Exposure to PM-bound PAHs: History
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Subjects: Public, Environmental & Occupational Health
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Lu Yang

Particulate matter (PM) is a major factor contributing to air quality deterioration that enters the atmosphere as a consequence of various natural and anthropogenic activities. In PM, polycyclic aromatic hydrocarbons (PAHs) represent a class of organic chemicals with at least two aromatic rings that are mainly directly emitted via the incomplete combustion of various organic materials. Numerous toxicological and epidemiological studies have proven adverse links between exposure to particulate matter-bound (PM-bound) PAHs and human health due to their carcinogenicity and mutagenicity. Among human exposure routes, inhalation is the main pathway regarding PM-bound PAHs in the atmosphere. Moreover, the concentrations of PM-bound PAHs differ among people, microenvironment, and areas. Hence, understanding the behaviour of PM-bound PAHs in the atmosphere is crucial.

  • particulate matter
  • polycyclic aromatic hydrocarbons
  • personal exposure
  • indoor
  • outdoor

1. Introduction

Air pollution has become a mainstream global environmental pollution problem in recent decades [1,2]. Particulate matter (PM) is one of the major factors contributing to air quality deterioration, leading to adverse health effects on humans [3–6]. PM is an extremely complex mixture defined in many ways, including formation pathway, emission source, chemical composition, and PM size [5]. The formation pathway of PM involves the direct release by emission sources into the atmosphere or secondary formation via nucleation, vapour condensation, adsorption, and absorption of gaseous precursors, primary PM, or secondary PM [5]. In terms of the source, natural sources include dust, sea salt, living vegetation, volcanic activity, and forest fires, whereas anthropogenic sources mainly involve combustion processes, including stationary sources (domestic, industrial, and agricultural activities) and mobile sources (vehicles, aircraft, and shipping traffic exhaust) [1][2][3][4][5][6][7][8][9][10][11][12][13][14][15][16][17][18][19][20][21][22][23][24][25][26][27][28][29][30][31][32][33][34][35][36][37][38][39][40][41][42][43][44][45][46][47][48][49][50][51][52][53][54][55][56][57][58][59][60][61][62][63][64][65][66][67][68][69][70][71][72][73][74][75][76][77][78][79][80][81][82][83][84][85][86][87][88][89][90][91][92][93][94][95][96][97][98][99][100][101][102][103][104][105][106][107][108][109][110][111][112][113][114][115][116][117][118][119][120][121][122][123][124][125][126][127][128][129][130][131][132][133][134][135][136][137][138][139][140][141][142][143][144][145][146][147][148][149][150][151][152][153][154][155][156][7,8]. Regarding the chemical composition, PM contains many kinds of inorganic and organic compounds, including water-soluble ions, trace elements, crustal material, elemental carbon, and organic carbon, many of which are harmful to human health [9–13]. According to the size, PM is generally divided into inhalable coarse particles (PM with an aerodynamic diameter smaller than 10 μm, PM10) and inhalable fine particles (PM with an aerodynamic diameter smaller than 2.5 μm, PM2.5) [14]. However, other sizes of PM also can be collected in the atmosphere by using different types of air sampler, such as PM1, PM2, PM2.1, and PM4, which aerodynamic diameters are smaller than 1, 2, 2.1, and 4 μm, respectively [15–18].

PM is of concern not only to researchers but also to the general public. Numerous toxicological and epidemiological studies have proven the adverse links between exposure to PM and health effects [19–26]. The cancer risk resulting from PM exposure has also been demonstrated by the International Agency for Research on Cancer (IARC) [27]. According to the report by the World Health Organization (WHO), PM in outdoor air is responsible for approximately 4.5 million premature deaths every year, or close to 10% of the total deaths on a global scale [28]. Among these deaths, approximately 2 million deaths, which represent approximately 5% of the global total deaths, are due to damage to the lungs and respiratory system directly attributable to PM [28]. Moreover, WHO also reported that almost 3 billion people worldwide still rely on solid fuels for cooking and heating, leading to approximately 4 million people premature deaths due to household indoor air pollution [29], which is almost equal to the deaths caused by outdoor PM pollution.

To protect public health, relatively strict indoor and outdoor air quality standards have been prescribed by WHO. For indoor household fuel combustion, WHO has strongly recommended that the emission rate target of PM2.5 should not exceed 0.23 mg/min under unvented conditions, and 0.80 mg/min under vented conditions [29]. For outdoor air, the annual average PM2.5 and PM10 concentrations cannot exceed 10 and 20 μg/m3, respectively, and 24-h average concentrations that cannot exceed 25 and 50 μg/m3, respectively [30]. Currently, outdoor PM regulation standards also have been implemented in many cities worldwide. Table 1 lists various PM2.5 and PM10 regulation standards set by the governments of several countries according to their national conditions.

Specifically, in North America, the 24-h and annual average PM2.5 concentration standards in the United States of America (USA) are 35 μg/m3 and 12 μg/m3, respectively [3]. Mexico (12 μg/m3) and Canada (8.8 μg/m3) have also defined relatively low standards for the annual average PM2.5 concentration [31,32]. In South America, Brazil determined the final standards of PM2.5 and PM10 in 2018, which 24-h and annual average are the same as those prescribed by WHO [33]. However, Chile had relatively higher standards of PM2.5 (24-h: 50 μg/m3; annual: 20 μg/m3) than these above countries [34]. In Australia, the 24-h average PM2.5 and PM10 concentrations are the same as ones prescribed by the WHO, while the annual average of PM2.5 (8 μg/m3) is lower and of PM10 (25 μg/m3) is higher than the WHO-prescribed levels [35]. In the European Union (EU), Russia, and some Asian countries, the PM2.5 and PM10 regulation standards are mostly higher than those in the above countries. The annual average PM2.5 concentration standards are 15 μg/m3 (Japan), 25 μg/m3 (EU, Russia, South Korea, and Mongolia), 35 μg/m3 (China), and 40 μg/m3 (India), respectively [36–38]. In South Africa, only the PM10 regulation standard has been defined, which does not exceed 40 μg/m3 for the annual average concentration and 75 μg/m3 for the 24-h average concentration [39].

 Table 1. Regulation standards of PM2.5 and PM10 set by several governments.

Continent

Country

PM2.5 (µg/m3)

PM10 (µg/m3)

References

24-h

Annual

24-h

Annual

North America

USA

35

12

150

None

[3]

Mexico

45

12

75

40

[31]

Canada

27

8.8

None

None

[32]

South America

Brazil

25

10

50

20

[33]

 

Chile

50

20

150

50

[34]

Australia

Australia

25

8

50

25

[35]

Africa

South Africa

None

None

75

40

[39]

Europe

EU

None

25

50

40

[37]

Russia

35

25

60

40

[38]

Asia

China

75

35

150

70

[36]

Japan

35

15

None

100

[36]

South Korea

25

25

100

50

[36]

Mongolia

50

25

150

50

[36]

India

60

40

100

60

[36]

 

In PM, polycyclic aromatic hydrocarbons (PAHs) are a class of persistent organic chemicals with at least two aromatic rings, mainly directly emitted as a result of the incomplete combustion of various organic materials, including both natural sources, such as forest fires and volcanic eruptions, and anthropogenic sources, such as the combustion of fossil fuels and biomass [40,41]. Several hundred PAHs have been detected worldwide, and the United States Environmental Protection Agency (US EPA) has classified 16 PAH species in a priority control pollutant list, which can be generally divided into low-molecular-weight PAHs (LMW PAHs, MW < 200 g/mol), medium-molecular-weight PAHs (MMW PAHs, 200 ≤ MW < 250 g/mol), and high-molecular-weight PAHs (HMW PAHs, MW ≥ 250 g/mol) these three categories. LMW PAHs exhibit a relatively high vapour pressure and easily occur in the gaseous phase, whereas HMW PAHs exhibit a much lower vapour pressure than that of LMW PAHs, and they mainly occur in the particle phase [42–47]. The vapour pressure of MMW PAHs is between those of LMW and HMW PAHs [47–50], suggesting that they may occur in both the gaseous and particle phases, and phase partitioning largely depends on factors such as meteorological conditions (their occurrence in the gaseous phase increases at a relatively high ambient temperature, and their occurrence in the particle phase increases at a relatively low ambient temperature) [48,51]. In contrast, a previous study has reported that the half-lives of PAHs range from a few hours to days, and the half-lives of MMW and HMW PAHs are longer than that of LMW PAHs [52], which indicates that particulate matter-bound (PM-bound) PAHs could be transported across long distances to other regions worldwide before attenuation [53–57].

PAHs are widely known for their carcinogenicity, mutagenicity and toxicity, and they pose a serious threat to human health [58,59]. A previous study has noted that with increasing MW, the carcinogenicity and acute toxicity of PAHs increase and decrease, respectively [60]. However, LMW and MMW PAHs can react with other gaseous air pollutants, such as ozone (O3, which is a strong oxidizing agent that can damage human lung function, thus threatening human health [61]) and NOx, to produce derivatives with a relatively low vapour pressure that more easily occur in the particle phase than their parent PAHs, and their mutagenicity and toxicity may be higher than those of the parent PAHs [62,63]. In addition to benzo[a]pyrene (BaP), which is classified in Group 1 (carcinogenic to humans), seven species are classified in Group 2A (probably carcinogenic to humans), and twenty-five species are classified in Group 2B (possibly carcinogenic to humans) [27,64,65]. Moreover, several emission sources in outdoor air, including coal combustion, coal tar pitch, coke production, diesel engine exhaust, tobacco smoke, and wood dust, are classified in Group 1 by the IARC [27], which may release many PAHs and derivatives. Due to these harmful effects on human health, it is necessary to clarify the concentrations, compositions, and major contributors of PAHs in the atmosphere.

2. Concentrations of Atmospheric PM-Bound PAHs

2.1. Personal Exposure to PM-Bound PAHs

Research data have indicated that several factors, including conditions of the body, exposure routes, and environmental conditions, can influence the mechanisms of PAHs that are absorbed or adsorbed by human bodies [70]. PAHs may affect human health via inhalation, ingestion, and dermal (skin) exposure, with inhalation being the main exposure pathway for PM-bound PAHs [70]. Exposure varies from person to person due to variations in various physical factors such as the breathing rate. Table 2 lists several studies of personal exposure to PM-bound PAHs in different cities, which have focused on different participants, including non-occupational and occupational exposure.

Table 2. Personal exposure concentration (average and/or range, ng/m3) of PM-bound PAHs in several studies.

Participants

PM

PAHs

Period

Concentration

Country, City

Rural residents

PM

28

July

655 ± 250

China, Laiyang [71]

Rural residents

PM2.5

10

9–12 March 2013

4.2–224

Thailand, Lampang [72]

Health residents

PM2.5

26

2014–2016

1.7 (0.4–5.2)

China, Hongkong [73]

Residents

PM2.5

16

2015–2018

8.27

China, Zhuhai [74]

Residents

PM2.5

16

2014–2017

11.9

China, Wuhan [74]

Children

PM2.5

8

11 April–9 May 2012

15 July–3 November 2012

0.65

0.63

Italy, Rome [75]

Children

PM2.5

16

17 May–23 June 2010

8 November–13 December 2010

27.31

58.18

China, Tianjin [76]

COPD patients

PM2.5

16

June 2017–October 2018

186.85

China, Harbin [77]

Drivers

PM4

12

2015–2018

9.97

Greece, Thessaloniki [15]

Office workers

PM2

13

2015

4.0 ± 2.3

Australia, Canberra [16]

Office workers

PM2.5

8

6–13 March 2009

10–19 June 2009

15.19 ± 15.15

3.04 ± 1.38

Czech Republic, Ostrava [78]

Policemen

PM2.5

8

8–20 February 2009

17–27 May 2009

4.27 ± 2.95

1.03 ± 0.61

Czech Republic, Prague [78]

Policemen

PM2.5

8

2–6 March 2009

6–10 June 2009

39.08 ± 17.33

4.27 ± 1.99

Czech Republic, Karvina [78]

Housewife

PM2.5

19

4–21 November 2016

310 ± 443

China, Xingping [79]

Housewife

PM2.5

19

January 2018

116 (32–224)

China, Xi’an, [80]

Highway toll station workers

PM2.5

16

March–May 2014

319.90

China, Tianjin [81]

Newsagent

PM

16

2013

5570

Iran, Tehran [82]

Seafarers

PM

32

July 2016

760–8400

Sweden [83]

Chinese kitchen worker

PM

16

4 September–1 November 2014

1794–12,108

China, Taiwan [84]

 

Regarding non-occupational exposure, residents living in rural areas [71,72] generally inhale higher PM-bound PAH concentrations (4.2–655 ng/m3) than residents living in urban areas (0.4–11.9 ng/m3) [73,74]. This is because solid fuels such as coal and wood are still mainly used by rural residents for cooking and heating an many PAHs are emitted from these sources due to their low combustion efficiency, while urban residents mostly use clean fuels such as liquefied petroleum gas and natural gas. Thus, children living in Italy exhibited very low exposure levels and slight seasonal variation (0.65 and 0.63 ng/m3) [75], whereas children living in China exhibited relatively high exposure levels and a relatively large seasonal variation (27.31 and 58.18 ng/m3) [76]. A possible reason is the different urban type, whereby Rome (Italy) is a typical commercial city and Tianjin is one of the largest industrial centers of China, leading to the background concentration of PM-bound PAHs in Tianjin being higher than that in Rome [75,76]. In addition, the heating systems in Tianjin can lead to more PM-bound PAHs being released into the atmosphere in the winter [76]. Harbin is a large industrial city in northeast China, and it has been reported that the annual average PM-bound PAH exposure concentration in patients with chronic obstructive pulmonary disease is 186.85 ng/m3 [77], which is much higher than that seen in residents living in Hong Kong, Zhuhai, and Wuhan, China [73,74]. In Harbin, the annual temperature range can be up to 60°C and the heating period can account for half of the year, thus, coal combustion for heating is also a large factor, in addition to industrial coal combustion, leading to the annual average concentration of PM-bound PAHs being higher than other cities in South and Central China [77].

Regarding occupational exposure, drivers, office workers, and policemen, as indicated in Table 2, did not exhibit very high PM-bound PAH exposure concentrations (1.03–39.08 ng/m3) but the values differed among cities and revealed seasonal variations [15,16,78]. In addition to the low background concentrations of PM-bound PAHs in cities, a possible reason is that few direct emission sources were impacting the above professionals in space. However, both housewives and highway toll station workers exhibited relatively high exposure levels [79–81], which were several times higher than those for the professionals described above. The high exposure levels for housewives (116 and 310 ng/m3) mainly occurred due to the use of fuel for cooking, while those for highway toll station workers (319.90 ng/m3) largely occurred due to the inhalation of traffic-related PAHs, such as vehicle exhaust [79–81]. Traffic emissions are also a major source for exposure for newsagents because their workplaces are usually located near roads. However, the exposure levels of PM-bound PAHs for newsagents were much higher in Tehran (Iran) (5570 ng/m3) [82] than those measured for highway toll station workers in Tianjin (China) [81]. A possible reason could be that the pollution levels may be higher in Tehran than that in Tianjin, and the background PM-bound PAH concentrations may be very high. Moreover, studies have revealed very high exposure levels of PM-bound PAHs for seafarers and Chinese kitchen workers, ranging from approximately 760 to 12,108 ng/m3 [83,84]. The high exposure levels for the former occur due to ship engine emissions and their long residence times on ships, while those for the latter occurred due to factors such as the use of various fuels, food type, cooking location and methods. Also, many PM-bound PAHs contained in food can be ingested by the human body through eating [85,86].

The results found that background concentrations in different cities could affect the exposure levels of residents, and the indoor microenvironment could also influence the exposure level because most people spend a lot of time indoors. In addition, the characteristics of the PAH phase distribution determine that the PM-bound PAH concentration is higher during cold periods than that during warm periods. Moreover, people with various jobs exhibit different exposure levels, especially regarding occupational exposure, and workers are exposed to very high health risks in spaces with high PM-bound PAH concentrations. Overall, although the PM type and determined PAH number differed among the above studies, personal exposure to PM-bound PAH depends on many complex factors, including indoor microenvironments, outdoor environments, daily activities (such as jobs), and body conditions.

2.2. Indoor Concentrations of PM-Bound PAHs

Indoor air quality is crucial to human health because people spend more than 80% of their time indoors [73]. The indoor PM-bound PAHs in different microenvironments are emitted by many sources, including cooking, smoking, heating, stoves, chemical spraying, machines such as laser printers, and outdoor sources. Studies have noted that indoor air may even be worse than the outdoor air with the PM-bound PAHs [71,87]. Table 3 summarizes the indoor concentrations of PM-bound PAHs in different microenvironments in certain cities in recent years. Some studies have simultaneously examined personal exposure, as described in Section 2.1. [71,73,79,80,83,84].

Regarding residential indoor air, the PM-bound PAH concentrations in rural households (738 ± 321 ng/m3) were much higher than those in urban households (0.186–276 ng/m3), and the concentrations observed in northern Chinese cities (15–276 ng/3) were higher than those observed in southern Chinese cities (1.0–7.3 ng/m3) [71,73,79,80,88], similar to those of personal exposure (Section 2.1.). The PM-bound PAH concentrations in households in Jeddah (Saudi Arabia) (18.5 ng/m3) [89] were comparable to those in Bursa (Turkey) (22 ng/m3) [90], while in Madrid (Spain) [91] they were much lower (0.186 ng/m3) than in the above cities. In both the Jeddah and Bursa studies air samplers were placed in the living room of residences, while in Madrid, Spain, it was placed on the first floor of a residential building [89–91]. The different locations of air samplers could lead to different PM-bound PAH concentration measurements. On the other hand, a high concentration of PM-bound PAHs (318 ± 314 ng/m3) was observed in infants’ rooms in Harbin (China) in winter, was not only suggested a high health risk for infants, but also for people who staying in the room [92].

Regarding school indoor air, Table 3 does not indicate a large concentration difference in Wuhan (China), among university dormitories (31.3 ng/m3), laboratories (27.0 ng/m3), and offices (32.4 ng/m3) [93]. The annual average concentration differences observed in universities between dormitories and offices in Beijing (34.1 and 32.1 ng/m3) were also not large [88], and comparable to those seen in Wuhan [93]. However, the concentrations observed in university offices in Jeddah (12.7 ± 5.1 ng/m3) [89] were lower than those observed in university offices in Beijing, consistent with the household results. Moreover, the PAH concentrations observed at the laboratories and offices in Harbin (115 and 96.6 ng/m3) in winter [94] were approximately three times higher than those observed in Beijing and Wuhan [88,93]. In contrast, the PM-bound PAH concentrations in primary or secondary classrooms ranged from 0.45 to 29.83 ng/m3, and the concentrations were comparable across Beijing, Warsaw and Gliwice (Poland), and Porto (Portugal), but they were very low in São Paulo (Brazil) [17,95–97]. Different from other sampling cities, São Paulo is close to the Equator and its hot climate can greatly influence the gaseous/particles phase distribution of PAHs. Moreover, the shorter sampling period used in São Paulo than in the other cities also had an impact on the PM-bound PAHs results [96].

Regarding public indoor air, the lowest concentration level (2.39–7.4 ng/m3) was observed in shopping malls in Islamabad (Pakistan), bakeries in Bari (Italy) and hotels in Jeddah [89,98,99]. This may occur because these public places contained few PAH sources. The second-highest concentration level (39.58–155.11 ng/m3) was observed in hotels in Jinan (China) in public bars in Warri (Nigeria) and in office buildings in Changchun (China) [100–102]. The higher concentration of PM-bound PAHs in hotels in Jinan than that in hotels in Jeddah possibly occurred due to the local background urban concentrations, and is consistent with the results for Beijing and Jeddah [88,89]. The relatively high concentration in public bars may be the result of factors such as tobacco smoking, while in office buildings, this may occur due to emissions from machines such as printers [103]. The seasonal differences observed in office buildings may occur due to the gaseous/particles phase distribution of PAHs at different temperatures and various outdoor pollution sources via window opening or ventilator operation [102]. The highest concentration level (550–39,000 ng/m3) was observed in fire stations, ships, and Chinese kitchens [83,84,104]. Smoke originating from fires contains a large number of PM-bound PAHs, which could be adsorbed onto the helmets and clothes of firefighters and transported to fire stations [104]. The high concentrations observed in ships and Chinese kitchens determine the personal exposure levels of seafarers and Chinese kitchen workers, respectively, because of engine fuel combustion and cooking, respectively [83,84]. However, the lower concentration in kitchens than the personal exposure concentration for kitchen workers possibly occurred because the air sampler was likely not located close to the cooking bench for safety reasons (high temperature) [84].

Table 3. Indoor concentrations (average and/or range, ng/m3) of PM-bound PAHs in several studies.

Place

Country, City

PM

PAHs

Periods

Concentration

Rural households

China, Laiyang [71]

PM

28

July

738 ± 321

Households

China, Hongkong [73]

PM2.5

26

2014–2016

3.0 (1.0–7.3)

China, Xingping [79]

PM2.5

19

4–21 November 2016

211 ± 120

China, Xi’an [80]

PM2.5

19

January 2018

92 (15–276)

China, Beijing [88]

PM2.5

16

December 2014–February 2016

39.8

Saudi Arabia, Jeddah [89]

PM10

13

-

18.5 ± 11.2

Turkey, Bursa [90]

PM

16

July 2014– January 2015

22

Spain, Madrid [91]

PM10

14

May 2017–April 2018

0.186

Infant room

China, Harbin [92]

PM

16

December 2013–March 2014

318 ± 314

University

(dormitory)

China, Beijing [88]

PM2.5

16

December 2014–February 2016

34.1

China, Wuhan [93]

PM2.5

16

December 2014–June 2015

31.3

University

(laboratory)

China, Wuhan [93]

PM2.5

16

December 2014–June 2015

27.0

China, Harbin [94]

PM2.5

16

January 2015

115

University

(office)

China, Beijing [88]

PM2.5

16

December 2014–February 2016

32.1

Saudi Arabia, Jeddah [89]

PM10

13

-

12.7 ± 5.1

China, Wuhan [93]

PM2.5

16

December 2014–June 2015

32.4

China, Harbin [94]

PM2.5

16

January 2015

96.6

Classroom

China, Beijing [95]

PM2.5

12

October 2016–March 2017

29.83

Brazil, São Paulo [96]

PM

15

7–11 November 2016

0.45

Poland, Warsaw [17]

PM1

16

April–June 2015

10.9

Poland, Gliwice [17]

PM1

16

April–June 2015

21.6

Portugal, Porto [97]

PM2.5

18

March–May 2014

5.03–23.6

Shopping malls

Pakistan, Islamabad [98]

PM2.5

16

February–April 2014

2.39 ± 1.45

Bakery

Italy, Bari [99]

PM2.5

7

7–19 April 2013

7.4

Hotels

Saudi Arabia, Jeddah [89]

PM10

13

-

6.3 ± 1.3

China, Jinan [100]

PM2.5

19

January 2016

39.58–115.63

Public bars

Nigeria, Warri [101]

PM

16

-

43.43–155.11

Office building

China, Changchun [102]

PM2.5

16

April–October 2018

December 2017–April 2018

48.6

67.9

Fire station

Poland, North Poland [104]

PM4

15

September 2018

1882–5924

Ship

Sweden [83]

PM

32

July 2016

550–39,000

Chinese kitchen

China, Taiwan [84]

PM

16

4 September–1 November 2014

1648–5342

In contrast to personal exposure studies, air samplers remain fixed indoors rather than being portable. Hence, the indoor concentrations of PM-bound PAHs are related to the type of space and sampler location. Different space uses to determine whether direct sources of PAH emissions occur, and different air sampler locations may determine whether more or less PAHs are collected. However, the air is exchanged between indoor and outdoor environments via opening windows or operating fans. Therefore, the air exchange speed and local outdoor concentration exert some considerable impacts on the indoor concentration. For example, if PM-bound PAHs are notably generated indoors and the air exchange speed is low, the indoor concentrations may be higher than the outdoor concentrations (such as in Chinese kitchens [84]). Although the PM type and determined PAH number differed among the above studies, the resulting impacts of PM-bound PAHs concentration levels were not large for the overall characteristics of different microenvironments, e.g., the concentrations were much higher in kitchens than those in classrooms, and higher in rural households than those in urban households.

2.3. Outdoor Concentrations of PM-Bound PAHs

The atmospheric behaviours of outdoor PM-bound PAHs are more complex because they are not only dependent on various direct emission sources including industrial and traffic emission [105], wood and biomass burning [59], but also many complex atmospheric physical and chemical factors, including interactions with other pollutants, photochemical degradation, and dry and wet deposition [106,107]. Due to the different conditions in countries worldwide, the concentrations of PM-bound PAHs also vary among regions. In this review, relevant reports (in English) were retrieved on outdoor PM-bound PAHs worldwide in recent years, and the available data are summarized in Table 4.

As indicated in Table 4, Auckland (New Zealand) attained the lowest annual average concentration of outdoor PM-bound PAHs (0.31 ± 0.19 ng/m3) [108] among all the cities/countries. Cities in the Americas also attained relatively low overall average concentrations levels of outdoor PM-bound PAHs, which ranged from 0.84 to 10.2 ng/m3 [67,109–113]. In Europe, the observations in A Coruña (Spain) (7.56 ng/m3) [114] exhibited a lower annual average concentration than that in cities in southern Spain (26.2 ng/m3) [115]. A seasonal variation was observed in Milan (Italy) (0.40 and 72.8 ng/m3) [116] and Wadowice (Poland) (10.5 and 80.6 ng/m3) [117] which was larger during cold periods than that during warm periods. During cold periods, the concentration of PM-bound PAHs in Nicosia (Cyprus) (1.62 ng/m3) [118] was the lowest among European cities. The concentrations observed in Brno (Czech Republic) (20.7 ng/m3) [119] were comparable to those observed in Zagreb (Croatia) (25.4 ng/m3) [120], whereas the concentrations in Sarajevo (Bosnia-Herzegovina) (64.8 ng/m3) [120] were much higher than those in Zagreb, even though the samples were collected during the same period. During warm periods, slight differences in the PM-bound PAH concentration were observed among Moscow (1.32–7.68 ng/m3), St. Petersburg (1.71–6.30 ng/m3), and Kazan (2.95–9.61 ng/m3) in Russia [121]. Although most countries in Europe as shown in Table 5 are developed countries, the overall concentration levels of outdoor PM-bound PAHs were relatively higher than those in the Americas. One reason was the different sampling periods that the concentrations were mostly annual average in the Americas, whereas were mostly seasonal average in Europe. In addition, different meteorological conditions in different periods can lead to the concentration differences of PM-bound PAHs. Moreover, the higher population density in Europe (~73/km2) than in the Americas (~21/km2), with more human activities including industrial and traffic emission also can increase PM-bound PAHs concentrations.

Table 4. Outdoor PM-bound PAHs concentrations (average and/or range, ng/m3) in some cities worldwide.

Country, City

PM

PAHs

Periods

Concentration

Oceania

New Zealand, Auckland [108]

PM2.5

15

2016–2017

0.31 ± 0.19

Americas

Mexico, South Mexico [67]

PM2.5

24

November 2016–March 2017

4.82 ± 1.97

Peru, Arequipa [109]

PM2.5

PM10

14

January–December 2018

7.4 ± 2.3

9.6 ± 3.9

Argentina, Cordoba [110]

PM10

14

August 2011–August 2013

4.5 ± 4.34

Brazil, Belo Horizonte [111]

PM2.5

16

May 2017–April 2018

1.68–6.24

Canada, Toronto [112]

PM10

17

August 2016–August 2017

10.2 ± 2.5

US, Washington [113]

PM10

19

April 2016–September 2018

0.84

Europe

Spain, Coruña [114]

PM10

12

Januray–December 2017

7.56

Spain, South Spain [115]

PM2.5

PM10

16

July 2014–June 2015

23.0

26.2

Italy, Milan [116]

PM2.5

-

December 2018–February 2019

May–July 2019

72.8 ± 16.6

0.40 ± 0.07

Poland, Wadowice [117]

PM10

9

March 2017

August 2017

80.6

10.5

Cyprus, Nicosia [118]

PM2.5

50

January–March 2018

1.62

Czech Republic, Brno [119]

PM1

15

January–February 2017

20.7

Croatia, Zagreb [120]

PM10

10

December 2017–February 2018

25.4

Bosnia and Herzegovina, Sarajevo [120]

PM10

10

December 2017–February 2018

64.8

Russia, Moscow [121]

PM10

9

June–July 2018

1.32–7.68

Russia, St. Petersburg [121]

PM10

9

June–July 2018

1.71–6.30

Russia, Kazan [121]

PM10

9

June–July 2018

2.95–9.61

Africa

Algeria, Algiers [122]

PM10

22

June–September 2016

7.47 ± 1.21

South Africa, Pretoria [123]

PM2.5

16

June–July 2016

4.11

Rwanda, Kigali [124]

PM2.5

15

May–June 2017

52.7

Asia

China, Hongkong [73]

PM2.5

26

2014–2016

3. 9 (1.5–9.6)

China, Xi’an [127]

PM2.5

PM10

16

December 2016–December 2017

63.1 (14.3–266)

66.8 (9.69–349)

China, Shanghai [18]

PM2.1

9

July 2017

January 2018

1.36 ± 0.20

7.72 ± 3.33

China, Beijing [128]

PM10

15

January 2017

98.1 ± 48.2

China, Zhengzhou [128]

PM10

15

January 2017

77.9 ± 29.6

China, Guangzhou [129]

PM2.5

16

June–July 2016

November–December 2016

5.49

10.5

China, Taiyuan [129]

PM2.5

16

June–July 2016

November–December 2016

29.5

197

China, Jinan [130]

PM2.5

18

March–December 2016

39.8 (8.18–246)

China, Shanxi [131]

PM10

PM2.1

17

January–February 2017

1056 ± 315

937 ± 294

China, Urumqi [132]

PM2.5

16

September 2017–September 2018

448

China, Chengdu [133]

PM10

16

March 2015–February 2016

82.0 ± 64.8

China, Changchun [134]

PM2.5

15

October–November 2016

81.4 ± 46.0

China, Harbin [135]

PM2.5

16

June 2017–May 2018

86.9

China, Lanzhou [136]

PM2.5

9

July 2017–October 2018

9.86

Japan, Kanazawa [137]

PM2.5

9

April 2017–February 2018

0.69

Japan, Chiba [138]

PM2.5

21

June 2016–October 2017

2.9

Japan, Kirishima [139]

PM2.5

9

November–December 2016

1.32 (0.36–2.90)

South Korea, Seoul [140]

PM2.5

14

January–December 2018

5.6 ± 7.9

South Korea, Gwangju [141]

PM2.5

17

October 2016–April 2017

1.04–29.5

Vietnam, Hanoi [142]

PM10

9

2016–2018

8.51

Singapore, Singapore [143]

PM10

16

May 2015–June 2016

0.68–5.97

Malaysia, Lumpur [144]

PM2.5

16

June 2015–May 2016

2.04 ± 0.28

Thailand, Chiang Mai [145]

PM2.5

8

February–April 2016

5.88 ± 1.97

Qatar, Doha [146]

PM2.5

PM10

36

May–December 2015

0.56

0.72

Lebanon, Beirut [147]

PM2.5

15

December 2018–October 2019

0.95

Mongolia, Ulaanbaatar [148]

PM10

15

January 2017

March 2017

September 2017

131–773

22.2–531

1.4–54.6

Pakistan, Islamabad [149]

PM2.5

PM10

16

January–September 2017

25.7 ± 12.0

40.1 ± 16.8

India, Jamshedpur [150]

PM2.5

16

December 2016–February 2017

March–May 2017

109 ± 18.2

81.1 ± 13.3

India, Delhi [151]

PM2.5

14

December 2016–December 2017

753 ± 252

India, Haryana [151]

PM2.5

14

December 2016–December 2017

259 ± 64.6

India, Uttar Pradesh [151]

PM2.5

14

December 2016–December 2017

535 ± 143

India, Pune [152]

PM2.5

PM10

16

March 2015–March 2016

342.4 ± 14.3

446.1 ± 25.6

Iran, Tehran [153]

PM10

16

February–March 2018

213 ± 145

Iran, Bushehr [154]

PM2.5

16

December 2016–September 2017

0.66–142.3

In Africa, the limited available studies involving short-term observations of outdoor PM-bound PAHs showed that the average concentrations in Algiers (Algeria) (7.47 ± 1.21 ng/m3) [122] and Pretoria (South Africa) (4.11 ng/m3) [123] were much lower than that in Kigali (Rwanda) (52.7 ng/m3) [124]. On the other hand, Ofori et al. [125] summarized the studies on the PAHs from 2005 to 2019 in Africa. Only 14 reports among 121 papers focused on the indoor and outdoor PM-bound PAHs. Of those, only four papers related to outdoor PM-bound PAHs in the last five years, in which air samples were collected in Rwanda (this report), Tunisia (urban area, 2.8 ± 3.4 ng/m3), Nigeria (industrial area, 73–143 ng/m3), and Egypt (rural: 323 ng/m3; suburban: 503 ng/m3; urban: 417 ng/m3) [125].

In Asia, China has the largest population globally. Since the severe haze event in 2013, researchers have paid special attention to the atmospheric environment. Yan et al. [126] reviewed 270 studies of PM-bound PAHs in 67 cities from 2001 to 2016, which covered seven typical regions, including North, Northwest, Northeast, East, Central, South, and Southwest China. It has been reported that the annual mean concentrations of PM-bound PAHs in these cities range from 3.35 to 910 ng/m3, and those in the northern regions are higher than those in the southern regions [126]. In this paper, Table 4 lists the concentrations of PM-bound PAHs in several major Chinese cities collected within 5 years. The results for these cities indicate average concentrations ranging from 1.36 to 1056 ng/m3 [18,73,127–136], and the seasonal and regional differences are consistent with the results reported by Yan et al. [126]. In other Asian cities, relatively low concentrations levels of PAHs have been observed in Japan [137–139], South Korea [140,141], Vietnam [142], Singapore [143], Malaysia [144], Thailand [145], Qatar [146], and Lebanon [147], ranging from 0.56 ng/m3 (Doha, Qatar) to 29.5 ng/m3 (Gwangju, South Korea). Relatively high PAH concentration levels have been observed in Mongolia [148], Pakistan [149], India [150–152], and Iran [153,154] ranging from 0.66 ng/m3 (Bushehr, Iran) to 773 ng/m3 (Ulaanbaatar, Mongolia), which countries are the most polluted areas in the world. In Mongolia, wood and biomass burning for cooking and heating were the largest emission sources of PM-bound PAHs in the winter [148], while in India and Iran, traffic emission was the main contributor for PM-bound PAHs through the years [151,153]. The air pollutants in Iran were also largely influenced by the air masses long-range transported from Iraq and other Middle East areas [155]. Moreover, due to the special location, frequent recirculation of air masses resulted in the increased residence time of PM in Iran and lead to high air pollutants levels [156].

According to the simultaneously obtained PM2.5 and PM10 data, as summarized in Table 4 [109,115,127,131,146,149,152], the concentrations of PM2.5-bound PAHs accounted for approximately 65% ~ 95% of the total PM-bound PAHs, proving that PAHs mostly occur in PM with a small size, which is consistent with previous studies reporting that the adsorption of PAHs depends on the PM type [69]. Although high outdoor concentrations of PM-bound PAHs were observed in certain cities, they were not as high as the personal exposure concentrations for seafarers and kitchen workers nor the indoor concentrations observed in ships and Chinese kitchens [83,84]. This occurs because expansive outdoor spaces facilitate the diffusion and dilution of air pollutants. Moreover, outdoor air pollutants easily degrade or transform in the atmosphere due to various meteorological conditions and other reasons. Moreover, although there have been a large number of environmental observation studies on PM-bound PAHs over the past few decades, different countries and regions have focused on different PM-bound PAH research aspects. Some major cities have published many reports on various PAH species, while other cities have published no research reports on PAHs at all.

This entry is adapted from the peer-reviewed paper 10.3390/ijerph18042177

References

  1. Lelieveld, J.; Evans, J.S.; Fnais, M.; Giannadaki, D.; Pozzer, A. The contribution of outdoor air pollution sources to premature mortality on a global scale. Nature 2015, 525, 367–371.
  2. Shah, A.S.; Langrish, J.P.; Nair, H.; McAllister, D.A.; Hunter, A.L.; Donaldson, K.; Newby, D.E.; Mills, N.L. Global associa-tion of air pollution and heart failure: A systematic review and meta-analysis. Lancet 2013, 382, 1039–1048.
  3. US EPA. National Ambient Air Quality Standards for Particulate Matter; United States Environmental Protection Agency: Washington, DC, USA, 2013.
  4. Esworthy, R. Air Quality: EPA’s 2013 Changes to the Particulate Matter (PM) Standard; Library of Congress, Congressional Re-search Service: Washington, DC, USA, 2013.
  5. Kim, K.-H.; Kabir, E.; Kabir, S. A review on the human health impact of airborne particulate matter. Environ. Int. 2015, 74, 136–143.
  6. WHO. Health Effects of Particulate Matter. Policy Implications for Countries in Eastern Europe, Caucasus and Central Asia; World Health Organization Regional Office for Europe: Copenhagen, Denmark, 2013.
  7. Grivas, G.; Cheristanidis, S.; Chaloulakou, A.; Koutrakis, P.; Mihalopoulos, N. Elemental composition and source apportion-ment of fine and coarse particles at traffic and urban background locations in Athens, Greece. Aerosol Air Qual. Res. 2018, 18, 1642–1659.
  8. Zhang, L.L.; Zhang, X.; Xing, W.L.; Zhou, Q.Y.; Yang, L.; Nakatsubo, R.; Wei, Y.J.; Bi, J.R.; Shima, M.; Toriba, A.; et al. Nat-ural aeolian dust particles have no substantial effect on atmospheric polycyclic aromatic hydrocarbons (PAHs): A laboratory study based on naphthalene. Environ. Pollut. 2020, 263, 114454.
  9. Zhou, Q.Y.; Zhang, L.L.; Yang, L.; Zhang, X.; Xing, W.L.; Hu, M.; Chen, B.; Han, C.; Toriba, A.; Hayakawa, K.; et al. Long-term variability of inorganic ions in TSP at a remote background site in Japan (Wajima) from 2005 to 2015. Chemosphere 2021, 264, 128427.
  10. Jiang, N.; Li, Q.; Su, F.; Wang, Q.; Yu, X.; Kang, P.; Zhang, R.; Tang, X. Chemical characteristics and source apportionment of PM2.5 between heavily polluted days and other days in Zhengzhou, China. J. Environ. Sci. 2018, 66, 188–198.
  11. Zhang, L.L.; Morisaki, H.; Wei, Y.J.; Li, Z.G.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Hu, M.; Shima, M.; et al. Charac-teristics of air pollutants inside and outside a primary school classroom in Beijing and respiratory health impact on children. Environ. Pollut. 2019, 255, 113147.
  12. Zhang, X.; Zhang, L.L.; Yang, L.; Zhou, Q.Y.; Xing, W.L.; Toriba, A.; Hayakawa, K.; Wei, Y.J.; Tang, N. Characteristics of polycyclic aromatic hydrocarbons (PAHs) and common air pollutants at Wajima, a remote background site in Japan. Int. J. Environ. Res. Public Health 2020, 17, 957.
  13. Zhang, H.; Zhang, L.L.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Kazuichi, H.; Toriba, A.; Tang, N. Impact of COVID-19 outbreak on the long-range transport of common air pollutants in KUWAMS. Chem. Pharm. Bull. 2020, in press.
  14. Lee, H.; Honda, Y.; Hashizume, M.; Guo, Y.L.; Wu, C.-F.; Kan, H.; Jung, K.; Lim, Y.-H.; Yi, S.; Kim, H. Short-term exposure to fine and coarse particles and mortality: A multicity time-series study in East Asia. Environ. Pollut. 2015, 207, 43–51.
  15. Karageorgou, K.; Manoli, E.; Kouras, A.; Samara, C. Commuter exposure to particle-bound polycyclic aromatic hydrocarbons in Thessaloniki, Greece. Environ. Sci. Pollut. Res. 2020, doi:10.1007/s11356-020-09475-9.
  16. Wang, X.; Banks, A.P.; He, C.; Drage, D.S.; Gallen, C.L.; Li, Y.; Li, Q.; Thai, P.K.; Mueller, J.F. Polycyclic aromatic hydrocar-bons, polychlorinated biphenyls and legacy and current pesticides in indoor environment in Australia–occurrence, sources and exposure risks. Sci. Total Environ. 2019, 693, 133588.
  17. Rogula-Kozłowska, W.; Kozielska, B.; Majewski, G.; Rogula-Kopiec, P.; Mucha, W.; Kociszewska, K. Submicron parti-cle-bound polycyclic aromatic hydrocarbons in the Polish teaching rooms: Concentrations, origin and health hazard. J. Envi-ron. Sci. 2018, 64, 235–244.
  18. Yang, L.; Zhang, X.; Xing, W.L.; Zhou, Q.Y.; Zhang, L.L.; Wu, Q.; Zhou, Z.J.; Chen, R.J.; Toriba, A.; Hayakawa, K.; et al. Yearly variation in characteristics and health risk of polycyclic aromatic hydrocarbons and nitro-PAHs in urban shanghai from 2010 to 2018. J. Environ. Sci. 2021, 99, 72–79.
  19. Ai, S.; Qian, Z.M.; Guo, Y.; Yang, Y.; Rolling, C.A.; Liu, E.; Wu, F.; Lin, H. Long-term exposure to ambient fine particles asso-ciated with asthma: A cross-sectional study among older adults in six low-and middle-income countries. Environ. Res. 2019, 168, 141–145.
  20. Beelen, R.; Hoek, G.; Raaschou-Nielsen, O.; Stafoggia, M.; Andersen, Z.J.; Weinmayr, G.; Hoffmann, B.; Wolf, K.; Samoli, E.; Fischer, P.H. Natural-cause mortality and long-term exposure to particle components: An analysis of 19 European cohorts within the multi-center escape project. Environ. Health Perspect. 2015, 123, 525–533.
  21. Akhavan, O.; Bijanzad, K.; Mirsepah, A. Synthesis of graphene from natural and industrial carbonaceous wastes. RSC Adv. 2014, 4, 20441–20448.
  22. Akhavan, O.; Ghaderi, E.; Akhavan, A. Size-dependent genotoxicity of graphene nanoplatelets in human stem cells. Bio-materials 2012, 33, 8017–8025.
  23. Akhavan, O.; Ghaderi, E.; Esfandiar, A. Wrapping bacteria by graphene nanosheets for isolation from environment, reacti-vation by sonication, and inactivation by near-infrared irradiation. J. Phys. Chem. B 2011, 115, 6279–6288.
  24. Hashemi, E.; Akhavan, O.; Shamsara, M.; Rahighi, R.; Esfandiar, A.; Tayefeh, A.R. Cyto and genotoxicities of graphene oxide and reduced graphene oxide sheets on spermatozoa. RSC Adv. 2014, 4, 27213–27223.
  25. Akhavan, O.; Ghaderi, E.; Hashemi, E.; Akbari, E. Dose-dependent effects of nanoscale graphene oxide on reproduction ca-pability of mammals. Carbon 2015, 95, 309–317.
  26. Niu, Z.; Liu, F.; Yu, H.; Wu, S.; Xiang, H. Association between exposure to ambient air pollution and hospital admission, in-cidence, and mortality of stroke: An updated systematic review and meta-analysis of more than 23 million participants. En-viron Health Prev. Med. 2021, 26, 1–14.
  27. IARC. Outdoor Air Pollution; IARC Monographs on the Evaluation of Carcinogenic Risks to Humans; International Agency for Research on Cancer: Lyon, France, 2015; Volume 109.
  28. Stanaway, J.D.; Afshin, A.; Gakidou, E.; Lim, S.S.; Abate, D.; Abate, K.H.; Abbafati, C.; Abbasi, N.; Abbastabar, H.; Abd-Allah, F. Global, regional, and national comparative risk assessment of 84 behavioural, environmental and occupational, and metabolic risks or clusters of risks for 195 countries and territories, 1990–2017: A systematic analysis for the global bur-den of disease study 2017. Lancet 2018, 392, 1923–1994.
  29. WHO. Who Guidelines for Indoor Air Quality: Household Fuel Combustion; World Health Organization: Geneva, Switzerland, 2014.
  30. WHO. Air Quality Guidelines: Global Update 2005: Particulate Matter, Ozone, Nitrogen Dioxide and Sulfur Dioxide; World Health Organization: Copenhagen, Denmark, 2006.
  31. INECC. Mexico-Ambient Air Quality Standard; Mexico’s Ecology and Climate Change Institute, Instituto Nacional de Ecologia y Cambio Climatico (INECC): Mexico City, Mexico, 2014.
  32. CAAQS. Canadian Ambient Air Quality Standards (CAAQS) for Fine Particulate Matter (PM2.5) and Ozone. 2020. Available online: https://www.ccme.ca/files/current_priorities/aqms_elements/caaqs_and_azmf.pdf (accessed on 02 July 2020).
  33. Siciliano, B.; Dantas, G.; Silva, C.M.D.; Arbilla, G. The updated Brazilian national air quality standards: A critical review. J. Braz. Chem. Soc. 2020, 31, 523–535.
  34. Díaz-Robles, L.; Saavedra, H.; Schiappacasse, L.; Cereceda-Balic, F. The Air Quality in Chile; Air, Waste Management Associ-ation: Pittsburgh, PA, USA, 2011.
  35. NEPM, Australia. Australia-National Air Quality Standards; National Environment Protection Measure for Ambient Air Qual-ity: Canberra, Australia, 2016.
  36. Sato, K.; Ohara, T. The status of PM2.5 pollution in Asia and direction toward solving the issue. Glob. Environ. Res. 2018, 22, 39–45.
  37. EU. Directive 2008/50/EC of the European Parliament and of the council of 21 May 2008 on ambient air quality and cleaner air for Europe. Off. J. Eur. Union 2008, Available online: http://news.cleartheair.org.hk/wp-content/uploads/2013/02/LexUriServ.pdf (accessed on 12 February 2021)
  38. Lisetskii, F.; Borovlev, А. Monitoring of emission of particulate matter and air pollution using lidar in Belgorod, Russia. Aero-sol Air Qual. Res. 2019, doi:10.4209/aaqr.2017.12.0593.
  39. SANS. South African National Standard (SANS). 2005. Available online: https://www.environment.gov.za/sites/default/files/docs/stateofair_executive_iaiquality_standardsonjectives.pdf (accessed on 02 July 2020).
  40. Zhang, Y.; Tao, S. Global atmospheric emission inventory of polycyclic aromatic hydrocarbons (PAHs) for 2004. Atmos. Envi-ron. 2009, 43, 812–819.
  41. Baek, S.; Field, R.; Goldstone, M.; Kirk, P.; Lester, J.; Perry, R. A review of atmospheric polycyclic aromatic hydrocarbons: Sources, fate and behavior. Water. Air Soil. Pollut. 1991, 60, 279–300.
  42. Ambrose, D.; Ellender, J.; Sprake, C.; Townsend, R. Thermodynamic properties of fluorine compounds. Part 15—Vapour pressures of the three tetrafluorobenzenes and 1,3,5-trichloro-2,4,6-trifluorobenzene. Faraday Transactions 1: Physical Chem-istry in Condensed Phases. J. Chem. Soc. 1975, 71, 35–41.
  43. Coover, M.P.; Sims, R.C. The effect of temperature on polycyclic aromatic hydrocarbon persistence in an unacclimated agri-cultural soil. Hazard. Waste Hazard. Mater. 1987, 4, 69–82.
  44. Goldfarb, J.L.; Suuberg, E.M. Vapor pressures and enthalpies of sublimation of ten polycyclic aromatic hydrocarbons deter-mined via the Knudsen effusion method. J. Chem. Eng. Data 2008, 53, 670–676.
  45. Lee, W.M.G.; Tong, H.C.; Yeh, S.Y. Partitioning model of PAHs between gaseous and particulate phases with consideration of reactivity of PAHs in an urban atmosphere. J. Environ. Sci. Health A 1993, 28, 563–583.
  46. Lei, Y.D.; Chankalal, R.; Chan, A.; Wania, F. Supercooled liquid vapor pressures of the polycyclic aromatic hydrocarbons. J. Chem. Eng. Data 2002, 47, 801–806.
  47. PubChem. National Library of Medicine, National Center of Biotechnology Information. Available online: https://pubchem.ncbi.nlm.nih.gov/ (accessed on 03 July 2020).
  48. Yamasaki, H.; Kuwata, K.; Miyamoto, H. Effects of ambient temperature on aspects of airborne polycyclic aromatic hydro-carbons. Environ. Sci. Technol. 1982, 16, 189–194.
  49. Hoyer, H.; Peperle, W. Vapor pressure measurements on organic compounds and their sublimation heats. Z. Electrochem. 1958, 62, 61–66.
  50. Sonnefeld, W.; Zoller, W.; May, W. Dynamic coupled-column liquid-chromatographic determination of ambient-temperature vapor pressures of polynuclear aromatic hydrocarbons. Anal. Chem. 1983, 55, 275–280.
  51. Hu, H.; Tian, M.; Zhang, L.; Yang, F.; Peng, C.; Chen, Y.; Shi, G.; Yao, X.; Jiang, C.; Wang, J. Sources and gas-particle parti-tioning of atmospheric parent, oxygenated, and nitrated polycyclic aromatic hydrocarbons in a humid city in southwest China. Atmos. Environ. 2019, 206, 1–10.
  52. Ray, D.; Ghosh, S.K.; Raha, S. Impacts of photochemical ageing on the half-lives and diagnostic ratio of polycyclic aromatic hydrocarbons intrinsic to PM2.5 collected from ‘real-world’ like combustion events of wood and rice straw burning. J. Hazard. Mater. 2019, 366, 10–15.
  53. Yang, L.; Tang, N.; Matsuki, A.; Takami, A.; Hatakeyama, S.; Kaneyasu, N.; Nagato, E.G.; Sato, K.; Yoshino, A.; Hayakawa, K. A comparison of particulate-bound polycyclic aromatic hydrocarbons long-range transported from the Asian continent to the Noto peninsula and Fukue island, Japan. Asian J. Atmos. Environ. 2018, 12, 369–376.
  54. Yang, L.; Zhang, L.L.; Zhang, H.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Takami, A.; Sato, K.; Shimizu, A.; Yoshino, A.; et al. Comparative analysis of PM2.5-bound polycyclic aromatic hydrocarbons (PAHs), nitro-NPAHs (NPAHs) and water-soluble inorganic ions (WSIIs) at two background sites in japan. Int. J. Environ. Res. Public Health 2020, 17, 8224.
  55. Singh, D.K.; Kawamura, K.; Yanase, A.; Barrie, L.A. Distributions of polycyclic aromatic hydrocarbons, aromatic ketones, carboxylic acids, and trace metals in arctic aerosols: Long-range atmospheric transport, photochemical degrada-tion/production at polar sunrise. Environ. Sci. Technol. 2017, 51, 8992–9004.
  56. Tang, N.; Hakamata, M.; Sato, K.; Okada, Y.; Yang, X.; Tatematsu, M.; Toriba, A.; Kameda, T.; Hayakawa, K. Atmospheric behaviors of polycyclic aromatic hydrocarbons at a Japanese remote background site, Noto Peninsula, from 2004 to 2014. Atmos. Environ. 2015, 120, 144–151.
  57. Zhang, L.L.; Yang, L.; Zhang, H.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Toriba, A.; Hayakawa, K.; Tang, N. Impact of the COVID-19 outbreak on the long-range transport of particulate PAHs in East Asia. Aerosol Air Qual. Res. 2020, 20, 2035–2046.
  58. WHO. Part II Evaluation of Risks to Human Health, Chapter 5 Organic Pollutants. In Air Quality Guidelines for Europe; WHO Regional Publications, European Series, No. 91; WHO: Geneva, Switzerland, 2000.
  59. Kim, K.-H.; Jahan, S.A.; Kabir, E.; Brown, R.J. A review of airborne polycyclic aromatic hydrocarbons (PAHs) and their hu-man health effects. Environ. Int. 2013, 60, 71–80.
  60. Armstrong, B.; Hutchinson, E.; Unwin, J.; Fletcher, T. Lung cancer risk after exposure to polycyclic aromatic hydrocarbons: A review and meta-analysis. Environ. Health Perspect. 2004, 112, 970–978.
  61. Zheng, Y.; Jiang, F.; Feng, S.; Cai, Z.; Shen, Y.; Ying, C.; Wang, X.; Liu, Q. Long-range transport of ozone across the eastern china seas: A case study in coastal cities in southeastern China. Sci. Total Environ. 2021, 768, 144520.
  62. Bandowe, B.A.M.; Meusel, H. Nitrated polycyclic aromatic hydrocarbons (nitro-PAHs) in the environment—A review. Sci. Total Environ. 2017, 581–582, 237–257.
  63. Taga, R.; Tang, N.; Hattori, T.; Tamura, K.; Sakai, S.; Toriba, A.; Kizu, R.; Hayakawa, K. Direct-acting mutagenicity of ex-tracts of coal burning-derived particulates and contribution of nitropolycyclic aromatic hydrocarbons. Mutat. Res. 2005, 581, 91–95.
  64. IARC. Bitumens and Bitumen Emissions, and Some N- and S-Heterocyclic Polycyclic Aromatic Hydrocarbons; IARC Monographs on the Evaluation of Carcinogenic Risks to Humans; International Agency for Research on Cancer: Lyon, France, 2013; Volume 103.
  65. IARC. Diesel and Gasoline Engine Exhausts and Some Nitroarenes; IARC Monographs on the Evaluation of Carcinogenic Risks to Humans; International Agency for Research on Cancer: Lyon, France, 2013; Volume 105.
  66. Li, Y.J.; Sun, Y.; Zhang, Q.; Li, X.; Li, M.; Zhou, Z.; Chan, C.K. Real-time chemical characterization of atmospheric particulate matter in China: A review. Atmos. Environ. 2017, 158, 270–304.
  67. Amador-Muñoz, O.; Martínez-Domínguez, Y.; Gómez-Arroyo, S.; Peralta, O. Current situation of polycyclic aromatic hy-drocarbons (PAH) in PM2.5 in a receptor site in Mexico City and estimation of carcinogenic PAH by combining non-real-time and real-time measurement techniques. Sci. Total Environ. 2020, 703, 134526.
  68. Cropper, P.M.; Overson, D.K.; Cary, R.A.; Eatough, D.J.; Chow, J.C.; Hansen, J.C. Development of the GC-MS organic aero-sol monitor (GC-MS OAM) for in-field detection of particulate organic compounds. Atmos. Environ. 2017, 169, 258–266.
  69. Zhang, L.L.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Wei, Y.J.; Hu, M.; Zhao, L.X.; Toriba, A.; Hayakawa, K.; et al. Size distribution of particulate polycyclic aromatic hydrocarbons in fresh combustion smoke and ambient air: A review. J. Environ. Sci. 2020, 88, 370–384.
  70. WHO. Ambient Air Pollution: A Global Assessment of Exposure and Burden of Disease; World Health Organization: Geneva, Switzerland, 2016.
  71. Zhang, J.; Liu, W.; Xu, Y.; Cai, C.; Liu, Y.; Tao, S.; Liu, W. Distribution characteristics of and personal exposure with polycy-clic aromatic hydrocarbons and particulate matter in indoor and outdoor air of rural households in northern China. Environ. Pollut. 2019, 255, 113176.
  72. Orakij, W.; Chetiyanukornkul, T.; Chuesaard, T.; Kaganoi, Y.; Uozaki, W.; Homma, C.; Boongla, Y.; Tang, N.; Hayakawa, K.; Toriba, A. Personal inhalation exposure to polycyclic aromatic hydrocarbons and their nitro-derivatives in rural residents in northern Thailand. Environ. Monit. Assess. 2017, 189, 510.
  73. Chen, X.-C.; Chuang, H.-C.; Ward, T.J.; Tian, L.; Cao, J.-J.; Ho, S.S.-H.; Lau, N.-C.; Hsiao, T.-C.; Yim, S.H.; Ho, K.-F. Indoor, outdoor, and personal exposure to PM2.5 and their bioreactivity among healthy residents of Hong Kong. Environ. Res. 2020, 188, 109780.
  74. Mu, G.; Fan, L.; Zhou, Y.; Liu, Y.; Ma, J.; Yang, S.; Wang, B.; Xiao, L.; Ye, Z.; Shi, T. Personal exposure to PM2.5-bound poly-cyclic aromatic hydrocarbons and lung function alteration: Results of a panel study in china. Sci. Total Environ. 2019, 684, 458–465.
  75. Gatto, M.P.; Gariazzo, C.; Gordiani, A.; L’Episcopo, N.; Gherardi, M. Children and elders exposure assessment to parti-cle-bound polycyclic aromatic hydrocarbons (PAHs) in the city of Rome, Italy. Environ. Sci. Pollut. Res. 2014, 21, 13152–13159.
  76. Han, J.; Zhang, N.; Niu, C.; Han, B.; Bai, Z. Personal exposure of children to particle-associated polycyclic aromatic hydro-carbons in Tianjin, China. Polycycl. Aromat. Compd. 2014, 34, 320–342.
  77. Che, C.; Li, J.; Dong, F.; Zhang, C.; Liu, L.; Sun, X.; Ma, L.; Qi, H.; Wang, K. Seasonal characteristic composition of inorganic elements and polycyclic aromatic hydrocarbons in atmospheric fine particulate matter and bronchoalveolar lavage fluid of COPD patients in northeast China. Respir. Med. 2020, 171, 106082.
  78. Svecova, V.; Topinka, J.; Solansky, I.; Rossner, P.; Sram, R.J. Personal exposure to carcinogenic polycyclic aromatic hydro-carbons in the Czech Republic. J. Expo. Sci. Env. Epid. 2013, 23, 350–355.
  79. Li, Y.; Xu, H.; Wang, J.; Ho, S.S.H.; He, K.; Shen, Z.; Ning, Z.; Sun, J.; Li, L.; Lei, R. Personal exposure to PM2.5-bound organic species from domestic solid fuel combustion in rural Guanzhong basin, China: Characteristics and health implication. Chem-osphere 2019, 227, 53–62.
  80. He, K.; Xu, H.; Feng, R.; Shen, Z.; Li, Y.; Zhang, Y.; Sun, J.; Zhang, Q.; Zhang, T.; Yang, L. Characteristics of indoor and per-sonal exposure to particulate organic compounds emitted from domestic solid fuel combustion in rural areas of northwest China. Atmos. Res. 2021, 248, 105181.
  81. Zhao, Y.-J.; Shou, Y.-P.; Mao, T.-Y.; Guo, L.-Q.; Li, P.-H.; Yi, X.; Li, Q.-Q.; Shen, L.-Z.; Zuo, H.-R.; Wang, J. PAHs exposure assessment for highway toll station workers through personal particulate sampling and urinary biomonitoring in Tianjin, China. Polycycl. Aromat. Compd. 2018, 38, 379–388.
  82. Rezaei, F.; Kakooei, H.; Ahmadkhaniha, R.; Azam, K.; Omidi, L.; Shahtaheri, S.J. Inhalation exposure and health risks for newsagents exposed to atmospheric polycyclic aromatic hydrocarbons in Tehran, Iran. Urban Clim. 2018, 24, 796–802.
  83. Strandberg, B.; Österman, C.; Koca Akdeva, H.; Moldanová, J.; Langer, S. The use of polyurethane foam (PUF) passive air samplers in exposure studies to PAHs in Swedish seafarers. Polycycl. Aromat. Compd. 2020, 1–12, doi:10.1080/10406638.2020.1739084.
  84. Wu, M.-T.; Lin, P.-C.; Pan, C.-H.; Peng, C.-Y. Risk assessment of personal exposure to polycyclic aromatic hydrocarbons and aldehydes in three commercial cooking workplaces. Sci. Rep. 2019, 9, 1–11.
  85. Ghorbani, M.; Saleh, H.N.; Barjasteh-Askari, F.; Nasseri, S.; Davoudi, M. The effect of gas versus charcoal open flames on the induction of polycyclic aromatic hydrocarbons in cooked meat: A systematic review and meta-analysis. J Environ. Health Eng. 2020, 18, 345–354.
  86. Afé, O.H.I.; Douny, C.; Kpoclou, Y.E.; Igout, A.; Mahillon, J.; Anihouvi, V.; Hounhouigan, D.J.; Scippo, M.-L. Insight about methods used for polycyclic aromatic hydrocarbons reduction in smoked or grilled fishery and meat products for future re-engineering: A systematic review. Food Chem. Toxicol. 2020, 141, 111372.
  87. Zhang, L.L.; Morisaki, H.; Wei, Y.J.; Li, Z.G.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Hu, M.; Shima, M.; et al. PM2.5-bound polycyclic aromatic hydrocarbons and nitro-polycyclic aromatic hydrocarbons inside and outside a primary school classroom in Beijing: Concentration, composition, and inhalation cancer risk. Sci. Total Environ. 2020, 705, 135840.
  88. Chen, Y.; Li, X.; Zhu, T.; Han, Y.; Lv, D. Pm2. 5-bound pahs in three indoor and one outdoor air in Beijing: Concentration, source and health risk assessment. Sci. Total Environ. 2017, 586, 255–264.
  89. Ali, N. Polycyclic aromatic hydrocarbons (PAHs) in indoor air and dust samples of different saudi microenvironments; health and carcinogenic risk assessment for the general population. Sci. Total Environ. 2019, 696, 133995.
  90. Esen, F.; Kayikci, G. Polycyclic aromatic hydrocarbons in indoor and outdoor air in turkey: Estimations of sources and expo-sure. Environ. Forensics 2018, 19, 39–49.
  91. Madruga, D.G.; Ubeda, R.M.; Terroba, J.M.; dos Santos, S.G.; García‐Cambero, J.P. Particle‐associated polycyclic aromatic hydrocarbons in a representative urban location (indoor‐outdoor) from south Europe: Assessment of potential sources and cancer risk to humans. Indoor Air 2019, 29, 817–827.
  92. Li, H.-L.; Liu, L.-Y.; Zhang, Z.-F.; Ma, W.-L.; Sverko, E.; Zhang, Z.; Song, W.-W.; Sun, Y.; Li, Y.-F. Semi-volatile organic compounds in infant homes: Levels, influence factors, partitioning, and implications for human exposure. Environ. Pollut. 2019, 251, 609–618.
  93. Wang, G.; Wang, Y.; Yin, W.; Xu, T.; Hu, C.; Cheng, J.; Hou, J.; He, Z.; Yuan, J. Seasonal exposure to PM2.5-bound polycyclic aromatic hydrocarbons and estimated lifetime risk of cancer: A pilot study. Sci. Total Environ. 2020, 702, 135056.
  94. Yury, B.; Zhang, Z.; Ding, Y.; Zheng, Z.; Wu, B.; Gao, P.; Jia, J.; Lin, N.; Feng, Y. Distribution, inhalation and health risk of PM2.5 related pahs in indoor environments. Ecotoxicol. Environ. Saf. 2018, 164, 409–415.
  95. Ouyang, R.; Yang, S.; Xu, L. Analysis and risk assessment of PM2.5-bound pahs in a comparison of indoor and outdoor envi-ronments in a middle school: A case study in Beijing, China. Atmosphere 2020, 11, 904.
  96. Pereira, D.C.A.; Custódio, D.; de Andrade, M.d.F.; Alves, C.; de Castro Vasconcellos, P. Air quality of an urban school in são Paulo city. Environ. Monit. Assess. 2019, 191, 659.
  97. Slezakova, K.; Oliveira, M.; Madureira, J.; Fernandes, E.d.O.; Delerue-Matos, C.; Morais, S.; Pereira, M.d.C. Polycyclic aro-matic hydrocarbons (PAH) in Portuguese educational settings: A comparison between preschools and elementary schools. J. Toxicol. Environ. A 2017, 80, 630–640.
  98. Hamid, N.; Syed, J.H.; Junaid, M.; Mahmood, A.; Li, J.; Zhang, G.; Malik, R.N. Elucidating the urban levels, sources and health risks of polycyclic aromatic hydrocarbons (PAHs) in Pakistan: Implications for changing energy demand. Sci. Total En-viron. 2018, 619, 165–175.
  99. Ielpo, P.; Taurino, M.R.; Buccolieri, R.; Placentino, C.M.; Gallone, F.; Ancona, V.; Di Sabatino, S. Polycyclic aromatic hydro-carbons in a bakery indoor air: Trends, dynamics, and dispersion. Environ. Sci. Pollut. Res. 2018, 25, 28760–28771.
  100. Li, Y.; Yang, L.; Chen, X.; Gao, Y.; Jiang, P.; Zhang, J.; Yu, H.; Wang, W. PM2. 5-bound pahs in indoor and outdoor of hotels in urban and suburban of Jinan, China: Concentrations, sources, and health risk impacts. Aerosol Air Qual. Res. 2017, 17, 2463–2473.
  101. Adesina, O.A.; Nwogu, A.S.; Sonibare, J.A. Indoor levels of polycyclic aromatic hydrocarbons (PAHs) from environment tobacco smoke of public bars. Ecotoxicol. Environ. Saf. 2021, 208, 111604.
  102. Bai, L.; Chen, W.; He, Z.; Sun, S.; Qin, J. Pollution characteristics, sources and health risk assessment of polycyclic aromatic hydrocarbons in pm2. 5 in an office building in northern areas, China. Sustain. Cities Soc. 2020, 53, 101891.
  103. Arı, A. A comprehensive study on gas and particle emissions from laser printers: Chemical composition and health risk as-sessment. Atmos. Pollut. Res. 2020, 11, 269–282.
  104. Rogula-Kozłowska, W.; Bralewska, K.; Rogula-Kopiec, P.; Makowski, R.; Majder-Łopatka, M.; Łukawski, A.; Brandyk, A.; Majewski, G. Respirable particles and polycyclic aromatic hydrocarbons at two polish fire stations. Build. Environ. 2020, 184, 107255.
  105. Yang, L.; Suzuki, G.; Zhang, L.; Zhou, Q.; Zhang, X.; Xing, W.; Shima, M.; Yoda, Y.; Nakatsubo, R.; Hiraki, T.; et al. The characteristics of polycyclic aromatic hydrocarbons in different emission source areas in Shenyang, China. Int. J. Environ. Res. Public Health 2019, 16, 2817.
  106. Zhang, L.L.; Tokuda, T.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Xing, W.L.; Wu, Q.; Zhou, Z.J.; Chen, R.J.; Kameda, T.; et al. Char-acteristics and health risks of particulate polycyclic aromatic hydrocarbons and nitro-polycyclic aromatic hydrocarbons at urban and suburban elementary schools in Shanghai, China. Asian J. Atmos. Environ. 2019, 13, 266–275.
  107. Amarillo, A.C.; Carreras, H. Quantifying the influence of meteorological variables on particle-bound PAHs in urban envi-ronments. Atmos. Pollut. Res. 2016, 7, 597–602.
  108. Kalisa, E.; Nagato, E.; Bizuru, E.; Lee, K.; Tang, N.; Pointing, S.; Hayakawa, K.; Archer, S.; Lacap-Bugler, D. Pollution char-acteristics and risk assessment of ambient PM2.5-bound PAHs and NPAHs in typical Japanese and New Zealand cities and rural sites. Atmos. Pollut. Res. 2019, 10, 1396–1403.
  109. Valdivia, A.E.L.; Larico, J.A.R.; Peña, J.S.; Wannaz, E.D. Health risk assessment of polycyclic aromatic hydrocarbons (PAHs) adsorbed in PM2.5 and PM10 in a region of Arequipa, Peru. Environ. Sci. Pollut. Res. 2020, 27, 3065–3075.
  110. Amarillo, A.C.; Mateos, A.C.; Carreras, H. Source apportionment of PM10-bound polycyclic aromatic hydrocarbons by posi-tive matrix factorization in Cordoba city, Argentina. Arch. Environ. Contam. Toxicol. 2017, 72, 380–390.
  111. Dos Santos, R.R.; de Lourdes Cardeal, Z.; Menezes, H.C. Phase distribution of polycyclic aromatic hydrocarbons and their oxygenated and nitrated derivatives in the ambient air of a Brazilian urban area. Chemosphere 2020, 250, 126223.
  112. Jariyasopit, N.; Tung, P.; Su, K.; Halappanavar, S.; Evans, G.J.; Su, Y.; Khoomrung, S.; Harner, T. Polycyclic aromatic com-pounds in urban air and associated inhalation cancer risks: A case study targeting distinct source sectors. Environ. Pollut. 2019, 252, 1882–1891.
  113. Kramer, A.L.; Campbell, L.; Donatuto, J.; Heidt, M.; Kile, M.; Simonich, S.L.M. Impact of local and regional sources of PAHs on tribal reservation air quality in the US pacific northwest. Sci. Total Environ. 2020, 710, 136412.
  114. Sánchez-Piñero, J.; Moreda-Piñeiro, J.; Concha-Graña, E.; Fernández-Amado, M.; Muniategui-Lorenzo, S.; López-Mahía, P. Inhalation bioaccessibility estimation of polycyclic aromatic hydrocarbons from atmospheric particulate matter (PM10): Influ-ence of PM10 composition and health risk assessment. Chemosphere 2020, 263, 127847.
  115. Pastor, R.P.; Salvador, P.; Alonso, S.G.; Alastuey, A.; dos Santos, S.G.; Querol, X.; Artíñano, B. Characterization of organic aerosol at a rural site influenced by olive waste biomass burning. Chemosphere 2020, 248, 125896.
  116. Hakimzadeh, M.; Soleimanian, E.; Mousavi, A.; Borgini, A.; De Marco, C.; Ruprecht, A.A.; Sioutas, C. The impact of biomass burning on the oxidative potential of PM2.5 in the metropolitan area of Milan. Atmos. Environ. 2020, 224, 117328.
  117. Skiba, A.; Styszko, K.; Furman, P.; Dobrowolska, N.; Kistler, M.; Kasper-Giebl, A.; Zięba, D. Polycyclic aromatic hydrocar-bons (PAHs) associated with PM10 collected in Wadowice, south Poland. E3S Web Conf. 2019, 108, 02007.
  118. Iakovides, M.; Iakovides, G.; Stephanou, E.G. Atmospheric particle-bound polycyclic aromatic hydrocarbons, n-alkanes, ho-panes, steranes and trace metals: PM2.5 source identification, individual and cumulative multi-pathway lifetime cancer risk assessment in the urban environment. Sci. Total Environ. 2021, 752, 141834.
  119. Křůmal, K.; Mikuška, P. Mass concentrations and lung cancer risk assessment of PAHs bound to pm1 aerosol in six industri-al, urban and rural areas in the Czech Republic, central Europe. Atmos. Pollut. Res. 2020, 11, 401–408.
  120. Pehnec, G.; Jakovljević, I.; Godec, R.; Štrukil, Z.S.; Žero, S.; Huremović, J.; Džepina, K. Carcinogenic organic content of par-ticulate matter at urban locations with different pollution sources. Sci. Total Environ. 2020, 734, 139414.
  121. Khalikov, I.; Korunov, A. The content of high-molecular polycyclic aromatic hydrocarbons in urban air during FIFA World Cup 2018. Russ. J. Gen. Chem. 2019, 89, 2821–2826.
  122. Rabhi, L.; Lemou, A.; Cecinato, A.; Balducci, C.; Cherifi, N.; Ladji, R.; Yassaa, N. Polycyclic aromatic hydrocarbons, phthalates, parabens and other environmental contaminants in dust and suspended particulates of Algiers, Algeria. Environ. Sci. Pollut. Res. 2018, 25, 24253–24265.
  123. Morakinyo, O.M.; Mukhola, M.S.; Mokgobu, M.I. Concentration levels and carcinogenic and mutagenic risks of PM2.5-bound polycyclic aromatic hydrocarbons in an urban–industrial area in South Africa. Environ. Geochem. Health 2019, doi:10.1007/s10653-019-00493-2.
  124. Kalisa, E.; Nagato, E.G.; Bizuru, E.; Lee, K.C.; Tang, N.; Pointing, S.B.; Hayakawa, K.; Archer, S.D.; Lacap-Bugler, D.C. Characterization and risk assessment of atmospheric PM2.5 and PM10 particulate-bound PAHs and NPAHs in Rwanda, Cen-tral-East Africa. Environ. Sci. Technol. 2018, 52, 12179–12187.
  125. Ofori, S.A.; Cobbina, S.J.; Doke, D.A. The occurrence and levels of polycyclic aromatic hydrocarbons (PAHs) in African en-vironments—A systematic review. Environ. Sci. Pollut. Res. 2020, 27, 32389–32431.
  126. Yan, D.; Wu, S.; Zhou, S.; Tong, G.; Li, F.; Wang, Y.; Li, B. Characteristics, sources and health risk assessment of airborne par-ticulate PAHs in Chinese cities: A review. Environ. Pollut. 2019, 248, 804–814.
  127. Wang, L.; Dong, S.; Liu, M.; Tao, W.; Xiao, B.; Zhang, S.; Zhang, P.; Li, X. Polycyclic aromatic hydrocarbons in atmospheric PM2.5 and PM10 in the semi-arid city of Xi’an, northwest china: Seasonal variations, sources, health risks, and relationships with meteorological factors. Atmos. Res. 2019, 229, 60–73.
  128. Cao, Z.; Wang, M.; Shi, S.; Zhao, Y.; Chen, X.; Li, C.; Li, Y.; Wang, H.; Bao, L.; Cui, X. Size-distribution-based assessment of human inhalation and dermal exposure to airborne parent, oxygenated and chlorinated PAHs during a regional heavy haze episode. Environ. Pollut. 2020, 263, 114661.
  129. Song, Y.; Zhang, Y.; Li, R.; Chen, W.; Chung, C.K.A.; Cai, Z. The cellular effects of PM2.5 collected in Chinese Taiyuan and Guangzhou and their associations with polycyclic aromatic hydrocarbons (PAHs), nitro-PAHs and hydroxy-PAHs. Ecotoxi-col. Environ. Saf. 2020, 191, 110225.
  130. Zhang, Y.; Yang, L.; Zhang, X.; Li, J.; Zhao, T.; Gao, Y.; Jiang, P.; Li, Y.; Chen, X.; Wang, W. Characteristics of PM2.5-bound pahs at an urban site and a suburban site in Jinan in north china plain. Aerosol Air Qual. Res. 2019, 19, 871–884.
  131. Song, W.; Cao, F.; Lin, Y.-C.; Haque, M.M.; Wu, X.; Zhang, Y.; Zhang, C.; Xie, F.; Zhang, Y.-L. Extremely high abundance of polycyclic aromatic hydrocarbons in aerosols from a typical coal-combustion rural site in china: Size distribution, source iden-tification and cancer risk assessment. Atmos. Res. 2019, 248, 105192.
  132. Wang, W.; Ding, X.; Turap, Y.; Tursun, Y.; Abulizi, A.; Wang, X.; Shao, L.; Talifu, D.; An, J.; Zhang, X. Distribution, sources, risks, and vitro DNA oxidative damage of PM2.5-bound atmospheric polycyclic aromatic hydrocarbons in Urumqi, NW Chi-na. Sci. Total Environ. 2020, 139518, doi:10.1016/j.scitotenv.2020.139518.
  133. Yang, J.; Xu, W.; Cheng, H. Seasonal variations and sources of airborne polycyclic aromatic hydrocarbons (PAHs) in Chengdu, China. Atmosphere 2018, 9, 63.
  134. Wu, X.; Cao, F.; Haque, M.; Fan, M.-Y.; Zhang, S.-C.; Zhang, Y.-L. Molecular composition and source apportionment of fine organic aerosols in northeast china. Atmos. Environ. 2020, 239, 117722.
  135. Ma, L.; Li, B.; Liu, Y.; Sun, X.; Fu, D.; Sun, S.; Thapa, S.; Geng, J.; Qi, H.; Zhang, A. Characterization, sources and risk as-sessment of pm2. 5-bound polycyclic aromatic hydrocarbons (PAHs) and nitrated PAHs (NPAHs) in Harbin, a cold city in northern China. J. Clean. Prod. 2020, 264, 121673.
  136. Zhang, L.L.; Yang, L.; Bi, J.R.; Liu, Y.Z.; Toriba, A.; Hayakawa, K.; Nagao, S.; Tang, N. Characteristics and unique sources of polycyclic aromatic hydrocarbons and nitro-polycyclic aromatic hydrocarbons inPM2.5 at a highland background site in northwestern China. Environ. Pollut. 2021, 116527, doi:10.1016/j.envpol.2021.116527.
  137. Xing, W.L.; Zhang, L.L.; Yang, L.; Zhou, Q.Y.; Zhang, X.; Toriba, A.; Hayakawa, K.; Tang, N. Characteristics of PM2.5-bound polycyclic aromatic hydrocarbons and nitro-polycyclic aromatic hydrocarbons at a roadside air pollution monitoring station in Kanazawa, Japan. Int. J. Environ. Res. Public Health 2020, 17, 805.
  138. Ichikawa, Y.; Watanabe, T.; Horimoto, Y.; Ishii, K.; Naito, S. Measurements of 50 non-polar organic compounds including polycyclic aromatic hydrocarbons, n-alkanes and phthalate esters in fine particulate matter (PM2.5) in an industrial area of Chiba prefecture, Japan. Asian J. Atmos. Environ. 2018, 12, 274–288.
  139. Yang, L.; Zhou, Q.Y.; Zhang, H.; Zhang, X.; Xing, W.L.; Wang, Y.; Bai, P.; Yamauchi, M.; Chohji, T.; Zhang, L.L.; et al. At-mospheric behaviour of polycyclic and nitro-polycyclic aromatic hydrocarbons and water-soluble inorganic ions in winter in kirishima, a typical japanese commercial city. Int. J. Environ. Res. Public Health 2021, 18, 688.
  140. Kang, M.; Kim, K.; Choi, N.; Kim, Y.P.; Lee, J.Y. Recent occurrence of pahs and n-alkanes in PM2.5 in Seoul, Korea and char-acteristics of their sources and toxicity. Int. J. Environ. Res. Public Health 2020, 17, 1397.
  141. Kim, I.; Lee, K.; Lee, S.; Kim, S.D. Characteristics and health effects of PM2.5 emissions from various sources in Gwangju, South Korea. Sci. Total Environ. 2019, 696, 133890.
  142. Pham, C.-T.; Boongla, Y.; Nghiem, T.-D.; Le, H.-T.; Tang, N.; Toriba, A.; Hayakawa, K. Emission characteristics of polycyclic aromatic hydrocarbons and nitro-polycyclic aromatic hydrocarbons from open burning of rice straw in the north of Vietnam. Int. J. Environ. Res. Public Health 2019, 16, 2343.
  143. Urbančok, D.; Payne, A.J.; Webster, R.D. Regional transport, source apportionment and health impact of PM10 bound poly-cyclic aromatic hydrocarbons in Singapore’s atmosphere. Environ. Pollut. 2017, 229, 984–993.
  144. Sulong, N.A.; Latif, M.T.; Sahani, M.; Khan, M.F.; Fadzil, M.F.; Tahir, N.M.; Mohamad, N.; Sakai, N.; Fujii, Y.; Othman, M. Distribution, sources and potential health risks of polycyclic aromatic hydrocarbons (PAHs) in PM2.5 collected during differ-ent monsoon seasons and haze episode in Kuala Lumpur. Chemosphere 2019, 219, 1–14.
  145. Thepnuan, D.; Yabueng, N.; Chantara, S.; Prapamontol, T.; Tsai, Y.I. Simultaneous determination of carcinogenic PAHs and levoglucosan bound to PM2.5 for assessment of health risk and pollution sources during a smoke haze period. Chemosphere 2020, 257, 127154.
  146. Javed, W.; Iakovides, M.; Stephanou, E.G.; Wolfson, J.M.; Koutrakis, P.; Guo, B. Concentrations of aliphatic and polycyclic aromatic hydrocarbons in ambient PM2.5 and PM10 particulates in Doha, Qatar. J. Air Waste Manag. 2019, 69, 162–177.
  147. Fadel, M.; Ledoux, F.; Farhat, M.; Kfoury, A.; Courcot, D.; Afif, C. PM2. 5 characterization of primary and secondary organic aerosols in two urban-industrial areas in the East Mediterranean. J. Environ. Sci. 2021, 101, 98–116.
  148. Byambaa, B.; Yang, L.; Matsuki, A.; Nagato, E.G.; Gankhuyag, K.; Chuluunpurev, B.; Banzragch, L.; Chonokhuu, S.; Tang, N.; Hayakawa, K. Sources and characteristics of polycyclic aromatic hydrocarbons in ambient total suspended particles in Ulaanbaatar city, Mongolia. Int. J. Environ. Res. Public Health 2019, 16, 442.
  149. Mehmood, T.; Tianle, Z.; Ahmad, I.; Li, X. Ambient PM2.5 and PM10 bound PAHs in Islamabad, Pakistan: Concentration, source and health risk assessment. Chemosphere 2020, 127187.
  150. Kumar, A.; Ambade, B.; Sankar, T.K.; Sethi, S.S.; Kurwadkar, S. Source identification and health risk assessment of atmos-pheric PM2.5-bound polycyclic aromatic hydrocarbons in Jamshedpur, India. Sustain. Cities Soc. 2020, 52, 101801.
  151. Gadi, R.; Sharma, S.K.; Mandal, T.K. Seasonal variation, source apportionment and source attributed health risk of fine car-bonaceous aerosols over national capital region, India. Chemosphere 2019, 237, 124500.
  152. Roy, R.; Jan, R.; Gunjal, G.; Bhor, R.; Pai, K.; Satsangi, P.G. Particulate matter bound polycyclic aromatic hydrocarbons: Tox-icity and health risk assessment of exposed inhabitants. Atmos. Environ. 2019, 210, 47–57.
  153. Norouzian Baghani, A.; Bahmani, Z.; Sorooshian, A.; Farzadkia, M.; Nabizadeh, R.; Delikhoon, M.; Barkhordari, A.; Rezaei Kalantary, R.; Golbaz, S.; Kermani, M. Characterization of polycyclic aromatic hydrocarbons associated with PM10 emitted from the largest composting facility in the Middle East. Toxin Rev. 2020, 1–15, doi:10.1080/15569543.2020.1737823.
  154. Akhbarizadeh, R.; Dobaradaran, S.; Torkmahalleh, M.A.; Saeedi, R.; Aibaghi, R.; Ghasemi, F.F. Suspended fine particulate matter (PM2.5), microplastics (MPS), and polycyclic aromatic hydrocarbons (PAHs) in air: Their possible relationships and health implications. Environ. Res. 2020, 192, 110339.
  155. Farahani, V.J.; Arhami, M. Contribution of Iraqi and Syrian dust storms on particulate matter concentration during a dust storm episode in receptor cities: Case study of Tehran. Atmos. Environ. 2020, 222, 117163.
  156. Tsiouri, V.; Kakosimos, K.E.; Kumar, P. Concentrations, sources and exposure risks associated with particulate matter in the Middle East area—A review. Air Qual. Atmos. Health 2015, 8, 67–80.
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