 2-NFR and 2-NP are formed through atmospheric reactions [87]. [2-NFR]/[2-NP] ratios close to 10 and 100 are typically used to determine if the formation pathway of NPAHs is dominated by OH and NO3 radical-initiated reactions, respectively [88,89]. As shown in Figure 5a, the [2-NFR]/[2-NP] ratio in this study ranged from 1.89 to 20.8, and closer to 10, the results of which are similar to those in our previous studies in Hanoi, Singapore and Xinxiang, China [32,58,63]. This result showed the importance of the OH radical-initiated reaction pathway in Hanoi in both summer and winter. [2-NFR]/[1-NP] ratios indicate atmospheric NPAHs mainly originate from primary emissions (≤5) or secondary formation (>5) [90]. Although this ratio reflects the aging of aerosols, it is also affected by the atmospheric reaction rate of 1-NP degradation and 2-NFR formation [91]. As shown in Figure 5b, most of the ratios of [2-NFR]/[1-NP] in DA-Winter were close to 5 and were significantly lower (p < 0.01) than in other sites and seasons. Higher 1-NP levels and a lower formation rate of 2-NFR during the rice straw burning period were considered (Table S4). Similar [2-NFR]/[1-NP] results were found in other studies in Lanzhou (mean = 4.70) that experienced a biomass burning period [92].

3.3. Health Risk Assessment

3.3.1. BaP Equivalent Concentration

Among all PAHs, BaP has been found to be the primary contributor to cancer risk, and the BaP equivalent concentration (TEQ) has been widely used as an indicator for modelling air quality and assessing the health risk [93,94]. Table S5 shows the TEQ of each PAH and NPAH. During winter, the average total TEQs in GL, XT and DA were 1730 ± 1577 pg m−3, 363 ± 319 pg m−3 and 3877 ± 2548 pg m−3, respectively, which were higher than those in summer, with corresponding values of 298 ± 218 pg m−3, 335 ± 152 pg m−3 and 571 ± 111 pg m−3, respectively. The highest total TEQ appeared in DA, followed by GL, emphasizing the health effect of biomass burning on the rural residents of Hanoi. In both winter and summer, BaA, Chr, BbF and BaP were the major contributors to the total TEQ, accounting for 58.0% to 85.2% of the total TEQ. 6-NC showed the highest contribution among all three NPAHs due to its high atmospheric concentration and high toxic equivalent factor (10 times higher than that of BaP) [95,96]. Compared with other motorcycle-dominated cities in Southeast Asia, the TEQs in winter in GL and DA were higher than the background site in winter in Ho Chi Minh City, Vietnam (0.91 ng m−3), and the Mae Sot District, Thailand (1.57 ng m−3), the samples of which were collected during a biomass burning period [73,97]. The results show the relatively high health risks in Hanoi, Vietnam, compared with other motorcycle-dominated cities.

3.3.2. ILCR Assessment

The ingestion, inhalation and dermal contact pathways have been reported to be the main pathways for human exposure to PAHs and NPAHs [98]. Based on the total TEQ, the mean ILCR values for the three exposure routes in Hanoi were assessed, as shown in Table 1. The mean ILCR values for males and females were all higher than the acceptable threshold of the ILCR recommended by the U.S. EPA, at 1 × 10−6 [99]. The mean ILCR in DA (3.00 × 10−5 and 3.26 × 10−5 for males and females, respectively) was the highest, followed by GL (1.37 × 10−5 and 1.49 × 10−5 for males and females, respectively), which indicates that approximately 30 and 33 cancer cases might occur among one million males and females, respectively, in rural Hanoi. The results showed relatively high health risks for the residents of Hanoi, especially in the rural site. The risks under the three exposure routes were ranked in the following order: ingestion > dermal > inhalation. In this study, females in Hanoi experienced slightly higher health risks than males, which is different from other studies [51,63]. The reason might be due to the relatively long lifetime of females (80 years) relative to males (75 years) in Vietnam, causing a more extended exposure period. The mean ILCR values in GL and DA in this study were higher than those in our previous study (the mean ILCR values ranged from 7.5 × 10−6 to 2.3 × 10−5), and the results were calculated by using the TEQ [50]. The comparison revealed that the lack of determination of the ILCR through ingestion and dermal ingestion could cause underestimations of the health risks in Hanoi. However, the mean ILCR value in this study was lower than the ILCR value in our previous study obtained by using the CALUX bioassay (the mean ILCR value ranged from 1.0 × 10−4 to 2.8 × 10−4) to determine the TEQ, which is based on the ligand-activated aryl hydrocarbon receptor (AhR)-mediated induction of biological responses to PAHs and can more particularly identify the AhR-mediated risk of PAHs [50,100]. Therefore, future studies in Hanoi should consider all three exposure routes for a more precise determination of ILCR and CALUX-based TEQ values.

4. Conclusions

This study surveyed the concentration, composition and health risk of PM-bound PAHs and NPAHs from 2019 to 2022 at three specific sites in Hanoi, Vietnam. The concentrations of PAHs and NPAHs were highest in a rural site (DA) in winter and were lowest in a suburban site (GL) in summer. Combining several diagnostic ratios of PAHs and NPAHs, we found that the atmospheric PAHs and NPAHs of all areas in Hanoi were affected by the burning of RS with varying degrees in winter, especially during the large-scale RS burning period, while in the summer, the motorcycle was considered the main contributor. The [2-NFR]/[2-NP] ratios showed that OH-initiated reactions comprised the main formation pathway of 2-NFR and 2-NP in Hanoi. The ILCR values in this study were higher than the safe limit, especially in the RS burning period. We first reported that the three exposure routes for ILCR determination showed that ingestion yielded higher health risks than the dermal and inhalation pathways in Hanoi. The implementation of this study will urge the Vietnamese government to take further measures to control RS combustion.

Supplementary Materials

The following supporting information can be downloaded at: https://www.mdpi.com/article/10.3390/atmos14050782/s1, Text S1. Pretreatment and instrumental analysis of PAHs and NPAHs. Table S1: The toxic equivalent factor (TEF) of PAHs and NPAHs; Table S2: Parameters of incremental lifetime cancer risks (ILCRs); Table S3: Average PAHs’ concentration (ng m−3) at three sites in Hanoi during the sampling periods; Table S4: Average NPAHs’ concentration (pg m−3) at DA and XT in Hanoi during the sampling periods; Table S5: The toxic equivalent concentration range (pg/m3) of PAHs and NPAHs (except 2-NP, 7-NbaA and 6-NbaP). References [57,101,102,103,104,105,106] are cited in the supplementary materials.

Author Contributions

Conceptualization, H.Z.; methodology, H.Z., C.-T.P., B.C., L.Z. and N.T.; validation, X.Z., Y.W., P.B., S.N., A.T. and T.-D.N.; investigation, H.Z., X.Z., Y.W., P.B., S.N., A.T. and T.-D.N.; data curation, H.Z.; resources, C.-T.P., B.C., X.Z., Y.W., P.B., L.Z., S.N., A.T., T.-D.N. and N.T.; writing—original draft preparation, H.Z.; writing—review and editing, C.-T.P., B.C., L.Z. and N.T.; supervision, C.-T.P., B.C., L.Z. and N.T. All authors have read and agreed to the published version of the manuscript.

Funding

This research was funded by Sasakawa Scientific Research Grant (2023-6028) from The Japan Science Society; the JST SPRING (JPMJSP2135); the Bilateral Open Partnership Joint Research Projects of the Japan Society for the Promotion of Science, Japan (JPJSBP120219914); the CHOZEN Project of Kanazawa University, Japan (2019); and the cooperative research programs of the Institute of Nature and Environmental Technology, Kanazawa University, Japan (21020, 22005).

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Data are contained within this article or the Supplementary Materials.

Conflicts of Interest

The authors declare no conflict of interest.

References

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Figure 1. Locations of the three sampling sites in Hanoi, Vietnam (Used by permission. Google maps: https://www.google.com/maps) (accessed on 3 March 2023).
Figure 1. Locations of the three sampling sites in Hanoi, Vietnam (Used by permission. Google maps: https://www.google.com/maps) (accessed on 3 March 2023).
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Figure 2. Average concentrations of PAHs (ng m−3) and NPAHs (pg m−3) at three sampling sites in Hanoi.
Figure 2. Average concentrations of PAHs (ng m−3) and NPAHs (pg m−3) at three sampling sites in Hanoi.
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Figure 3. Distribution of PAHs’ isomer ratios among different emission sources at three sampling sites in Hanoi: (a) [FR]/([FR] + [Pyr]) and [BaA]/([BaA] + [Chr]) [64,65,66]; (b) [BbF]/([BbF] + [BkF]) and [IDP]/([BgPe] + [IDP]) [67,68].
Figure 3. Distribution of PAHs’ isomer ratios among different emission sources at three sampling sites in Hanoi: (a) [FR]/([FR] + [Pyr]) and [BaA]/([BaA] + [Chr]) [64,65,66]; (b) [BbF]/([BbF] + [BkF]) and [IDP]/([BgPe] + [IDP]) [67,68].
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Figure 4. The [BaP]/[BgPe] ratios of three sampling sites in Hanoi in this study with biomass burning [74,75,76,77,78,79], motorcycle exhausts [42,69,70,71,72,73] and vehicle exhausts [6,80,81,82,83,84,85].
Figure 4. The [BaP]/[BgPe] ratios of three sampling sites in Hanoi in this study with biomass burning [74,75,76,77,78,79], motorcycle exhausts [42,69,70,71,72,73] and vehicle exhausts [6,80,81,82,83,84,85].
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Figure 5. Distribution of NPAHs’ isomer ratios among different emission sources at three sampling sites in Hanoi: (a) [2-NFR]/[2-NP]; (b) [2-NFR]/[1-NP].
Figure 5. Distribution of NPAHs’ isomer ratios among different emission sources at three sampling sites in Hanoi: (a) [2-NFR]/[2-NP]; (b) [2-NFR]/[1-NP].
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Table
Table 1. The mean ILCRs for three exposure routes from 2019 to 2022 in Hanoi, Vietnam.
Table 1. The mean ILCRs for three exposure routes from 2019 to 2022 in Hanoi, Vietnam.
GLXTDA
MaleFemaleMaleFemaleMaleFemale
Ingestion1.07 × 10−51.17 × 10−53.69 × 10−64.01 × 10−62.36 × 10−52.56 × 10−5
Inhalation5.22 × 10−94.76 × 10−91.79 × 10−91.64 × 10−91.14 × 10−81.04 × 10−9
Dermal2.94 × 10−63.19 × 10−61.01 × 10−61.10 × 10−66.44 × 10−67.00 × 10−6
Total1.37 × 10−51.49 × 10−54.70 × 10−65.11 × 10−63.00 × 10−53.26 × 10−5
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MDPI and ACS Style

Zhang, H.; Pham, C.-T.; Chen, B.; Zhang, X.; Wang, Y.; Bai, P.; Zhang, L.; Nagao, S.; Toriba, A.; Nghiem, T.-D.; et al. Main Emission Sources and Health Risks of Polycyclic Aromatic Hydrocarbons and Nitro-Polycyclic Aromatic Hydrocarbons at Three Typical Sites in Hanoi. Atmosphere 2023, 14, 782. https://doi.org/10.3390/atmos14050782

AMA Style

Zhang H, Pham C-T, Chen B, Zhang X, Wang Y, Bai P, Zhang L, Nagao S, Toriba A, Nghiem T-D, et al. Main Emission Sources and Health Risks of Polycyclic Aromatic Hydrocarbons and Nitro-Polycyclic Aromatic Hydrocarbons at Three Typical Sites in Hanoi. Atmosphere. 2023; 14(5):782. https://doi.org/10.3390/atmos14050782

Chicago/Turabian Style

Zhang, Hao, Chau-Thuy Pham, Bin Chen, Xuan Zhang, Yan Wang, Pengchu Bai, Lulu Zhang, Seiya Nagao, Akira Toriba, Trung-Dung Nghiem, and et al. 2023. "Main Emission Sources and Health Risks of Polycyclic Aromatic Hydrocarbons and Nitro-Polycyclic Aromatic Hydrocarbons at Three Typical Sites in Hanoi" Atmosphere 14, no. 5: 782. https://doi.org/10.3390/atmos14050782

APA Style

Zhang, H., Pham, C. -T., Chen, B., Zhang, X., Wang, Y., Bai, P., Zhang, L., Nagao, S., Toriba, A., Nghiem, T. -D., & Tang, N. (2023). Main Emission Sources and Health Risks of Polycyclic Aromatic Hydrocarbons and Nitro-Polycyclic Aromatic Hydrocarbons at Three Typical Sites in Hanoi. Atmosphere, 14(5), 782. https://doi.org/10.3390/atmos14050782

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