The spatial distribution of PM_2.5 shows a clear trend of "higher in the north and lower in the south, higher inland and lower along the coast"^[15]. In different regions, the concentrations of EPFRs in PMs generally follow a pattern where heavy industrial cities have higher levels than light industrial cities, and sites near highways have higher levels than urban sites. The types of EPFRs in PM_2.5 are primarily carbon-centered radicals containing oxygen atoms, mainly semiquinone radicals with a g factor ranging from 2.0030 to 2.0040^[16-18]. The average concentration of EPFRs in PMs from the Three Gorges Reservoir area in Wanzhou is 7.0×10¹³ spins/m³, with an average g factor ranging from 2.0032 to 2.0038 across different seasons, predominantly consisting of carbon-centered radicals with adjacent oxygen atoms, such as semiquinone and phenoxy radicals^[16]. In Nanjing, EPFRs in PM_2.5 are also mainly carbon-centered radicals with adjacent oxygen atoms, with concentrations ranging from 2.78×10¹² to 1.72×10¹³ spins/m^3[17]. Wanzhou is a smaller city primarily focused on hydraulic engineering development, especially in the Three Gorges Reservoir area, resulting in lower pollution levels. In contrast, Nanjing, as an important industrial base in China, has more complex PM sources and more severe pollution, leading to significantly higher concentrations of EPFRs in PM_2.5^[17]. In Xi'an, the concentration range of EPFRs in PM_2.5 is 1.23×10¹⁸ to 2.56×10¹⁹ spins/g, with a g factor around 2.0023^[18]. Compared to Xi'an, the g factor of EPFRs in PMs from Jinzhong is higher, ranging from 2.00437 to 2.00429. However, unlike most studies that use direct EPR measurement, the study in Jinzhong includes an extraction step, which may lead to significant differences in the g factor^[9]. As a heavily industrialized city in Northeast China with relatively severe air pollution, the concentration of EPFRs in PM_2.5 in Harbin reaches 3.01×10¹⁴ spins/m³ during winter, posing a high potential health exposure risk^[19]. In Beijing, the average concentration of EPFRs in PM_2.5 samples is 2.30×10¹³ spins/m³, with a g factor of 2.00339±0.00025. Combined with the single unstructured signal obtained from EPR measurements, this indicates that the EPFRs in particulate matter are semiquinone radicals centered on oxygen and carbon^[20]. In Zhoukou, the g factor range of EPFRs in road dust from different areas is 2.0032 to 2.0039, mainly consisting of carbon-centered radicals with oxygen atoms. The average concentration of EPFRs in highway road dust is 2.20×10¹⁹ spins/g, while the average concentration in residential street dust is 8.15×10¹⁷ spins/g, indicating that the concentration of EPFRs in highway particulate matter is significantly higher than that in residential areas^[21].

In summary, the gfactor range of EPFRs in PMs is generally between 2.0030 and 2.0040, primarily consisting of carbon-centered free radicals containing oxygen atoms, such as semiquinone radicals. The concentration of EPFRs in particles from traffic sources is usually higher than in other urban areas, but their gfactor is relatively lower. In heavy industrial cities like Harbin, where reliance on fossil fuel combustion is high, PM_2.5pollution is severe, and EPFR concentrations are also relatively high.

PM_2.5EPFR concentrations show a trend of higher levels in winter and lower levels in summer^[12,16,19], with carbon-centered radicals predominating in winter and oxygen-centered radicals dominating in summer^[12,22]. In Xi'an, the g factor for EPFRs in PMs during winter is 2.0033, while in spring and summer it is 2.0034, primarily consisting of radicals with oxygen atoms adjacent to carbon centers. The concentration of EPFRs shows a trend of winter (1.79×10¹⁴ spins/m³) > spring (1.65×10¹⁴ spins/m³) > autumn (1.04×10¹⁴ spins/m³) > summer (9.52×10¹³ spins/m³)^[6]. In Wanzhou District, Chongqing, the concentrations of EPFRs in PM_2.5 across different seasons, from highest to lowest, are: autumn (8.4×10¹³ spins/m³), winter (8.1×10¹³ spins/m³), spring (6.5×10¹³ spins/m³), and summer (4.8×10¹³ spins/m³). The g factor ranges from 2.0031 to 2.0039, mainly comprising carbon-centered radicals adjacent to oxygen atoms, such as semiquinone and phenoxy radicals^[16]. Although the g values of EPFRs in PM_2.5 vary little across seasons, there are differences in the linewidth (∆Hp-p) of the EPR spectra. The ∆Hp-p for EPFRs in summer and autumn is similar, at 4.67 G and 4.63 G respectively; that for winter and spring is also similar, at 4.37 G and 4.49 G respectively. A higher ∆Hp-p indicates a greater variety and complexity of radical species^[16]. The g factor and ∆Hp-p characteristics suggest that the EPFRs bound to PM_2.5 in spring and winter have similar carbon-centered structures, but their chemical compositions may differ significantly^[16]. In Zhengzhou, the concentration range of EPFRs in PM_2.5 is 1.73×10¹² spins/m³ to 7.182×10¹⁴ spins/m³. In autumn (g factor: 2.0031~2.0036) and summer (g factor: 2.0031~2.0035), only semiquinone radicals were detected in PM_2.5, whereas in winter (g factor: 2.0029~2.0035) and spring (g factor: 2.0032~2.0041), carbon-centered radicals, carbon-centered radicals adjacent to oxygen atoms, and oxygen-centered radicals were all detected^[22]. Coal and biomass combustion can produce carbon-centered radicals with adjacent oxygen atoms (g=2.0030~2.0033). As a typical heavy industrial megacity in central China, Zhengzhou relies heavily on coal and biomass combustion for energy supply, especially during winter heating when energy demand increases, leading to a continuous presence of large amounts of carbon-centered radicals with adjacent oxygen atoms in PM_2.5 ^[22-23]. In Beijing, the average concentration of EPFRs in PM_2.5 during spring, autumn, and winter is 6.00×10¹⁷ spins/m³, and during severe winter smog, the maximum concentration of EPFRs in PM_2.5 can reach 2.95×10¹⁸ spins/m³, about four orders of magnitude higher than the concentrations in PM_2.5 from Wanzhou Three Gorges Reservoir Area and Zhengzhou^{[16,20,22,24]}. Additionally, EPFR concentrations in particulate matter during dust storms are generally lower than those during non-dust days. During dust storms in Erlianhot and Jinan, EPFR concentrations were 3.40×10¹³ spins/m³ and 5.50×10¹³ spins/m³, respectively, lower than those in particulate matter from cities like Wanzhou, Xi'an, and Zhengzhou during non-dust days^[12,25]. This may be related to the lower levels of organic precursors and metal ions in dust. Besides sources, meteorological conditions such as precipitation and humidity also affect the concentration of EPFRs in the atmosphere. From July to September in Beijing, abundant rainfall reduces the concentration of EPFRs in the air, resulting in no detectable EPFRs in PM_2.5 during summer^[20]. Xu et al.^[12] found in their study on individual inhalation exposure to EPFRs on PM_2.5 in Beijing that the daily human exposure in summer was 4.79×10¹⁴~7.76×10¹⁶ spins/m³, and in winter heating season it was 1.11×10¹⁷~7.42×10¹⁷ spins/m³. The difference between the lowest concentration in summer and the highest in winter is three orders of magnitude. The g value of EPFRs during the winter heating season is 2.0038 (range: 2.0033~2.0046), indicating that carbon-centered radicals with adjacent heteroatoms predominate; the average g value during the non-heating season is higher, at 2.0041 (range: 2.0032~2.0051), with an increased proportion of oxygen-centered radicals. This may be due to the strong atmospheric oxidation capacity in summer. Atmospheric Oxidation Capacity (AOC) in the troposphere is mainly reflected in

In summary, EPFR concentrations in winter PM_2.5 are higher and dominated by carbon-centered radicals, whereas EPFR concentrations in summer PM_2.5 are lower and primarily consist of more stable oxygen-centered radicals^[12,13]. The primary reasons for this difference are the extensive coal combustion during the winter heating season and the stronger atmospheric oxidation capacity in summer.

Particulate matter with smaller particle sizes contains higher concentrations of EPFRs, predominantly oxygen-centered free radicals^[5,19,26]. Wang et al.^[26]conducted experiments to collect PMs of different particle sizes and found that the g factor of EPFRs in PMs ranged from 2.0031 to 2.0038, with a ∆H p-p range of 8.28 to 9.78 G, mainly consisting of carbon-centered free radicals containing oxygen atoms. A distinct single peak was observed in the EPR spectrum of PM_<2.1, with EPFR concentrations in PM_<2.1ranging from 8.01×10¹³ to 24.49×10¹³ spins/m³. In contrast, the peaks in PM_>2.1showed slight fluctuations, with EPFR concentrations ranging from 1.60×10¹³ to 8.64×10¹³ spins/m³. The highest EPFR content was found in PM_0.4-0.7, reaching 5.1×10¹⁴ spins/m³, approximately 27.0 to 38.9 times higher than the average EPFR concentration (1.6±0.29×10¹³ spins/m³) in PM_5.8-9.0. Jia et al.^[19]obtained PMs of different particle sizes through graded sampling and determined that the g factor in Harbin's PM_0.056-1.8was 2.0034, higher than the g factor of 2.0033 in PM_1.8-10. During the heating season, EPFRs were primarily distributed in PM_0.18-1.8 and PM_1.8-10, with average concentrations of 1.49×10¹⁴ spins/m³ and 8.42×10¹³ spins/m³, respectively. In the non-heating season, EPFRs were mainly distributed in PM_1.8-10 and PM_>10, with average concentrations of 2.14×10¹³ spins/m³ and 1.97×10¹³ spins/m³, respectively. This indicates that smaller particles contain more EPFRs. The composition of PM_2.5-10mainly originates from the Earth's crust, such as dust and soil, while PM_2.5is primarily composed of carbonaceous aerosols generated by anthropogenic combustion, containing abundant EPFRs^[27-28]. Smaller-sized PM_2.5has a larger specific surface area, providing more adsorption sites and oxidative reaction active sites, resulting in a higher distribution of EPFRs on fine particulate matter^[26]. Yang et al.^[5]found that the g factor of EPFRs in TSP extracts ranged from 2.00316 to 2.00371, with concentrations from 3.1×10¹⁹ to 6.2×10²⁰ spins/g, while the g factor of EPFRs in PM_<1μmranged from 2.00323 to 2.00371, with concentrations from 7.4×10¹⁹ to 3.9×10²⁰ spins/g, indicating that more oxygen-containing free radicals with higher g values exist in the extracts of PM_<1μm. Wang et al.^[26]reported that the g factor of EPFRs in PM_<2.1ranged from 2.0031 to 2.0035, mainly consisting of carbon-centered EPFRs with adjacent oxygen atoms; the g factor of PM_>2.1ranged from 2.0025 to 2.0039, higher than that of PM_<2.1, suggesting that the g factor range of EPFRs in PM_>2.1is wider and more complex, potentially influenced by particle surface morphology and metal content. Researchers measured EPFR concentrations in various types of particulate matter from the Xuanwei region of Kunming, where TSP contained significant amounts of transition metals that react with polycyclic aromatic hydrocarbons (PAHs) under strong ultraviolet light to generate stable EPFRs at a concentration of 4.74×10¹⁸ spins/g, with a g factor of 2.0041, identified as oxygen-centered free radicals^[25]. In summary, particulate matter with smaller particle sizes mainly originates from anthropogenic combustion emissions, has a more complex composition, and features more surface adsorption sites and oxidative reaction active sites, leading to more active free radical reactions. Consequently, it contains higher concentrations of EPFRs and is more prone to generating oxygen-centered free radicals.

Since EPFRs were not detected in highly polar or water-soluble extracts, researchers typically use non-polar organic solvents to extract EPFRs from PM_2.5, with dichloromethane (DCM) being the most commonly employed extraction agent^[45]. Yang et al.^[5] mixed 20 mL of DCM with PM_2.5 samples, followed by shaking, light-protected soaking, and centrifugation. The mixture was then evaporated under a gentle nitrogen stream down to approximately 100 μL before being analyzed. To verify the extraction efficiency of EPFRs, researchers performed a secondary extraction on the remaining residue, finding that the EPR signal in the secondary extract was about 2.5% of that from the primary extraction, indicating that a single extraction with DCM was relatively complete^[5]. Using DCM to extract semiquinone radicals from PMs collected in Jinzhong and Beijing, recovery rates of 75.42% to 86.99% were achieved^[9]. However, some studies have reached different conclusions. Truong et al.^[46] extracted EPFRs from 5%CuO/silica particle samples with a particle size of 125~150 μm using 1.5 mL each of methanol (MeOH), isopropanol (IPA), dichloromethane, toluene (TOL), and tert-butylbenzene (TBB), followed by ultrasonic treatment for 1 hour and centrifugal separation. The results showed that the recovery rates of EPFRs using the polar organic solvents IPA and MeOH exceeded 90%, whereas the extraction efficiencies of non-polar solvents (TBB and TOL) were less than 20%. The extraction efficiency of DCM ranged from 20% to 55%, demonstrating significant variations in extraction efficiency among different solvents, with differences reaching up to 70%. Furthermore, these extractable EPFRs can participate in various molecular reactions in solution, generating molecular reaction products such as catechol, hydroquinone, phenol, and chlorophenols. Chen et al.^[45] extracted EPFRs from PM_2.5 samples collected on polytetrafluoroethylene (PTFE) filters, sequentially washing the samples with 10 mL of MeOH, 10 mL of DCM, and 10 mL of n-hexane under pressure to obtain the extracts. The extracts were then concentrated using a rotary evaporator and transferred to quartz filters for EPFR analysis. Comparison with direct measurement results indicated that the g-factor range of EPFRs in the original sample was 2.0028~2.0033, with concentrations ranging from 2.58×10¹⁴ to 5.41×10¹⁴ spins/m³, primarily consisting of carbon-centered organic radicals with heteroatoms. After extraction, the g-factor range of EPFRs was 2.0036~2.0040, predominantly higher-g-factor semiquinone radicals, and the concentration of EPFRs was only 12% of that in the original sample, suggesting that organic solvents struggle to effectively extract EPFRs. Solvent extraction methods involve cumbersome procedures, exhibit large variations in recovery rates among different solvents, and may damage EPFRs during the extraction process, thereby affecting measurement results^[47]. Additionally, the safety of organic solvents is another issue that needs consideration. In contrast, direct measurement methods offer greater advantages.

Currently, direct measurement using EPR is the primary method for determining EPFRs in PM_2.5. Before measurement, researchers first cut the filter membrane into an appropriate size and place it in a quartz tube or quartz plate for analysis^[48]. This method is simple to operate, can fully preserve the sample, and facilitates recovery.

Xu et al^[12]quantitatively analyzed EPFRs in PM_2.5 collected on PTFE filters using a portable personal aerosol exposure monitoring device (microPEM 3.2 A, USA) for 24 hours, employing EPR (Bruker EMXPlus X-band, Billerica, MA, USA). The results indicated that the daily average individual exposure of urban residents in Beijing during the heating period (winter) was 1.11×10¹⁷~7.42×10¹⁷ spins/m³, and during the non-heating period (summer) was 4.79×10¹⁴~7.76×10¹⁶ spins/m³. Qian et al^[16]collected a total of 111 PM_2.5 samples from Wanzhou District, Chongqing, in April, July, October 2017, and February 2018. Custom-made wedges were used to cut the sampled quartz filters into 28mm×5mm strips, which were then placed in quartz tubes for EPR measurement (EPR MS5000 Freiberg Instruments Co., Ltd., Germany). The magnetic field strength was set at 332~341 mT, detection time at 60 s, modulation amplitude at 0.2 mT, and microwave frequency at 8 mW. The results showed that the annual average concentration of EPFRs in PM_2.5 in Wanzhou District was 7.0 ×10¹³ spins/m^{3 [16]}. He et al^[22]collected 28 PM_2.5 samples from Zhengzhou during spring, summer, autumn, and winter seasons in 2019 using quartz fiber filters. Using Bruker EPR EMX plus (X-band) with a dual-cavity model, 1/8 of each sample was placed in a quartz tube for detecting EPFRs on PM_2.5. The results indicated that the concentration range of EPFRs in Zhengzhou PM_2.5 was 1.73×10¹²~7.18×10¹⁴ spins/m³. In addition to using quartz tubes, Chen et al^[48]proposed an electron paramagnetic resonance spectroscopy method based on quartz plates, comparing it with quartz tubes and DCM extraction methods. By replacing the quartz tubes used in EPR measurements with custom-made quartz grooves, placing the filter membrane inside the groove, and fixing it with a custom cover slip before analysis, they found that the EPR spectra obtained using quartz plates were more complete and standardized compared to those from quartz tubes, although the signal intensity was slightly lower; meanwhile, the signals obtained were stronger than those from solvent extraction methods^[48].

It is estimated that by 2050, the number of global premature deaths caused by PMs will reach 5.84 million. Toxicological studies have confirmed that PM_2.5can cause DNA damage^[54-55]. Multiple studies have shown that redox-active transition metals and EPFRs in PM_2.5can synergistically generate ROS. Excessive ROS can induce oxidative stress, and these free radicals can disrupt mucopolysaccharides and lipid membranes in connective tissues within the human body, leading to oxidative damage of lipids and tissues^[7,56-59]. Semiquinone radicals in PMs can also exert damaging effects on human lung epithelial cells and myeloid leukemia cells^[60]. Researchers simulated the preparation of EPFRs by adding iron oxide/SiO₂and Fe³⁺/SiO₂matrices with different concentrations into catechol solutions, followed by stirring, drying, and heating, and conducted in vitro toxicological experiments. The results showed that EPFRs can increase endogenous hydroxyl radicals (⋅OH) in human alveolar cells, causing DNA damage, promoting tumorigenesis, and triggering inflammatory responses^[8,20,22]. A study by Hwang et al.^[28]indicated that the endogenous ⋅OH induced by highway PM_2.5in aqueous phase showed a strong correlation with EPFRs (r ²=0.73, p<0.05), further demonstrating that EPFRs can induce the generation of ⋅OH.

Currently, there is still a lack of standardized models for assessing the exposure risk of EPFRs, resulting in limited related research. The most commonly used risk assessment model available is the cigarette exposure model, which converts the exposure level of EPFRs in PM_2.5 into the number of cigarettes^[43,61-63], as shown in Equation (1):

in the formula: C _EPFRsrepresents the average concentration of EPFRs in PM_2.5(spins/m³), Vis the daily air volume inhaled by an adult male (20 m³/day), RC_cigis the concentration of free radicals in cigarette tar, approximately 4.75×10¹⁶ spins/g^[64], and C _tarindicates the tar content per cigarette, approximately 0.013 g/cig^[63]. The 24-hour average PM_2.5concentration in Louisiana, USA, is 26.9 μg/m³, with an average EPFRs concentration of 3.84×10¹⁷ spins/g. The amount of EPFRs inhaled per person per day is equivalent to 0.3 cigarettes. Based on the annual average PM_2.5concentration in the USA of 10.6 μg/m³, the equivalent number of cigarettes inhaled per person annually due to EPFRs is 47^[43]. In Wanzhou, the daily average EPFRs concentration is 1.73×10¹² spins/m³, and the average daily EPFRs inhalation per person is equivalent to 2.3 cigarettes, totaling 840 cigarettes annually^[16].

Additionally, the concentration used for calculating the equivalent number of cigarettes for EPFRs can also be expressed as EPFRs per gram of PMs. Xu et al.^[20]used another more detailed assessment formula in their study, combining body weight and lung deposition fraction. They first estimated the daily exposure to EPFRs in PM_2.5 using Equation (2), and then calculated it based on the concentration of EPFRs in cigarette tar, as shown in Equation (3), to assess the equivalent number of cigarettes inhaled for health risk evaluation.

in equation (2): Inh_PM represents the daily exposure to EPFRs from inhaled PM_2.5 (spins/kg/day); RC_PM denotes the average concentration of EPFRs in unit mass of PM_2.5 (spins/g); F is the conversion factor from g to μg (1×10⁶); PC_PM is the concentration of PM_2.5 (μg/m³); IR is the adult inhalation rate (20 m³/day); F _r is the fraction of alveolar particles retained in the lungs (0.75); body weight is 70 kg for adults; in equation (3), EQ represents the equivalent number of cigarettes smoked per day by a 70-kg adult due to EPFR exposure from inhaled PM_2.5; RC_cig is the concentration of EPFRs in cigarette tar (9×10⁶ spins/g), and C _tar denotes the tar content per cigarette, approximately 0.013 g/cigarette^[63]. Xu et al. found that during winter, the daily exposure to EPFRs from PM_2.5 for adults ranged from 1.47×10¹⁸ to 6.33×10²⁰ spins/g, with an average equivalent of smoking between 0.53 cigarettes per day on non-haze days in spring, autumn, and winter, up to 226.9 cigarettes per day on haze days, resulting in an average daily exposure equivalent to 33.1 cigarettes' worth of EPFRs from tar^[20]. In Shihezi City, the concentration range of EPFRs in PM_2.5 was 1.35×10¹³ to 4.65×10¹³ spins/g, with a daily inhalation exposure equivalent to 0.66 to 8.40 cigarettes' worth of EPFRs from tar. Specifically, the number of cigarettes inhaled varied by season: 1.69 to 4.61 cigarettes per day in spring, 0.66 to 4.47 cigarettes per day in summer, 1.49 to 4.37 cigarettes per day in autumn, and 2.26 to 8.40 cigarettes per day in winter. The winter exposure levels were significantly lower than the 226.9 cigarettes observed during severe haze events in Beijing^[20,65].

Although the cigarette exposure model can assess the exposure levels of EPFRs, it remains imperfect. The material composition and chemical components of PMs differ from those of cigarette particles; therefore, the types of EPFRs in cigarettes differ from those in PMs, resulting in varying exposure risks. Moreover, this model only focuses on exposure levels, lacking further health risk assessments and standardized health risk reference values.