Tianjin is a coastal megacity located in northeast part of the North China Plain (NCP), bordering the Bohai Sea in the east, leaning against the Yanshan Mountain in the north (Fig. S1 in the Supplement). It is 120 km from Beijing, the capital of China, in the northwest. The sampling was conducted by utilizing the 255 m meteorological tower at the Tianjin Atmospheric Boundary Layer Observatory of the China Meteorological Administration, located at the southern area of urban Tianjin. Two automatic high-volume aerosol samplers (DIGITEL DHA-80) were set at two floors in the meteorological tower, i.e., the near-ground (2 m) and a higher altitude (220 m), respectively. PM₁₀ samples were collected on pre-combusted (450 °C for 6 h) quartz fiber filter (Φ 150 mm, Pall, USA) during the autumn (18–24 September) and winter (8–14 December) of 2020, and the spring (26–31 May) and summer (2–8 August) of 2021. After collection, the samples were stored at -20 °C in a freezer until analysis.

The extraction and detection methods of 3-OH-FAs are available in our previous study (Niu et al., 2024). Briefly, aliquots of aerosol samples (Φ 28 mm) were cut and placed in a centrifuge tube, followed by adding 4 mL of ethyl acetate (EtOAc), 1 min of vortex, and 15 min of sonicating at room temperature. Then, 4 mL of Milli-Q water was added, and the resulting solution was vortexed for remixing. Centrifugation was performed for 5 min at 5000 rpm, and the supernatant was extracted afterward, and then dried under nitrogen. Subsequently, 100 µL of acetonitrile (ACN) was added to the dried centrifuge tube and vortexed for 2 min, then derivatization reagents (10 µL of 20 µmol mL⁻¹ 2-chloro-1-methylpyridinium iodide (CMPI), 20 µL of 20 µmol mL⁻¹ triethylamine (TEA), and 20 µL of 20 µmol mL⁻¹ 2-dimethylaminoethylamine (DMED)) were added and vortexed for 2 min, and the mixture was incubated for 1 h at 40 °C and 1500 rpm, then dried under nitrogen. Milli-Q water and internal standards (d4-DMED-labeled hydroxy fatty acid standard) (v/v, 1/9) were added to each dried centrifuge tube, vortexed to redissolve, and then centrifuged for 5 min at 4 °C and 12 000 rpm to extract the supernatant. Details of the chemicals, reagents and method validation are described in Sect. S1 in the Supplement.

The UHPLC-MS system comprised an ACQUITY UPLC I-Class LC from Waters (Milford, MA, USA) coupled with an AB SCIEX 6500+ triple quadrupole MS (Framingham, MA, USA). Separation of the target compounds was achieved using ACQUITY HPLC HSS T3 (1.8 µm, 2.1 mm × 50 mm) column (Waters, Milford, MA, USA). The mobile phases consisted of solvent A (0.1 % formic acid in water) and solvent B (ACN), delivered at a flow rate of 0.4 mL min⁻¹. The gradient program was set as follows: 0–3 min at 5 % B, 3–15 min from 22 % to 60 % B, 15–17 min at 60 % B, 17–24 min at 95 % B, 24–40 min at 5 % B. Mass spectrometric analysis was performed in positive ion mode using electrospray ionization (ESI) and operated in multiple reaction monitoring (MRM) mode. The ESI conditions were set as follows: curtain gas, 35 L min⁻¹; collision gas, 8 L min⁻¹; ion spray voltage, 5500 V; temperature, 500 °C; ion source gas 1, 50 L min⁻¹; ion source gas 2, 45 L min⁻¹. Data acquisition and quantification were conducted using SCIEX OS version 2.0.1 software. The mass concentrations of 3-OH-FAs (C₈–C₁₈) were quantified using internal standard method, employing d4-DMED-labeled 3-OH-FA standards as internal standards. The detailed analytical procedure is described in Niu et al. (2024).

The amounts of endotoxins and GNB dry mass were estimated using the following formulas: 1Endotoxin(ngm-3)=∑3-OH-FAsC10–C18(nmolm-3)/4×8000(gmol-1) where 4 means each lipid A carries 4 mol 3-OH-FAs (C₁₀–C₁₈) (Laitinen et al., 2001; Rietschel et al., 1984), while the multiplication factor 8000 represents the average molecular weight of endotoxins (Laitinen et al., 2001). 2GNB dry mass(mgm-3)=∑3-OH-FAsC10–C18(nmolm-3)/15(nmolmg-1) where 15 refers to the biomarker-to-microbial mass conversion factor of 15 nmol of 3-OH-FAs per mg dry cell weight (Lee et al., 2004; Balkwill et al., 1988).

The concentrations of organic carbon (OC) and elemental carbon (EC) were determined using a Thermal/Optical Carbon Analyzer (Model RT-4, Sunset Laboratory Inc., Oregon, USA). Analytical errors were controlled within ±10 % via a duplicate analysis of each filter. The concentrations of water-soluble organic carbon (WSOC) were measured using Total Organic Carbon (TOC) Analyzer (Shimadzu 5000A, Japan). Water-soluble inorganic anions and cations (Cl⁻, SO42-, NO3-, Na⁺, NH4+, K⁺, Mg²⁺, Ca²⁺) were analyzed using an Ion Chromatography system (Dionex Aquion, Thermo Scientific, Waltham, MA, USA). Gradient meteorological parameters were continuously recorded by automatic weather stations installed on the 255 m high meteorological tower, while both gaseous and particulate pollutants were collected simultaneously at the ground platform.

Concentration-weighted trajectories (CWTs) were calculated per hour, and starting height of 220 m, based on the 3 d (72 h) backward trajectories of air masses with the Hybrid Single Particle Lagrangian Integrated Trajectory (HYSPLIT) model (Stein et al., 2015) in conjunction with the measured concentrations of 3-OH-FAs as follows: 3CWTij=∑lClnijl∑lnijl where l is the number of the trajectory; Cl is the average concentration of 3-OH-FAs, and nijl is the number of trajectory endpoints that lie in the grid cell (i,j). The CWT value indicates the potential distant sources impacting the sampling site. The analyses were performed for 1° × 1° grids covering the area encompassed by 20–80° N and 60–150° E. A weight function (Wij) in Eq. (4) was applied in the CWT analysis to increase statistical stability, and Nij is the number of trajectory endpoints that lie in the grid cell (i,j). 4Wij=0.1,Nij<20.4,2≤Nij<40.7,4≤Nij<81.0,elsewhere

Source apportionment of 3-OH-FAs was carried out with U.S. EPA PMF 5.0. PMF solves X≈GF by weighted least-squares minimization of: 5Q=∑i=1n∑j=1mxij-∑k=1pgikfkjuij2 where X is the observed concentration matrix, u the corresponding uncertainty matrix, G the factor-contribution matrix, and F the factor-profile matrix (Paatero and Tapper, 1994; Bhandari et al., 2022). Uncertainties (u) were calculated as recommended by the EPA guidance:

If the concentration is less than or equal to the species-specific method detection limit (MDL) provided, the uncertainty is calculated using a fixed fraction of the MDL (Eq. 6).

If the concentration is greater than the MDL provided, the calculation is based on a user provided fraction of the concentration and MDL (Eq. 7). 6ifx≤MDL,u=56×MDL7ifx>MDL,u=(Error Fraction×Concentration)22+(0.5×MDL)2‾ An error fraction of 0.2 was adopted for most species; values up to 0.6 were assigned to constituents near the detection limit.

A total of 19 measured species in 87 samples were used, including bulk carbonaceous fractions (OC, EC), major inorganic ions (K⁺, Na⁺, Ca²⁺, Mg²⁺, NH4+, NO3-, SO42-, Cl⁻) and nine 3-OH-FA homologues (C₁₀–C₁₈). Species selection followed three criteria: (1) Signal-to-noise ratio (S/N) – species with S/N≤0.5 were excluded; 0.5<S/N≤1 were down-weighted (“weak”) (Paatero and Hopke, 2003). Accordingly, C₁₂ 3-OH-FA was removed. (2) Goodness-of-fit (R2) – after trial runs, variables with persistently low R2 between modelled and observed concentrations were discarded. (3) Q evaluation – models with 4–9 factors (p) were explored; the lowest Q value at five factors when moving from four to nine factors. The solution giving Qtrue/Qexp ratio closest to 1 was selected (Bhandari et al., 2022). The final configuration (five factors) yielded stable and physically interpretable profiles.

Spearman rank analysis was conducted to evaluate the correlation between airborne endotoxins and air pollutants. Data were analyzed using unpaired t-tests, two-tailed for single comparisons. Graphs were generated using RStudio software version 2023.09.1+494, R version 4.3.2 and Origin 2021. Mantel test was performed between 3-OH-FA homologues and those of meteorological/pollution variables using Spearman's correlation coefficient (Spearman's r) with 999 permutations, computed with the vegan package (v2.6-4) in R (v4.3.2). Significance was accepted at p<0.05. Hierarchical-cluster analysis was performed via Ward's minimum-variance method on Euclidean distance using OriginPro. The distribution characteristics of 3-OH-FAs and airborne endotoxins were described in range (minimum–maximum) and arithmetic mean.

The molecular distribution of 3-OH-FAs with carbon numbers from 8 to 18 (C₈–C₁₈) is illustrated in Fig. 1. In terms of diurnal variations, the mass concentrations of most 3-OH-FAs were higher at night than during the daytime (Fig. 1a). This feature, observed predominantly at ground level may be partly due to the lower boundary layer height at night, which contributes to the accumulation of pollutants (Li et al., 2017; Qiu et al., 2019).

The molecular distribution of 3-OH-FAs also showed apparent seasonal variations (Fig. 1b). Even-carbon homologues (C₁₆, C₁₈) were predominant in all seasons, especially in winter, with average mass concentrations of 1355.2 ± 737.6 pg m⁻³ (C₁₆) and 855.4 ± 461.8 pg m⁻³ (C₁₈) at near ground, 882.6 ± 523.2 pg m⁻³ (C₁₆) and 467.7 ± 299.8 pg m⁻³ (C₁₈) at 220 m. C₁₁, C₁₃, and C₁₅ 3-OH-FAs were the predominant odd-carbon 3-OH-FAs, particularly in summer, with average mass concentrations of 367.4 ± 70.5 pg m⁻³ (C₁₁), 412.7 ± 126.6 pg m⁻³ (C₁₃), and 405.8 ± 143.8 pg m⁻³ (C₁₅) at near ground, respectively. In terms of 3-OH-FAs with different carbon numbers, most 3-OH-FAs were more abundant at near ground, suggesting that these homologues were mainly influenced by near-surface emissions. In contrast, C₉ at 220 m remained higher than at 2 m in autumn and summer, regardless of diurnal changes (Fig. 1c, d). Other short-chain homologues (C₈, C₁₀, C₁₁) also presented higher concentrations at 220 m during summer, indicating that they were possibly contributed by regional transport or photochemical oxidation processes (Lei et al., 2021).

The result of concentration-weighted trajectory (CWT) analysis revealed that the mass concentration of 3-OH-FAs at 220 m was influenced by marine sources to varying extents in spring, autumn, and summer (Fig. 2). During autumn and winter, the major source areas were the Beijing-Tianjin-Hebei region, with transport pathways originating from the northwest of Tianjin, including Mongolia, Inner Mongolia and Beijing, before reaching the sampling site. In summer, predominant origin areas included Beijing-Tianjin-Hebei region, Heilongjiang province, and the Bohai Sea in the northeast. Photochemical oxidation of organic matter is prone to occur in summer due to high temperatures and intense solar radiation. The elevated concentrations of short-chain 3-OH-FAs (C₈–C₁₁) at 220 m may result from secondary transformation processes during long-range transport or photochemical oxidation of the local emissions. Bikkina et al. (2019) reported that photochemical oxidation of marine organic matter likely accounts for the predominance of odd-carbon and short-chain 3-OH-FAs in sea-spray aerosols.

Total mass concentrations of endotoxins in PM₁₀ from urban Tianjin during 2020–2021 were estimated based on the measured 3-OH-FAs (C₁₀–C₁₈) (Table 1, Fig. 3a). There was no obvious diurnal difference in total endotoxin levels (Fig. 3a). The endotoxin level was significantly higher (p<0.05) at 2 m than at 220 m during each season (Fig. 4a), indicating that endotoxin emissions were mainly dominated by near-ground sources. It was suggested that the majority of airborne endotoxins may originate from soils (Brooks et al., 2006), as aliphatic 3-OH-FAs (< C₂₀) originate from soil microorganisms (Zelles, 1999; Bikkina et al., 2021a). 3-OH-FAs are also detected in marine ultrafiltered dissolved organic matter (Wakeham et al., 2003), groundwater sediments and estuarine sediments (Parker et al., 1982). Marine source could also be a dominant contributor to elevated endotoxin levels in coastal areas. Lang-Yona et al. (2014) observed high endotoxins content correlated with cyanobacteria at a coastal site of the eastern Mediterranean Sea, assumed that higher wind speeds with low pressure system led to increased sea spray and the consequent primary aerosol emission. Therefore, the endotoxins at 220 m in urban Tianjin might originate from the vertical transport of ground emissions, long-range transport from northwest of Tianjin or from marine emissions (e.g., Bohai Sea) as depicted in CWT results (refer to Sect. 3.1).

In the near-ground (2 m) air, the average endotoxin mass concentrations were the highest (26.5 ± 13.0 ng m⁻³) in winter, followed by 23.7 ± 6.1 ng m⁻³ in summer, 20.5 ± 7.1 ng m⁻³ in fall, and the lowest 14.0 ± 6.6 ng m⁻³ in spring (Table 1, Fig. 4a). While at a higher altitude (220 m), the average endotoxin level was the highest in summer (18.3 ± 5.3 ng m⁻³), followed by winter (16.8 ± 8.7 ng m⁻³) and fall (12.8 ± 2.9 ng m⁻³). The endotoxin concentration ratio between 220 and 2 m during summer was 0.8 ± 0.1, higher than those during autumn (0.6 ± 0.2) and winter (0.7 ± 0.2), suggesting an increased contribution of possible regional transport or photochemical oxidation processes during summer. Most airborne bacteria could not survive during long-range transport due to ultraviolet radiation, which further lead to photodegradation of proteinaceous matter and significant losses of microbial viability (Hu et al., 2021; Yin et al., 2021). LPS is then released into the air during cell death (Petsch and Anspach, 2000), cause the rising in higher endotoxin level. Moreover, the ratio of endotoxin concentrations measured at 220 and 2 m was higher during the daytime (0.8 ± 0.1) than nighttime (0.6 ± 0.1). The average endotoxin mass concentration was slightly higher at night than during the daytime in spring, summer and autumn, but lower at night during winter (Fig. 4a). The increasing concentration of certain bacteria at night may also attribute to the diurnal changes in endotoxins. Hu et al. (2020b) observed richer operational taxonomic units per PM_2.5 sample in Hangzhou at night during spring, with higher relative abundance of Thermomicrobia, a Gram-negative bacterial class, generally distributed away from anthropogenic regions. Abdel Hameed et al. (2009) found the diurnal variations in number concentration of airborne bacteria peaked in the evening at Helwan, Egypt.

The relative contributions of different carbon-numbered 3-OH-FAs to total endotoxin mass concentrations in PM₁₀ of Tianjin presented seasonal and altitudinal variations (Fig. 4b). Overall, the even-carbon long-chain C₁₆ and C₁₈ 3-OH-FAs made up the largest proportion in all seasons, with higher relative contributions (46.5 ± 10.1 %) to total endotoxins at near-ground level than at 220 m (39.7 ± 13.7 %), suggesting their local emissions. The relative contribution of the even-carbon C₁₆ and C₁₈ homologues were the highest in winter (56.0 ± 6.9 % at near ground and 53.3 ± 8.1 % at 220 m), followed by autumn (49.1 ± 5.4 %, 38.6 ± 6.2 %), and the lowest in summer (36.7 ± 7.4 %, 27.0 ± 9.8 %). This finding is consistent with those obtained in suburban Tokyo (Bikkina et al., 2021a), Chichijima island (Bikkina et al., 2021b), and Jeju island (Tyagi et al., 2017), in which 3-OH-FAs were also characterized by a strong even-carbon predominance. Hydroxy fatty acids exhibiting a strong even-carbon preference are usually derived from biological pathways (Bikkina et al., 2019) or indicative of biogenic sources (Tyagi et al., 2015a), such as the lipid fractions of microorganisms (bacteria, algae, fungi, etc.) and vascular plant surface waxes (Kawamura et al., 2003; Simoneit et al., 2004).

However, in this study, the odd-carbon homologues (C₁₁, C₁₃ and C₁₅) accounts for 36.7 % and 41.9 % of total endotoxins at near ground and at 220 m, respectively. Their relative contribution was maximum in summer (45.0 % at near ground and 51.6 % at 220 m), followed by autumn (34.5 % at near ground and 41.4 % at 220 m), and the lowest in winter (30.2 % at near ground and 32.5 % at 220 m). Compared with remote islands dominated by natural emissions such as microorganisms and vegetation (Bikkina et al., 2021b; Tyagi et al., 2015a, 2017), the urban aerosols in Tianjin are largely affected by both anthropogenic emissions and secondary formation (Li et al., 2022). These odd-carbon 3-OH-FAs (C₁₁, C₁₃ and C₁₅) may originate from various sources, such as anthropogenic activities (e.g., fossil fuel combustion, biomass emissions), or from the photochemical oxidation of long-chain fatty acids or hydroxy fatty acids during atmospheric transport (Bikkina et al., 2019).

In addition, long-chain C₁₆, C₁₈ and odd-carbon homologues (C₁₁, C₁₃ and C₁₅) 3-OH-FAs exhibited similar variation trend at the two heights. Long-chain C₁₆, C₁₈ 3-OH-FAs contributed more at 2 m, while odd-carbon homologues (C₁₁, C₁₃ and C₁₅) 3-OH-FAs contributed more at 220 m, regardless of diurnal changes (Fig. 4c).

The airborne endotoxin level at near-ground observed in Tianjin is compared with results in literatures (Fig. 5). Different sampling methods and endotoxin assay methodologies may affect the results of the endotoxin assessment (de Rooij et al., 2017). Herein only the endotoxin level in aerosol samples assessed by chemical analytical method with 3-OH-FAs as biomarkers are compared.

During all seasons, the mean mass concentration of airborne endotoxins in Tianjin was much higher than that in remote islands such as Gosan (except for spring) (Tyagi et al., 2017), Chichijima (Bikkina et al., 2021b; Tyagi et al., 2015a), but close to the endotoxin levels measured in developed areas such as Tokyo (Bikkina et al., 2021a), and the Pearl River Delta in China (Cheng et al., 2012). These results imply that high endotoxin mass concentrations in urban areas may be caused by enhanced human activities (Lee et al., 2004). Guan et al. (2014) discovered a higher endotoxin level in a high-traffic urban setting than in a low-traffic residential area, and similarly Madsen (2006) reported that the endotoxin concentrations on congested streets were higher than in residential areas.

Different from higher endotoxin levels observed in winter and summer in Tianjin, the endotoxin level was much higher in spring at Gosan, Korea, which was probably due to the enhanced mobilization of mineral dust through high altitudinal long-range atmospheric transport (Tyagi et al., 2017). Dutch Expert Committee on Occupational Safety (DECOS), a Committee of the Health Council of the Netherlands stated that exposure to endotoxin above 9 ng m⁻³ for 8 h or more increases the risk of respiratory diseases (DECOS, 2010). The endotoxin concentrations reported in this study and in previous studies exceeded this value, except for remote islands such as Chichijima (Fig. 5), raising an environmental health issue caused by airborne endotoxins.

It should be noted that the endotoxin concentration proposed by DECOS refers to the mass concentration of bioactive endotoxin rather than total mass concentration (DECOS, 2010). In light of the multiple studies, bioactive endotoxin appears to be correlated with C₁₀, C₁₄, C₁₆ 3-OH-FAs in coarse mode particles (PM2.5-10) in the Pearl River Delta region (Cheng et al., 2012); C₁₀, C₁₂, C₁₄ 3-OH-FAs in house dust (Saraf et al., 1997, 1999; Park et al., 2004; Uhlig et al., 2016); C₁₄, C₁₆ 3-OH-FAs in airplane seat dust (Hines et al., 2003); and C₁₂, C₁₄ 3-OH-FAs in agricultural dust (Reynolds et al., 2005). In this regard, C₁₀, C₁₂, C₁₄, C₁₆ 3-OH-FAs were selected to estimate the bioactive endotoxin potential in this study. Correspondingly, the estimated concentrations of bioactive endotoxins were substantially lower compared to total endotoxins (Fig. 5). The level of bioactive endotoxins at 2 m was 6.6 ng m⁻³ in spring, 8.7 ng m⁻³ in autumn, lower than the endotoxin exposure threshold (9.0 ng m⁻³). While the level of bioactive endotoxins exceeded the threshold in summer (9.1 ng m-3) and winter (12.4 ng m⁻³), implying possible health risks of airborne endotoxin exposure.

Based on the empirical equations, the dry mass loading of airborne GNB was estimated from the endotoxin mass concentration. The mass concentrations of airborne GNB measured in this study and previous literatures measured in urban, marine and alpine aerosols are listed in Table 2. The dry mass concentration of airborne GNB showed obvious regional variations. The airborne GNB dry mass concentration in Tianjin was the highest in winter (882.5 ± 167.4 ng m⁻³), followed by summer (787.6 ± 538.0 ng m⁻³), autumn (684.2 ± 464.6 ng m⁻³), and the lowest in spring (467.8 ± 220.4 ng m⁻³) at 2 m. Dry-mass concentrations of GNB in Tianjin PM₁₀ are markedly higher than those reported for PM₁₀ from Hong Kong in 2002–2003 (Lee et al., 2004) and even exceed GNB levels measured in TSP at Chichijima (Japan) (Bikkina et al., 2021b), Tokyo (Japan) (Bikkina et al., 2021a), and Gosan (Korea) (Tyagi et al., 2017). These findings underscore the exceptionally heavy bioaerosol burden in urban Tianjin and highlight the urgency of mitigating biological, as well as chemical, particulate pollution.

Amato et al. (2007) reported that cultivable airborne GNB in cloud water were more abundant in summer than in winter. The phenomenon was further explained by the preferential development of microbial populations on vegetation in summer, coupled with the relative resistance of GNB to ultraviolet (UV) damage. Heavy biomass burning activities may also result in higher GNB mass concentration. GNB mass concentration measured at Mount Tai, China during the high biomass burning events, were 5 times higher than that during the low biomass burning period (Tyagi et al., 2016). During biomass burning events (i.e., wildfires and prescribed fires), microbes can be aerosolized from terrestrial sources into the atmosphere (Bonfantine et al., 2024; Kobziar et al., 2024). It was previously assumed that pyrogenic carbon or smoke produced by biomass burning provides temporary habitat for microbes aerosolized from soils and vegetations (Bonfantine et al., 2024; Kobziar and Thompson, 2020). Meanwhile, particulate matter in smoke confers attenuation of UV-B and UV-A radiation, further leading to increasing bioaerosol viability and higher microbial cell concentration (Mims et al., 1997; Moore et al., 2021). Thus, the high GNB dry mass loading in Tianjin, extremely during summer and winter, could be resulted from the contributions of specific strains of GNB and/or microbial emissions during biomass burning. There are few descriptions of the atmospheric GNB mass concentrations in Tianjin in the literatures, and the distribution of different GNB community, namely the species distribution and their effects on human health remain to be further investigated.

Based on the PMF factor profiles, we identified 5 source factors (microorganisms, secondary nitrate formation process, secondary sulfate formation process, biomass burning, and dust) of PM₁₀ from Tianjin.

Factor 1 was highly loaded on 3-OH-FAs, thus was mainly identified as microbial sources. Factor 2 was characterized by high loading of K⁺ and Cl⁻, which were mainly emitted from biomass burning (Srivastava et al., 2021; Hays et al., 2005). Factor 3 was heavily weighted by NH4+ and NO3-, which was typical of secondary nitrate formation process (Wang et al., 2019a). Factor 4 presented high loaded of two crustal elements Ca²⁺ and Mg²⁺, mainly emitted from dust sources (Huang et al., 2016). Factor 5 indicated secondary sulfate formation process, which was identified by high concentrations of NH4+ and SO42-.

The contributions of different sources to 3-OH-FAs in PM₁₀ were estimated (Fig. 8). Mid-chain 3-OH-FAs (C₁₀, C₁₁, C13) were predominantly contributed by microorganisms (42.6 %–54.4 %), and also associated with secondary sulfate formation process (17.3 %–29.7 %) and dust sources (20.6 %–27.7 %). This result indicates substantial inputs from microorganisms associated with soil dust resuspension and aging of primary aerosols. Long-chain 3-OH-FAs (C₁₄–C₁₈) were likewise primarily originated from microorganisms (41.5 %–57.1 %), with notable contribution from biomass burning (16.5 %–31.4 %) and associated with secondary nitrate formation process (5.2 %–23.2 %). Overall, microorganisms represent the primary biological source of both mid- and long-chain 3-OH-FAs (C₁₀–C₁₈), with additional contributions from secondary processes and biomass burning.

The different behaviors of sulfate and nitrate further highlight source-specific processes. Sulfate-related processes are mainly driven by photochemical oxidation, aqueous phase or cloud chemistry at a regional scale (Slowik et al., 2010; Zheng et al., 2015), whereas nitrate-related processes are typically controlled by gas-particle partitioning, rapid secondary photochemical formation at a local scale (Vispute et al., 2025; Guo et al., 2010; Hu et al., 2016). Furthermore, the higher relative contributions of mid-chain 3-OH-FAs at 220 m (Fig. 4b) support their attribution to regional transport processes, similar to sulfate. Conversely, the enrichment of long-chain homologues at 2 m (Fig. 4b), together with their association with nitrate formation process, suggests a stronger influence from local combustion and secondary formation. In general, the PMF results, consistent with the results of Mantel test, indicate that mid-chain 3-OH-FAs are mainly influenced by natural sources and secondary processes, representing aged, regionally transported aerosols. In contrast, long-chain homologues show stronger links to anthropogenic emissions, particularly biomass burning and local secondary formation, reflecting fresher and more locally derived contributions.