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Article

Seasonal Variability and Risk Assessment of Atmospheric Polycyclic Aromatic Hydrocarbons and Hydroxylated Polycyclic Aromatic Hydrocarbons in Kanazawa, Japan

by
Masato Honda
1,*,
Kazuichi Hayakawa
2,
Lulu Zhang
2,3,
Ning Tang
2 and
Hiroyuki Nakamura
4
1
Botanical Garden, Institute of Nature and Environmental Technology, Kanazawa University, Kakuma-machi, Kanazawa 920-1192, Japan
2
Institute of Nature and Environmental Technology, Kanazawa University, Kakuma-machi, Kanazawa 920-1192, Japan
3
School of Civil Engineering, Architecture and Environment, Hubei University of Technology, Wuhan 430068, China
4
Department of Hygiene and Public Health, Faculty of Medicine, Institute of Medical, Pharmaceutical and Health Sciences, Kanazawa University, Kanazawa 920-8640, Japan
*
Author to whom correspondence should be addressed.
Appl. Sci. 2022, 12(19), 9469; https://0-doi-org.brum.beds.ac.uk/10.3390/app12199469
Submission received: 31 August 2022 / Revised: 8 September 2022 / Accepted: 15 September 2022 / Published: 21 September 2022
(This article belongs to the Special Issue PM2.5 and PM10: Atmospheric Behaviors and Correlation with Health Effects)
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Abstract

:
Polycyclic aromatic hydrocarbons (PAHs) and hydroxylated PAHs (OH-PAHs) are ubiquitous atmospheric pollutants that are a concern because of their endocrine disrupting activities. In this study, seasonal air sampling was conducted in 2017 and 2018 in Kanazawa, Ishikawa Prefecture, Japan. The concentrations and seasonal variations of PAHs and OH-PAHs were analyzed, and health risks of individual congeners were evaluated based on their relative endocrine activity. The atmospheric concentrations of PAHs and OH-PAHs showed seasonal trends with higher concentrations in the winter (daily average ± standard deviation: 1.00 ± 0.26 ng/m3 for PAHs and 75.06 ± 23.38 pg/m3 for OH-PAHs) and lower concentrations in the summer (0.30 ± 0.09 ng/m3 for PAHs and 17.08 ± 4.83 pg/m3 for OH-PAHs). There were significant positive correlations between the concentrations of atmospheric PAHs and OH-PAHs. Additionally, the health risk from the endocrine disrupting potential of each OH-PAH was evaluated using relative estrogenic and antiestrogenic activities. OH-PAHs with four rings, such as OH-chrysenes and OH-benz[a]anthracene, had particularly high health risks. These results suggest that atmospheric OH-PAHs are a potential health risk for organisms and thus warrant further research.

1. Introduction

Polycyclic aromatic hydrocarbons (PAHs) are a group of chemicals that contain two or more fused benzene rings without any functional group modifications. PAHs are generated through the incomplete combustion of organic matter such as wood, crop residues, and fossil fuels [1,2,3], and are released into the atmosphere through human activity. Therefore, PAHs are ubiquitous environmental pollutants, particularly in the atmosphere [2,3,4,5,6]. PAHs are of concern because of their mutagenicity, carcinogenicity and endocrine disrupting potential [7,8,9,10,11,12]. Because of its carcinogenicity, benzo[a]pyrene (one of major PAHs) is classified as a Group 1 carcinogen (carcinogenic to humans) by the International Agency for Research on Cancer. The United States Environmental Protection Agency has selected 16 representative PAHs, including those with between two and six rings, as priority compounds for environmental monitoring. Similar to other atmospheric combustion-related pollutants, PAHs show a distinct seasonal variability. The atmospheric concentrations of PAHs in the winter are generally higher than in the summer [4,13,14] because of the increased use of fuels for heating purposes.
Interest in the health risks and atmospheric behavior of oxidized derivatives of PAHs is growing rapidly, and many studies have been conducted [15,16]. One group of PAH derivatives, hydroxylated PAHs (OH-PAHs), has been detected in the atmosphere [17,18,19,20,21,22,23,24]. OH-PAHs are oxidized derivatives of PAHs that are generated via combustion of organic matter such as tobacco, wood, and fossil fuel [17,21,25]. The OH-PAHs can also be generated via hydroxyl (OH) radical-initiated reactions in the atmosphere [26,27,28,29]. In a study in China, Lin et al. [22] found that biomass burning is the dominant source (79.6%) of atmospheric OH-PAHs during colder seasons when domestic heating appliances are used, whereas traffic emissions (59.6%) and secondary formation (21.7%) are the dominant sources in the warmer seasons.
OH-PAHs have high health risks, with their endocrine disrupting potential being of particular concern [8,12,20,30,31,32]. In a previous study, Hayakawa et al. [30] evaluated the relative estrogenic and antiestrogenic activities of 14 PAHs and 63 OH-PAHs. In their study, OH-PAHs with four rings, such as hydroxylated benz[a]anthracenes, hydroxylated chrysenes, and hydroxylated benzo[c]phenanthrene, showed strong estrogenic and antiestrogenic activity. Hydroxylated PAHs with four rings, especially hydroxylated chrysene, have a similar chemical structure to 17β-estradiol (estrogen). Consequently, several OH-PAHs can interact with the human estrogen receptor (hER). The hER includes ERα and ERβ, which have similar binding affinities but different distributions in the tissues. It is believed that endocrine disrupting chemicals, such as OH-PAHs, bind to ERs and induce estrogenic and antiestrogenic activity [20,30]. Previous studies have revealed that OH-PAHs have much higher teratogenicity and developmental toxicity in fish compared to their parent PAHs [32,33]. While OH-PAHs are of toxicological concern, little is known about the seasonal variations in atmospheric OH-PAH concentrations [18,21]. In addition, a comprehensive toxicological risk assessment, especially on the estrogenic and antiestrogenic activities of atmospheric OH-PAHs towards humans, has not been conducted.
In this study, seasonal air sampling was conducted in 2017 and 2018 at Kanazawa, Ishikawa Prefecture, Japan. The concentrations and seasonal variations of airborne particle-bound PAHs and OH-PAHs were analyzed. Human health risks of the individual PAHs and OH-PAHs were evaluated using their relative estrogenic and antiestrogenic activities. While human health risk from atmospheric pollution is a global concern, few studies focus on the environmental level and health risk of endocrine disrupting compounds such as OH-PAHs. This is the first study to evaluate the contributions of atmospheric OH-PAHs to endocrine disruption in humans.

2. Materials and Methods

2.1. Chemicals and Regents

A Supelco PAH standard mixture was purchased from Sigma–Aldrich (Waltham, MA, US) and contained the following nine target PAHs: fluoranthene (Flt), pyrene (Pyr), benz[a]anthracene (BaA), chrysene (Chr), benzo[b]fluoranthene (BbF), benzo[k]fluoranthene (BkF), benzo[a]pyrene (BaP), benzo[g,h,i]perylene (BgPe), and indeno[1,2,3-cd]pyrene (IDP). The following OH-PAH standards were purchased from Toronto Research Chemicals (Toronto, Canada): 1- and 2-hydroxylated naphthalenes (1OH- and 2OH-Nap); 1-, 2-, and 3-hydroxylated fluorenes (1OH-, 2OH-, and 3OH-Flu); 1-, 2-, 3-, 4-, 9-hydroxylated phenanthrenes (1OH-, 2OH-, 3OH-, 4OH-, and 9OH-Phe); 1-hydroxylated pyrene (1OH-Pyr); 2-, 3-, and 6-hydroxylated chrysenes (2OH-, 3OH-, and 6OH-Chr); 3-hydroxylated benz[a]anthracene (3OH-BaA); and 3-hydroxylated benzo[a]pyrene (3OH-BaP). The deuterated PAHs; Pyr-d10 and BaP-d12 were purchased from Wako Pure Chemical Industries, Ltd. (Osaka, Japan). The deuterated OH-PAHs; 2OH-Nap-d7, 2OH-Flu-d9, 1OH- and 4OH-Phe-d9, 3OH-BaA-d11, 3OH-Chr-d11, 1OH-Pyr-d9, and 3OH-BaP-d11 were purchased from Toronto Research Chemicals. All organic solvents, chemicals, and ultrapure water used in the experiments were analytical or special grade.

2.2. Air Sampling

Air sampling was conducted next to a main road in a residential area of Kanazawa, Japan, using a high-volume air sampler (HV-RW, Sibata Scientific Technology LTD., Tokyo, Japan) equipped with a quartz fiber filter with a cut-off for particulate matter with sizes less than 2.5 μm (PM2.5) (2500QAT-UP, Pallflex Products corporation, Putnam, CT, US) at a flow rate of 1000 L/min [2,34]. There were no obvious sources of combustion PM2.5 near the monitoring station except for traffic. Filter samples were collected daily (every 24 h) for 1 week in spring (24–30 April), summer (21–27 August), and autumn (6–12 November) in 2017, and winter (19–25 February) and spring (24–30 April) in 2018. The filters were preheated at 600 °C for 4 h before using to minimize background PAH contamination. The amounts of PM2.5 were calculated based on the difference of the filter weight before and after sampling. The filters were kept in a freezer at −30 °C until further analysis.

2.3. PAHs Analysis

The sample preparation and analytical methods for particle-bound PAHs are detailed in a previous study [2]. Briefly, the filters were cut into pieces and placed in separate flasks. An internal standard solution was added to each flask. Benzene/ethanol (3:1, v/v) was added as the extraction solvent and sonicated. The extraction was repeated twice. The extracts were filtered to remove solid materials and then washed with a NaOH solution (5%, w/v) and H2SO4 solution (20%, v/v). This was followed by washing with distilled water. An amount of 100 µL Dimethyl sulfoxide was added, and a rotary evaporator was used to concentrate the extracts to 100 µL. Ethanol (900 μL) was added, and the solutions were then filtered through a 0.45 μm pore size filter membrane. The filtrates were injected into separate vials and analyzed by high-performance liquid chromatography combined with fluorescence detection (Shimadzu Inc., Kyoto, Japan) using Inertsil ODS-P column (250 × 4.6 mm, 5 µm; GL Sciences Inc., Tokyo, Japan).

2.4. OH-PAHs Analysis

For preparation of the airborne particle-bound OH-PAHs, a previously established method was modified [21]. Briefly, the filters were cut into pieces and then placed in glass centrifuge tubes. An internal standard solution was added to each tube. Ethanol/pentane (1:1, v/v) was added for extraction, and the samples were sonicated. The supernatant was collected after centrifugation (652× g). The extraction was repeated three times. The extracts were concentrated with N2 gas to approximately 1 mL. Next, the concentrated samples were loaded onto Bond Elut NH2 (Agilent, Santa Clara, CA, USA) that had been prewashed and conditioned with ethanol. The target compounds were eluted with ethanol/water (3:1, v/v). Each elute was filtered through a 0.21 μm pore size syringe filter. For sample analysis, we used a slight modification of an established method [35]. The target analytes were separated by high-performance liquid chromatography using a Kinetex F5 column (100 × 2.1 mm, 5 µm; Phenomenex, Torrance, CA, USA) and detected using tandem-mass spectrometry (LC/MS; Shimadzu Inc.).

2.5. Health Risk Assessment

To assess the endocrine disrupting risks posed by airborne particle-bound PAHs and OH-PAHs, the relative estrogenic and antiestrogenic activities and binding affinity to hER of each compound were calculated using the data reported in a previous study (Tables S1 and S2) [30]. In the previous study, the estrogenic and antiestrogenic activities were assessed using yeast (Saccharomyces cerevisiae) in a two-hybrid assay, and the binding affinity was assessed in a competition binding assay to hERα with 17β-estradiol as a competitor. Relative estrogenic and antiestrogenic activities, and relative binding potential were calculated from atmospheric concentrations simply multiplied with reported relative effective potency of estrogenic activity (REPE), relative effective potency of antiestrogenic activity (REPAE) and relative binding affinity.

2.6. Quality Control and Quality Assurance

A method blank and a matrix-spiked sample were analyzed with each batch of 10 samples. Method blanks contained the target analyte at a concentration of less than the limit of detection (LOD) for PAHs, or between the LOD and 0.08 ng/mL for OH-PAHs. The recovery rates for the spiked samples were 85%–110% (PAHs) and 81%–100% (OH-PAHs). The instrumental LODs for the analytes were 9.1–83 pg/mL (PAHs) and 0.05–1.00 ng/mL (OH-PAHs). PAHs with two to three rings were excluded as target analytes because they had high vapor pressures and were difficult to quantify. Because of the difficulty of peak separation and identification by LC/MS, the following appeared as overlapping peaks: 2OH- and 3OH-Phe, 1OH- and 9OH-Phe, 2OH- and 3OH-Chr, and 6OH-Chr and 3OH-BaA.

2.7. Statistical Analysis

The concentrations of PAHs and OH-PAHs below the LOD were substituted with a value of LOD/2 for further statistical analysis. R software (version 4.0.2, R Development Core Team, Vienna, Austria) was used for statistical analysis with the significance level set at p < 0.05. The differences in PAHs and OH-PAHs concentrations among sampling seasons were examined using Tukey’s test [36]. As a non-parametric statistical test, the correlations between total PAHs concentrations and total OH-PAHs concentrations were examined using Spearman’s rank order correlation coefficient.

3. Results and Discussion

3.1. Atmospheric Concentrations of PAHs and OH-PAHs

The highest concentration of total PAHs was found in winter 2018 (daily average ± standard deviation: 1.00 ± 0.26 ng/m3), followed by spring 2017 (0.86 ± 0.56 ng/m3), autumn 2017 (0.66 ± 0.23 ng/m3), spring 2018 (0.55 ± 0.30 ng/m3), and summer 2017 (0.30 ± 0.09 ng/m3) (Figure 1 and Figure S1) [2,34]. The total PAH concentrations in spring 2017 and winter 2018 were significantly higher than those in summer 2017 (p = 0.02 and <0.01, respectively). These results are consistent with previous reports of higher PAH concentrations with combustion-derived PAH emissions in the winter and lower concentrations in the summer [4,13,14]. The concentrations of atmospheric PAHs in this study were similar to or lower than those in previous atmospheric surveys in Japan [14,30,37], China [38,39], and other Asian countries [40]. The current results were considered to reflect the regulation of emissions in Japan and neighboring countries [4]. Since 1966, the Japanese government has gradually strengthened regulations regarding the emissions of pollutants from new vehicles [41]. Similarly, the Chinese government has adopted a series of policies aimed at reducing emissions [42]. Pyr was the dominant PAH (23.7 ± 5.5%), followed by Flt (19.6 ± 5.3%), BbF (12.6 ± 1.9%), and BgPe (12.1 ± 2.8%). This trend was almost consistent throughout the year. Low molecular weight PAHs (4 rings PAHs) had higher percentages than high molecular weight PAHs (5 and 6 rings PAHs), which was considered to be caused by emission source and differences in the vapor pressures among PAHs.
Similarly, the total OH-PAH concentration was highest in winter 2018 (75.06 ± 23.38 pg/m3), followed by autumn 2017 (57.36 ± 33.84 pg/m3), spring 2017 (49.81 ± 24.12 pg/m3), spring 2018 (36.62 ± 18.33 pg/m3) and summer 2017 (17.08 ± 4.83 pg/m3) (Figure 2 and Figure S2). The total OH-PAH concentration in autumn 2017 was significantly higher than in summer 2017 (p = 0.02), and in winter 2018 was significantly higher than those in summer 2017 and spring 2018 (p < 0.01 and p = 0.03, respectively). Similar to previous studies [22,43], the concentration of total OH-PAHs was approximately 6.9–24.8 times lower than that of the total PAHs (11.93–129.3 and 177.9–1757 pg/m3, respectively). On average, 2OH-Nap was the dominant PAH (20.1 ± 11.7%), followed by 2/3OH-Phe (16.5 ± 3.0%), 1OH-Flu (14.8 ± 3.4%), and 2/3OH-Chr (12.0 ± 2.9%). Atmospheric OH-PAHs (OH-Pyr and OH-Flu) were detected at lower concentrations (1OH-Pyr: 5.53 ± 4.22 and 2OH-Flu: 1.45 ± 1.12 pg/m3) to those previously reported for Nagasaki, Japan, in 1997–1998 (1OH-Pyr: 21.48 ± 14.97 and 2OH-Flu: 10.40 ± 10.71 pg/m3) [21]. The dominant component was 2OH-Nap in spring 2017/2018 (17.4%/22.8%) and summer 2017 (35.9%), but 2/3OH-Phe in autumn 2017 (19.6%) and winter 2018 (15.9%). This was probably caused by differences in the emission source [22,23] and atmospheric conditions [26,27]. Toriba et al. [24] reported more than 40 times lower concentration of 2OH-Nap compared to 2/3OH-Phe and these were considered to be emitted from exhaust gas. Moreover, Avagyan et al. [17] found the same levels of 2/3OH-Phe as 2OH-Nap from wood combustion. The current results may reflect these emission source differences. Compared to the results reported by Toriba et al. [24] for samples collected in Kanazawa in July 2013, the concentration of 2OH-Nap detected in the present study (6.14 ± 2.59 pg/m3 in August) was much higher (20 ± 16 fg/m3 in July in previous study). Additionally, the proportion of 1OH-Pyr in the present study was much lower than in the study by Toriba et al. (8.9% in the present study, and 48% calculated from reported values in the previous study). These differences may reflect differences in the sampling locations between this study (roadside) and the previous study (rooftop of a building on a university campus). To the best of our knowledge, this is the first report on annual seasonal variations of multiple OH-PAHs in the atmosphere in Japan.
The ratio of total PAHs/total OH-PAHs did not show significant differences among the seasons (ratio: 11.5–14.4, p = 0.67–0.99 by Tukey’s test), though these were weak differences between warmer seasons (spring and summer: 15.8–17.8) and colder seasons (autumn and winter: 13.4–13.7). These results are not consistent with those of Lin et al. [22], who reported clear seasonal differences with ratios of 33.3 in the warmer seasons (April–October) and 14.2 in the colder seasons (November–March) in Beijing, China. The samples in the present study were collected near a main road, and the major emission source was traffic. Additionally, compared with China, coal and biomass burning are rarely conducted in cities in Japan. Therefore, this inconsistency probably reflects differences in the emission sources of OH-PAHs.
The total PAH and total OH-PAH concentrations were significantly correlated (Figure 3, p < 0.01, r = 0.885). Additionally, there were significant correlations between Pyr and 1OH-Pyr (p < 0.01, r = 0.920) and BaA + Chr and OH-BaA + OH-Chrs (p < 0.01, r = 0.937). These results indicate that these two groups of chemicals have common emission sources. The OH-PAHs primarily originate as oxidized derivatives of PAHs through the combustion of organic matter, including fossil oil [17,21,25]. They can also form from PAHs after photochemical reactions in the atmosphere [26,27,28,29]. Both pathways likely caused the significant correlations between PAHs and OH-PAHs. Unlike the other PAHs and OH-PAHs, 3OH-BaP was not detected in any of the samples, even though BaP was detected in all analyzed samples (0.04 ± 0.02 ng/m3). This indicated that 3OH-BaP was not contained in the emission source of BaP and/or was unlikely to form from BaP in the atmosphere.
To group and estimate emission sources, principal component analysis (PCA) and diagnostic ratio analysis were conducted. The results of the PCA for PAHs and OH-PAHs found some seasonal differences. For PAHs, spring, summer, autumn 2017 and spring 2018 formed a related group and winter 2018 was relatively separated (Figure S3). Similarly, for OH-PAHs 4 seasons formed a related group and the winter of 2018 was relatively separated (Figure S4). In both cases, winter 2018 was separated from summer, autumn 2017 and spring 2018, and spring 2017 was located between winter 2018 and other 3 seasons. These results likely indicate that these three seasons had similar contamination sources that differed from the winter 2018 samples. In addition, it may be that some abrupt event occurred in spring 2017, especially 29–30 April, such as the long-range transport from China [2]. The PAH diagnostic ratio analysis indicates clear seasonal differences, as well (Figure S5). The ratio of IDP and IDP+BgPe (IDP/(IDP+BgPe)) was used to assess the contribution difference between diesel vehicles and gasoline vehicles [34], and no clear difference was found with the five periods. However, combustion source estimation using the ratio of Flt and Flt+Pyr (Flt/Flt+Pyr) [44] showed a remarkable difference, indicating that fossil fuel combustion was the dominant source in the spring and autumn, while petrogenic sources were dominant in the summer and organic matter combustion was the dominant source in the winter. In Japan, grass, wood, and related organic matter combustion (open burning) has been regulated [45]. The sampling point was located next to a main road in a residential area of Kanazawa city and there are low PAH and OH-PAHs emissions from residential areas. Hayakawa et al. [2] reported that the long-range transport of PM and PAHs from China has relatively high contributions to atmospheric PM and PAH contamination in Kanazawa. Therefore, the current results, which are consistent with the previous study, and seasonal variations in atmospheric PAH and OH-PAH concentrations were caused by local emissions and long-range transportation.
In this study, the detected atmospheric PAHs and OH-PAHs were bound to airborne particles, such as PM2.5. Atmospheric concentrations of PM2.5 were highest in spring 2017 (21.7 ± 12.1 μg/m3), spring 2018 (17.5 ± 9.6 μg/m3), winter 2018 (14.2 ± 2.6 μg/m3), autumn 2017 (12.3 ± 5.1 μg/m3), and summer 2017 (10.6 ± 4.3 μg/m3) [2]. There were no significant seasonal variations (Tukey’s test: p = 0.07–0.99). However, the total atmospheric PAHs and OH-PAHs concentrations were significantly correlated with atmospheric PM2.5 (p < 0.01, r = 0.496 for PAHs; and p = 0.02, r = 0.394 for OH-PAHs; Figure 4a,b). The observed significant correlation between atmospheric PAHs and PM2.5 concentrations is consistent with a previous study conducted in Japan [46]. The correlation between OH-PAHs and PM2.5 was weaker than that between PAHs and PM2.5. In this study, PM2.5 was thought to be released mainly from vehicle exhaust with PAHs and OH-PAH adsorbed on the surface. The correlation weakness observed in the current study suggests that OH-PAHs are formed by secondary mechanisms after release into the atmosphere [27]. Additionally, the weakness of the correlation between the PAHs/OH-PAHs and PM2.5 also suggests the degradation and/or chemical change in target chemicals in the atmosphere and on the filter [47].

3.2. Health Risk Assessment

The relative health risks caused by the endocrine-disrupting activities of atmospheric PAHs and OH-PAHs were evaluated using the relative estrogenic and antiestrogenic activities and binding affinities to hERα, which were reported by Hayakawa et al. [30]. None of the PAHs showed any estrogenic and antiestrogenic activities in the previous study. Therefore, only the estrogenic and antiestrogenic activities of the OH-PAHs were analyzed in the present study.
The total estrogenic and total antiestrogenic activities showed general trends of higher levels in winter and lower levels in summer, as was the case for the atmospheric concentrations. Four-ring PAH derivatives, such as OH-Chrysene and OH-benz[a]anthracene, had particularly high health risks.
The calculated, relative, estrogenic activities of the atmospheric OH-PAHs are shown in Figure 5. The total estrogenic activity was highest in winter 2018 (1.58 ± 0.43), followed by autumn 2017 (1.16 ± 0.72), spring 2017 (1.12 ± 0.66), spring 2018 (0.70 ± 0.44), and summer 2017 (0.28 ± 0.12). The total estrogenic activity in summer was significantly lower than that in spring 2017 (p = 0.04), autumn 2017 (p = 0.03), and winter 2018 (p < 0.01), and that in winter 2018 was significantly higher than that in spring 2018 (p = 0.03). Because of the higher activity of the four-ring OH-PAHs [30], the activities of 2/3OH-Chr and 6OH-Chr/3OH-BaA were dominant in the total activity (72.7 ± 5.6%). In particular, 2/3OH-Chr showed high proportions in all samples (53.6 ± 12.9%). 2OH-Nap was dominant for the total concentration (20.1 ± 11.7%); however, it has low estrogenic activity (0%). The second most dominant component, in terms of total concentration, was 2/3OH-Phe (16.5 ± 3.0%), and it had a similar contribution to total estrogenic activity (16.7 ± 3.7%). The compound 1OH-Pyr had a low contribution to total estrogenic activity (6.3 ± 1.5%), although it appears consistently in all seasons.
The calculated, relative, antiestrogenic activities of atmospheric OH-PAHs are shown in Figure 6. Unlike with estrogenic activity, only OH-PAHs with four rings were antiestrogenic [30]. Therefore, only 2/3OH-Chr and 6OH-Chr/3OH-BaA had health risks (78.6 ± 16.7% and 21.4 ± 16.7%, respectively). Similar to estrogenic activity, the total antiestrogenic activity was highest in winter 2018 (1.02 ± 0.27), followed by spring 2017 (0.72 ± 0.41), autumn 2017 (0.71 ± 0.40), spring 2018 (0.43 ± 0.26), and summer 2017 (0.20 ± 0.08). The differences in total antiestrogenic activity among the seasons was the same as that for the estrogenic activity. In addition, the total estrogenic activity and total antiestrogenic activity were significantly correlated (p < 0.01, r = 0.985).
Among the OH-PAHs, OH-chrysenes and OH-benz[a]anthracene had small contributions to the total concentration (9.8–26.1%). However, their high contributions to estrogenic activity (57.1–83.1%) and antiestrogenic activity (100%) suggest that they pose relatively high health risks for humans. The current results indicated that the health risks of endocrine disruption caused by atmospheric OH-PAHs are higher in the winter than in summer.
The binding potentials of PAHs and OH-PAHs to hER calculated from the reported binding affinities are shown in Figure 7. Several PAHs (BbF, BkF, BaP, BgPe, and IDP) and OH-PAHs (1/2OH-Nap, 1/3OH-Flu, 1/4/9OH-Phe, and 3OH-BaP) were considered to have no binding potential. Similar to the estrogenic and antiestrogenic activities, 2/3OH-Chr had the highest contribution to the total estimated binding potential (26.1 ± 5.5%), followed by 2/3OH-Phe (19.5 ± 4.2%). Unlike the estrogenic and antiestrogenic activities, Flt and Pyr had high contributions to the total, estimated, binding potential (13.8 ± 3.6% and 13.8 ± 4.5%, respectively), and 6OH-Chr/3OH-BaA had a low contribution to the total, estimated, binding potential (5.8 ± 4.5%). OH-PAHs contain a phenol group, which is considered an important functional group for binding with hERα, but not necessary for exhibiting antiestrogenic activity [30]. Previous research has suggested that the binding affinity to hER is not the only factor affecting the estrogenic and antiestrogenic activities. The current results indicate that the high binding potentials of several atmospheric PAHs and OH-PAHs to the hER are not directly connected to their estrogenic and antiestrogenic activities. Instead, the results imply that the endocrine disruption risk with PAHs and OH-PAHs arises because of their relatively high atmospheric concentrations.
OH-PAHs are intermediate metabolites of PAHs in organisms [48], and are of concern because they are endocrine disrupting chemicals [30]. Consequently, exposure to PAHs in the environment could result in endocrine disruption caused by OH-PAHs. However, over time, the OH-PAHs are metabolized into conjugated forms, such as the glucuronic acid conjugated form, by detoxification enzymes in the liver [10,49] and excreted mainly via urine [35,50]. These conjugated forms can be transported through the vascular system, but they generally have higher water solubility and lower toxicity than the non-conjugated forms. Therefore, metabolized OH-PAHs may have low health risks. However, it is also possible that OH-PAHs in the atmosphere can be taken in via inhalation and directly enter the vascular system through alveoli. This means that OH-PAHs from the atmosphere could enter the body without being metabolized in the liver and then may cause endocrine disruption. These atmospheric PAHs and derivatives have high risks for lung cancer and respiratory disease [51,52,53,54]. Therefore, endocrine disruption caused by OH-PAHs can occur both indirectly through exposure to atmospheric PAHs, and directly through inhalation of atmospheric OH-PAHs. Wenger et al. [55] reported that OH-PAHs contributed to the in vitro estrogenicity of ambient particulate matter. Therefore, the monitoring of atmospheric OH-PAHs and the assessment of their estrogenic and antiestrogenic activities should be performed.

3.3. Research Limitations

In this study, all of the OH-PAHs evaluated in a previous study [30] were not analyzed. In future studies, hydroxylated benzo[c]phenanthrene should be analyzed. High molecular weight PAHs in the atmosphere are mainly distributed in the particle phase; however, low molecular weight PAHs are mainly distributed in the gas phase [56]. Therefore, monitoring of atmospheric PAHs and OH-PAHs should involve both the gas and particle phases. Additionally, atmospheric environmental conditions, such as wind direction, humidity, ozone, temperature, and sunlight, are important in the secondary formation of OH-PAHs from PAHs in the atmosphere. To assess the health risks of environmental pollutants via inhalation, uptake factors are also important, in addition to their toxicities. Therefore, the uptake factors of PAHs and OH-PAHs, and especially differences among isomers, should be analyzed in future studies.

4. Conclusions

The concentrations of atmospheric PAHs and OH-PAHs showed strong correlations, with a seasonal increase in winter and a decrease in summer. Additionally, the health risk of each substance was evaluated in terms of estrogenic and antiestrogenic activities, with a focus on endocrine disruption. The current results indicated that the health risks of endocrine disruption caused by atmospheric OH-PAHs are higher in the winter than in the summer. Among the OH-PAHs, OH-chrysenes and OH-benz[a]anthracene had particularly high health risks. This is the first study to focus on the contribution of atmospheric OH-PAHs to endocrine disruption in humans. Further studies are needed for more target substances such as OH-benzo[c]phenanthrene, which show strong antiestrogenic activity, so as to evaluate the uptake factors and the effects of environmental conditions on secondary formation.

Supplementary Materials

The following supporting information can be downloaded at https://0-www-mdpi-com.brum.beds.ac.uk/article/10.3390/app12199469/s1: Figure S1: box plot of seasonal atmospheric concentrations of PAHs in Kanazawa; Figure S2: box plot of seasonal atmospheric concentrations of OH-PAHs in Kanazawa; Figure S3: PCA score plot in which the first two PCs for atmospheric PAHs; Figure S4: PCA score plot in which the first two PCs for atmospheric OH-PAHs; Figure S5: Diagnostic ratios calculated for atmospheric PAHs; Table S1: Estrogenic and antiestrogenic activities of OH-PAHs from a yeast two-hybrid assay conducted by Hayakawa et al., (2007); Table S2: Binding affinities of PAHs and OH-PAHs to hER conducted by Hayakawa et al., (2007). Reference [30] is cited in the supplementary materials.

Author Contributions

Conceptualization, K.H. and H.N.; methodology, M.H. and L.Z.; validation, K.H.; formal analysis, M.H.; investigation, L.Z. and N.T.; resources, K.H. and H.N.; writing—original draft preparation, M.H.; writing—review and editing, K.H., L.Z., N.T. and H.N.; supervision, K.H.; project administration, K.H. and H.N.; funding acquisition, K.H. and H.N. All authors have read and agreed to the published version of the manuscript.

Funding

This research was financially supported by Scientific Research (No. 17H06283) from Japan Society for the Promotion Science, the Environment Research, and the Environment Research and Technology Development Fund (Project Nos. 5-1951) of the Ministry of the Environment, Japan.

Institutional Review Board Statement

Not applicable.

Informed Consent Statement

Not applicable.

Data Availability Statement

Not applicable.

Acknowledgments

We thank Hina Satone and staff of Food Industry Department, Niigata Agro-Food University for the analysis of OH-PAHs. We thank Edward G. Nagato, Faculty of Life and Environmental Sciences, Shimane University for review and editing. We thank Gabrielle David, from Edanz (https://jp.edanz.com/ac: accessed on 22 August 2022) for editing a draft of this manuscript. We thank the Kanazawa City Environment Division Environmental Policy Department for arranging the PM2.5 sampling site.

Conflicts of Interest

The authors declare no conflict of interest.

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Applsci 12 09469 g001 550
Figure 1. Daily atmospheric concentrations of PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
Figure 1. Daily atmospheric concentrations of PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
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Figure 2. Daily atmospheric concentrations of OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
Figure 2. Daily atmospheric concentrations of OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
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Figure 3. Correlations between the atmospheric concentrations of total PAHs and OH-PAHs.
Figure 3. Correlations between the atmospheric concentrations of total PAHs and OH-PAHs.
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Figure 4. Correlation between the atmospheric concentrations of the target compounds and PM2.5: (a) total PAHs versus PM2.5, and (b) total OH-PAHs versus PM2.5.
Figure 4. Correlation between the atmospheric concentrations of the target compounds and PM2.5: (a) total PAHs versus PM2.5, and (b) total OH-PAHs versus PM2.5.
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Figure 5. Daily relative estrogenic activities of atmospheric OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
Figure 5. Daily relative estrogenic activities of atmospheric OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
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Figure 6. Daily relative antiestrogenic activities of atmospheric OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
Figure 6. Daily relative antiestrogenic activities of atmospheric OH-PAHs in April, August, and November 2017, and February and April 2018 in Kanazawa.
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Figure 7. Daily relative binding potentials of atmospheric PAHs and OH-PAHs to the hER in April, August, and November 2017, and February and April 2018 in Kanazawa.
Figure 7. Daily relative binding potentials of atmospheric PAHs and OH-PAHs to the hER in April, August, and November 2017, and February and April 2018 in Kanazawa.
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Honda, M.; Hayakawa, K.; Zhang, L.; Tang, N.; Nakamura, H. Seasonal Variability and Risk Assessment of Atmospheric Polycyclic Aromatic Hydrocarbons and Hydroxylated Polycyclic Aromatic Hydrocarbons in Kanazawa, Japan. Appl. Sci. 2022, 12, 9469. https://0-doi-org.brum.beds.ac.uk/10.3390/app12199469

AMA Style

Honda M, Hayakawa K, Zhang L, Tang N, Nakamura H. Seasonal Variability and Risk Assessment of Atmospheric Polycyclic Aromatic Hydrocarbons and Hydroxylated Polycyclic Aromatic Hydrocarbons in Kanazawa, Japan. Applied Sciences. 2022; 12(19):9469. https://0-doi-org.brum.beds.ac.uk/10.3390/app12199469

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Honda, Masato, Kazuichi Hayakawa, Lulu Zhang, Ning Tang, and Hiroyuki Nakamura. 2022. "Seasonal Variability and Risk Assessment of Atmospheric Polycyclic Aromatic Hydrocarbons and Hydroxylated Polycyclic Aromatic Hydrocarbons in Kanazawa, Japan" Applied Sciences 12, no. 19: 9469. https://0-doi-org.brum.beds.ac.uk/10.3390/app12199469

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