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,
4,
5,
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) [
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,
10,
11,
12,
13]. According to the size, PM is generally divided into inhalable coarse particles (PM with an aerodynamic diameter smaller than 10 μm, PM
10) and inhalable fine particles (PM with an aerodynamic diameter smaller than 2.5 μm, PM
2.5) [
14]. However, other sizes of PM also can be collected in the atmosphere by using different types of air sampler, such as PM
1, PM
2, PM
2.1, and PM
4, which aerodynamic diameters are smaller than 1, 2, 2.1, and 4 μm, respectively [
15,
16,
17,
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,
20,
21,
22,
23,
24,
25,
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 PM
2.5 should not exceed 0.23 mg/min under unvented conditions, and 0.80 mg/min under vented condition [
29]. For outdoor air, the annual average PM
2.5 and PM
10 concentrations cannot exceed 10 and 20 μg/m
3, respectively, and 24-h average concentrations that cannot exceed 25 and 50 μg/m
3, respectively [
30]. Currently, outdoor PM regulation standards also have been implemented in many cities worldwide.
Table 1 lists various PM
2.5 and PM
10 regulation standards set by the governments of several countries according to their national conditions.
Specifically, in North America, the 24-h and annual average PM
2.5 concentration standards in the United States of America (USA) are 35 μg/m
3 and 12 μg/m
3, respectively [
3]. Mexico (12 μg/m
3) and Canada (8.8 μg/m
3) have also defined relatively low standards for the annual average PM
2.5 concentration [
31,
32]. In South America, Brazil determined the final standards of PM
2.5 and PM
10 in 2018, which 24-h and annual average are the same as those prescribed by WHO [
33]. However, Chile had relatively higher standards of PM
2.5 (24-h: 50 μg/m
3; annual: 20 μg/m
3) than these above countries [
34]. In Australia, the 24-h average PM
2.5 and PM
10 concentrations are the same as ones prescribed by the WHO, while the annual average of PM
2.5 (8 μg/m
3) is lower and of PM
10 (25 μg/m
3) is higher than the WHO-prescribed levels [
35]. In the European Union (EU), Russia, and some Asian countries, the PM
2.5 and PM
10 regulation standards are mostly higher than those in the above countries. The annual average PM
2.5 concentration standards are 15 μg/m
3 (Japan), 25 μg/m
3 (EU, Russia, South Korea, and Mongolia), 35 μg/m
3 (China), and 40 μg/m
3 (India), respectively [
36,
37,
38]. In South Africa, only the PM
10 regulation standard has been defined, which does not exceed 40 μg/m
3 for the annual average concentration and 75 μg/m
3 for the 24-h average concentration [
39].
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.
Table 2 lists the information of these 16 PAHs and certain non-priority PAHs, 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,
43,
44,
45,
46,
47]. The vapour pressure of MMW PAHs is between those of LMW and HMW PAHs [
47,
48,
49,
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,
54,
55,
56,
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 (O
3, which is a strong oxidizing agent that can damage human lung function, thus threatening human health [
61]) and NO
x, 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].
Table 3 summarizes the evaluation of PAHs (including heterocyclic PAHs) and their derivatives by the IARC [
27,
64,
65]. 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). 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.
In contrast to PM observations, real-time observations of PAHs have not been commonly analysed. According to Li et al. [
66] real-time observation of PM can offer a high temporal resolution, but low specificity of chemical characterization. Amador-Muñoz et al. [
67] performed real-time PAHs sampling in parallel to non-real-time PM sampling. The non-real time concentrations of PAHs in PM were all lower than the results from real-time PAHs sampling, possibly because PAH derivatives and heterocyclic species were also detected as PAHs when using the real-time method, and the PAHs concentration in the atmosphere is overestimated [
67]. On the other hand, the techniques with high temporal resolution and high specificity for the chemicals in PM are very expensive for routine air quality monitoring [
68]. Therefore, the non-real-time determination method is still a well-recognized method to determine the PAHs, including sampling, experimental treatment, and instrumental analysis. However, although the non-real-time observation method can measure different PAH species as needed, sample processing is a time-consuming task, the information of PAHs such as the atmospheric concentration cannot be promptly provided. To clarify the characteristics of the PAH species in PM, in our previous review, we summarized the size distributions of PM-bound PAHs freshly released from combustion sources and the distribution patterns of PM-bound PAHs in the atmosphere [
69]. It was found that PAHs released from stationary sources are mainly bound to fine particles, but the sizes were slightly larger than that of PAHs released from mobile sources. In the atmosphere, PM-bound PAHs are more likely to be bound to large particles than to initial-mode particles, and the size decreases with increasing PAH MW [
69]. Because the concentrations of PM-bound PAHs may differ according to people, the microenvironment, and area, in this review, we summarize the personal exposure concentrations of PM-bound PAHs and indoor and outdoor PM-bound PAHs in cities worldwide in recent years to further examine the exposure routes to atmospheric PM-bound PAHs and their health effects.
3. Health Effects and Assessments of PAHs
Over the last decades, many studies have been performed to better understand the health effects of PAHs. The health effects of PAH exposure can be divided into acute (short-term) effects and chronic (long-term) effects [
59]. The acute effects mainly depend on the exposure time and PAH concentrations during exposure, while other factors, including pre-existing health conditions and age, may also influence health impacts [
59]. Short-term exposure to high PAH levels may cause impaired lung function in patients with asthma and thrombotic effects in people suffering from coronary heart disease [
157]. Acute occupational exposure to high PAH concentrations could cause eye irritation, nausea, vomiting, and diarrhea [
158]. Repeated skin contacts with certain PAHs, such as anthracene (Ant) and naphthalene (NaP), are also known to cause skin irritation and inflammation [
159]. However, PAHs to induce acute (short-term) human health effects at environmental concentrations are not fully understood.
In terms of chronic effects, long-term exposure to PAHs may induce DNA adduct formation in vitro and in vivo, in which the formation of DNA adducts is a key event regarding the mutagenicity and carcinogenicity of PAHs [
160]. Rota et al. [
161] summarized the studies of the respiratory and urinary tract cancers published between 1958 and 2014, that the high respiratory cancers (mainly lung cancer) were found in the workers who worked in iron and steel foundries for a long time. Petit et al. [
162] investigated 93 exposure groups in nine industries and the results showed the highest lung cancer risk level was found in coke and silicon production, the lowest was in bitumen manufacture. Moreover, long-term exposure to PAHs can increase cardiovascular diseases (CVDs) risks and/or risk factors. Poursafa et al. [
163] reviewed the related reports on exposure to PAHs and CVDs from 2000 to 2017. Most longitudinal with long-term follow-up studies indicated significant positive correlations of exposure to PAHs with CVDs increased risks. On the other hand, some studies reported that exposure to PAHs had negative effects on the development of children. For example, Kalantary et al. [
164] summarized the studies of long-term exposure to PAHs with attention deficit hyperactivity disorder in children until 2018. Although overall studies did not show consistent results, the harm of exposure to PAHs on children is still worth noting and further research.
Regarding the assessment of PAHs, a large PAH database has been structured in various test systems including carcinogenicity, mutagenicity, and genotoxicity in the past decades [
58,
160]. Certain PAHs have been classified as carcinogenic to humans (Groups 1, 2A, and 2B, as indicated in
Table 3) by IARC [
27,
64,
65]. Several LMW PAHs, such as Ant and fluorene (Flu), have not been classified as carcinogenic. The carcinogenicities of some PAHs, such as acenaphthene (Ace), phenanthrene (Phe), and pyrene (Pyr), remain questionable [
160]. On the other hand, because of the toxicity difference between different PAH species, their concentrations are insufficient to indicate the toxicity of PAHs. Hence, it is necessary to choose an appropriate index species for comparison. BaP is one of the most typical carcinogenic PAHs with the largest body of available data describing its exposure and health effects that BaP has commonly been adopted as the index species [
160].
Table 7 summarizes the reference data of certain PAHs including toxicity equivalency factor (TEF) and relative potency factors (RPF) regarding their cancer risks based on BaP [
160,
165,
166,
167,
168,
169,
170,
171,
172,
173]. There is more available data for the 16 PAHs prioritized by the US EPA than for non-priority PAHs, and
Table 7 indicates that among these 16 PAHs, the factor values are relatively low for most LMW PAHs and relatively high for most HMW PAHs. In addition to BaP, dibenz[
a,h]anthracene (DBA) exhibits the highest factor value, which is 5 (TEF) and 10 (RPF) times higher than that reported by Nisbet and LaGoy [
167] and the US EPA [
160], respectively. Regarding the non-priority PAHs, certain PAHs exhibit relatively high factor values, such as benzo[
c]fluorene (BcF), benz[
j]aceanthrylene (BjA), and dibenzo[
a,l]pyrene (DBalP), whose RPF values are 20, 60, and 30 times higher than that of BaP [
160], respectively, suggesting that even if the concentrations of these species in the atmosphere are very low, their health effects on humans are extremely notable.
Currently, it is difficult to accurately assess the health effects of PAHs. The inhalation lifetime cancer risk (ILCR) model has been widely applied to estimate the health risk in people induced by inhalation exposure to PAHs [
72], which is usually calculated according to two methods. One method is the occupational exposure assessment method, and the ILCR is calculated with Equation (1):
where BaP
eq is the BaP-equivalent concentration (ng/m
3), calculated by multiplying the PAH concentration by TEF. UR
BaP is the unit cancer risk resulting from BaP, which is estimated as 8.7 × 10
−5 per ng/m
3, which is based on epidemiological data retrieved from studies on coke oven workers [
58]. The other method is the non-occupational exposure assessment method, and the ILCR is calculated with Equation (2):
where SF is the cancer slope factor (mg·kg
−1·day
−1) for BaP inhalation exposure, IR is the inhalation rate (m
3·day
−1), EF is the exposure frequency (350 days·year
−1), ED is the exposure duration (year), CF is the conversion factor with a value of 10
−6, BW is the body weight (kg), and AT is the average lifespan for carcinogens (25,550 days). The population can be divided into males and females based on gender and subdivided into children, adolescents, adults, and senior adults based on age.