Measurements were conducted at an urban site in the National Center for Nanoscience (39.99∘ N, 116.32∘ E) in Beijing, which is close to the fourth ring of Beijing and surrounded by residential, commercial and traffic areas. All instruments were deployed on the roof of a five-story building (∼20 m above the ground) and the measurements were performed from 15 August to 4 December 2015.

NR-PM1 species including organics, sulfate, nitrate, ammonium and chloride were continuously measured by an Aerodyne quadrupole ACSM (Q-ACSM) with a time resolution of ∼30 min. Detailed descriptions of ACSM operation can be found elsewhere (Ng et al., 2011a; Wang et al., 2017). Briefly, the ambient aerosol was sampled at a flow rate of ∼3 L min-1 through a 3/8 in. stainless steel tube and a University Research Glass ware (URG) cyclone (Model: URG-2000-30ED); a size cut of 2.5 µm in front of the sampling inlet was used to remove coarse particles. A Nafion dryer (MD-110-48S; Perma Pure, Inc., Lakewood, NJ, USA) was applied to dry aerosol particles before they entered the ACSM, and the submicron aerosol was subsampled into the ACSM with a flow rate of 85 cc min-1 fixed by a 100 µm diameter critical aperture. The submicron particles were focused into a narrow beam by an aerodynamic lens and impacted a hot vaporizer (∼600∘). The resulting vapor was ionized with electron impact and chemically characterized with a quadrupole mass spectrometer. Monodispersed 300 nm ammonium nitrate particles, generated by an atomizer (Model 9302, TSI Inc., Shoreview, MN, USA) and selected by a differential mobility analyzer (DMA; TSI model 3080), were used to determine the response factor (RF) and calibrate the ionization efficiency (IE) (Ng et al., 2011a).

An aethalometer (Model AE-33, Magee Scientific) was used for the determination of black carbon (BC) concentration with a time resolution of 1 min. SO2 was measured by an Ecotech EC 9850 sulfur dioxide analyzer, CO by a Thermo Scientific Model 48i carbon monoxide analyzer, NOx by a Thermo Scientific Model 42i NO-NO2-NOx analyzer and O3 by a Thermo Scientific Model 49i ozone analyzer. Meteorological parameters, including wind speed, wind direction, relative humidity (RH) and temperature, were measured by an automatic weather station (MAWS201, Vaisala, Vantaa, Finland) and a wind sensor (Vaisala Model QMW101-M2).

Aerosol liquid water content (ALWC) was predicted using the ISORROPIA-II model (Fountoukis and Nenes, 2007) with ACSM aerosol composition and meteorological parameters (temperature and relative humidity) as input. The ISORROPIA-II model then calculated the composition and phase state of a NH4+–SO42-–NO3-–Cl-–H2O system in thermodynamic equilibrium, and the concentration of H+ and ALWC could be resolved.

Figure 1 shows the time series of meteorological parameters, trace gases and PM1 composition during the entire measurement period. The relatively clean events and polluted episodes occurred alternatively during the entire campaign. As shown in Fig. 1, the variations in PM1 species are strongly associated with meteorological conditions. For example, clean periods were generally associated with northerly and northwesterly winds with high wind speeds. However, serious pollution episodes were related to southerly winds with low wind speeds (< 1 m s-1), indicating the important role of stagnant meteorological conditions in haze pollution (Takegawa et al., 2009; Huang et al., 2010; Sun et al., 2014). The mass concentration of PM1 varied from 0.4 to 260.7 µg m-3. Considering that the long-term measurements in our study have different meteorological conditions, we separated the entire study into three periods as late summer (15 August to 10 September), autumn (11 September to 10 November) and early winter (11 November to 4 December) in order to discuss the seasonal variations in PM1 mass concentration and chemical composition.

The average mass concentration of PM1 was 21.6 µg m-3 in late summer (Fig. S3), which was much lower than that measured in July–August 2011 (50.0 µg m-3; Sun et al., 2012) and in August–September 2011 (84.0 µg m-3; Hu et al., 2016) (see Table 1). This lower PM1 concentration was likely associated with the 2015 China Victory Day parade control from 23 August to 3 September, which significantly improved air quality in Beijing (Zhao et al., 2017). OA constituted a major fraction of PM1 mass (64 %), followed by sulfate (14 %), BC (8 %), ammonium (7 %), nitrate (6 %) and chloride (1 %). During autumn, the mean concentration of PM1 increased to 43.3 µg m-3, which was 2 times higher than that in late summer. OA contributed a mass fraction of 49 %, followed by nitrate, sulfate, ammonium, BC and chloride with the mass fractions of 22 %, 11 %, 8 %, 8 % and 2 %, respectively. Compared to late summer, the mass fraction of OA decreased to 49 % (but the OA mass increased from 13.8 to 21.2 µg m-3), and the mass fraction of inorganic species increased correspondingly. The increase in inorganics was particularly noticeable for nitrate, which increased from 6 % to 22 % (or from 1.3 to 9.5 µg m-3). The mean concentration of PM1 was 64.3 µg m-3 in early winter, even higher than those in late summer and autumn. This PM1 average concentration in wintertime Beijing is similar to other studies, such as Hu et al. (2016) (60.0 µg m-3), Sun et al. (2013) (66.8 µg m-3) and Sun et al. (2016) (64.0 µg m-3). OA accounted for 46 % of PM1 mass in early winter, followed by 20 % of nitrate, 15 % of sulfate, 10 % of ammonium, 6 % of BC and 3 % of chloride (Fig S3).

As shown in Figs. 1f and S3, OA dominated PM1 mass in late summer. In autumn and early winter, however, the contribution of OA decreased and secondary inorganic aerosol increased to be equally important. It should also be noted that nitrate had a more important contribution than sulfate to PM1 during autumn and early winter, with nitrate/sulfate mass ratios of 2.0 and 1.3 in autumn and early winter, respectively. This phenomenon is likely due to the efficient emission reduction of SO2 and the continuous increase in NOx because of dramatic growth of the vehicle fleets and large emissions from industries (Xu et al., 2015). Therefore, nitrate is expected to play a more important role in PM pollution in the near future and controlling NOx emission would greatly help mitigate air pollution in Beijing.

The diurnal cycles of PM1 species during different seasons are shown in Fig. S4. OA was characterized by three peaks occurring in the morning (06:00–09:00 LT), at noon (12:00–14:00 LT) and in the evening (19:00–22:00 LT) during all three seasons. Such diurnal patterns were partially influenced by the emission behavior of pollution sources, i.e., traffic, cooking and/or coal burning emissions (Huang et al., 2012; Sun et al., 2012; Crippa et al., 2013). Due to the relatively flat planetary boundary layer (PBL) height related to stagnant meteorological conditions in early winter compared to that in autumn and late summer, the noon peak of OA was more evident in early winter. The morning peak of OA was even more pronounced than the noon peak in late summer. Such a diurnal cycle was likely related to the efficient photochemical oxidation in the morning and efficient dilution effect resulted from PBL height increase at noon.

The diurnal cycle of nitrate varied significantly during different seasons due to the seasonal difference in photochemical production and gas-particle partitioning (Sun et al., 2015). Compared to nitrate, sulfate showed a relatively flat diurnal cycle in all seasons. A clear increase in sulfate in the afternoon was observed during late summer and autumn due to enhanced photochemical processes (Takegawa et al., 2009). In the winter, however, sulfate showed a decreasing trend in the afternoon, suggesting low photochemical production as discussed below. Chloride presented a morning peak and then rapidly decreased to a low concentration level at ∼ 18:00 LT during late summer, while in both autumn and winter, chloride displayed a diurnal cycle with higher concentrations at nighttime, which may be related to the local emission from coal combustion. BC also showed a similar diurnal cycle with higher concentrations at nighttime and lower concentrations in the daytime during all three seasons.

Three POA factors were resolved in this study: HOA, COA and CCOA. As shown in Fig. 2a, the HOA mass spectrum is characterized by prominent hydrocarbon ion series of CnH2n-1 and CnH2n+1, particularly m/z 27, 29, 41, 43, 55, 57, 67 and 71. The HOA spectrum is similar to previously reported HOA spectra at various urban sites (He et al., 2011; Ng et al., 2011; Sun et al., 2012). The time series of HOA is also correlated well with that of BC, which is an external tracer of incomplete combustion (R2=0.56). The mass fractions of HOA (10 %–13 %) and diurnal cycles in different seasons are rather consistent. There are two peaks from rush hours, i.e., 07:00–09:00 LT in the morning and around 20:00 LT in the evening. The nighttime concentrations are generally high (Fig. S4), likely due to increased diesel fleets, which are allowed in urban Beijing only at night, and the decrease in PBL during nighttime.

The COA profile is characterized by prominent ion peaks at m/z 55 and m/z 57 (Fig. 2b) and a higher ratio of intensity at m/z 55 over that at m/z 57 (=2.3) compared to the other two primary OA components (∼1), which have been shown to be robust markers for COA (He et al., 2010; Mohr et al., 2012; Crippa et al., 2013; Elser et al., 2016). This COA mass spectrum is highly correlated with other COA profiles reported in previous studies (Crippa et al., 2013; Elser et al., 2016; Wang et al., 2017), and the time series correlated well with that of m/z 55 with R2=0.81. The COA diurnal cycle showed two obvious peaks at lunch (12:00 LT) and dinner (20:00 LT) time and a smaller peak at breakfast time (07:00 LT) (Fig. S4). Similar diurnal behaviors of COA have been observed in Beijing and at other urban sites (Allan et al., 2010; Sun et al., 2010, 2013). COA had a lower mass fraction of 11 % during late summer compared to autumn (20 %) and early winter (16 %).

The mass spectrum of CCOA is dominated by unsaturated hydrocarbons, particularly PAH-related ion peaks (e.g., 77, 91 and 115) (Dall'Osto et al., 2013; Hu et al., 2013). It shows a similar spectral pattern with the ambient CCOA mass spectra in Beijing and Xi'an (Elser et al., 2016). The presence of CCOA can be further validated by the good correlation with external combustion tracer chloride (R2=0.77) (Zhang et al., 2012). The time series of CCOA shows that the mass concentration of CCOA was much lower in August and September but increased dramatically after November, indicating the large emissions from residential coal combustion for domestic heating. Also, the nighttime CCOA concentrations were much higher than the daytime concentrations, further confirming the enhanced coal combustion emissions from domestic heating in wintertime nights. Specifically, on average, the mass fraction of CCOA increased from 5 % (0.7 µg m-3) in late summer to 9 % (2.0 µg m-3) in autumn and then to 26 % (7.7 µg m-3) in early winter (Fig. S3).

In order to analyze sources of sulfate in our study period, the bivariate polar plots of sulfate during different seasons are displayed in Fig. 3. During late summer, the high mass concentration of sulfate was mainly located in the south and southwest regions from the sampling site, suggesting regional transport was the major source of sulfate in late summer. However, high sulfate was located both at the sampling site and in the south and southeast regions from the sampling site in autumn, which indicates that both local formation and regional transport contributed to the sulfate concentration. When it comes to the early winter, high mass concentration of sulfate was mainly located at the sampling site coming from local formation and there was almost no contribution from regional transport. These results indicate that transported sulfate at a large regional scale was more important during late summer, while local formation was the major source of sulfate in early winter due to residential heating.

Two oxygenated OA factors with very different time series were identified in our study, which we defined as local SOA (LSOA) and regional SOA (RSOA) as characterized below in detail. As shown in Fig. 3, different correlations between sulfate and RSOA or LSOA were found during different seasons. The time series of RSOA correlated well with that of sulfate during late summer with R2=0.71. This correlation coefficient decreased to 0.62 during autumn, and there was almost no correlation between RSOA and sulfate (R2=0.02) in early winter. By contrast, the correlations between LSOA and sulfate displayed the opposite variation with the correlation coefficient (R2) increasing from 0.40 in late summer to 0.66 in autumn and 0.86 in early winter (Fig. 3a). As we have discussed that sulfate mainly came from regional transport during late summer, while the contribution of local formation increased during autumn and further became the dominant source of sulfate, these correlation variations (i.e., better correlation with RSOA in late summer, similar correlations with RSOA and LSOA in autumn and close correlation with LSOA in early winter) suggested that RSOA is related to the regional source of OOA and LSOA indicates a local source and subsequent local formation. These two SOA factors show similar mass spectra with high ratios of intensity at m/z 44 over that at m/z 43 (f44/43), and the f44/43 of RSOA (4.8) is higher than that of LSOA (2.9), suggesting that RSOA from regional transport is more oxygenated (more aged) than locally formed SOA (Sun et al., 2014, 2015). The attribution of LSOA and RSOA is further supported by the bivariate polar plots (Fig. S5), which show clearly that LSOA is mainly located at the sampling site while RSOA is mainly from the south of the sampling site. The average mass concentration of LSOA increased from 3.2 in late summer to 9.2 µg m-3 in autumn and to 12.1 µg m-3 in early winter with an increase in mass fraction from 23 in late summer to 43 % in autumn and 41 % in early winter. By contrast, the average mass concentration of RSOA decreased from 6.6 in late summer to 3.8 µg m-3 in autumn and to 1.8 µg m-3 in early winter, with a dramatic decrease in mass fraction from 48 % in late summer to 18 % in autumn and to 6 % in early winter (Fig. S3). These seasonal variations in LSOA and RSOA indicate that RSOA related to regional transport was more important during late summer, while locally formed LSOA played a dominant role in autumn and early winter.

The average PM1 concentration increased from late summer (21.6 µg m-3) to early winter (64.3 µg m-3) (Fig. S3) and the chemical composition showed a seasonal difference. The mass concentrations of secondary species increased from 15.7 µg m-3 in late summer to 30.8 µg m-3 in autumn and to 42.8 µg m-3 in early winter, but the mass fraction in PM1 decreased from 72 % in late summer to 66 % in early winter. In particular, SOA had a dominant contribution in late summer (9.8 µg m-3, 46 % of PM1), while secondary inorganic aerosol (SIA) played a key role during autumn (17.8 µg m-3, 41 % of PM1) and early winter (28.9 µg m-3, 45 % of PM1) (Fig. S3). The high SOA fraction in summer is likely associated with active photochemical oxidation, while the increased SIA fraction in autumn and early winter is likely due to enhanced gas-particle partitioning of nitrate and aqueous-phase formation of sulfate.

Figure 4 shows the PM1 composition and OA sources on clean days (daily average PM1 < 20 µg m-3), medium-pollution days (M-pollution; 40 µg m-3 < daily average PM1 < 80 µg m-3) and high-pollution days (H-pollution; daily average PM1 > 80 µg m-3) during different seasons. The mass concentrations of PM1 species and OA factors, gaseous pollutants, and meteorological parameters during different periods are summarized in Table S1 in the Supplement. The average concentration of PM1 was 46.9 µg m-3 during M-pollution days, about 3 times higher than that during clean days (15.6 µg m-3) in late summer. In autumn and early winter, the average PM1 concentrations during H-pollution days (110.5 and 109.7 µg m-3, respectively) were 2 times higher than those on M-pollution days (54.2 and 43.5 µg m-3, respectively) and 10 times higher than those on clean days (9.3 and 8.1 µg m-3, respectively). As shown in Fig. 4, the mass fraction of secondary aerosol species (SIA and SOA) increased from clean days (52 %–70 %) to M-pollution days (67 %–76 %) and H-pollution days (66 %–74 %) during all three seasons, emphasizing the significant enhancements of secondary aerosol formation in haze pollution events (Huang et al., 2014; Jiang et al., 2015; Zheng et al., 2015). In late summer, the mass concentration of LSOA increased from 2.2 µg m-3 (21 % of OA) during clean days to 6.7 µg m-3 (24 % of OA) during M-pollution days, and the mass concentration of RSOA increased from 5.0 µg m-3 (48 % of OA) during clean days to 13.8 µg m-3 (49 % of OA) during M-pollution days, suggesting that regional transport played a more important role than local formation in both clean and haze pollution events during late summer. The mass concentration of LSOA increased from 1.5 on clean days to 10.2 µg m-3 on M-pollution days and to 25.4 µg m-3 on H-pollution days during autumn and increased from 1.5 µg m-3 on clean days to 7.5 µg m-3 on M-pollution days and to 20.7 µg m-3 on H-pollution days during early winter. In comparison, the mass concentration of RSOA increased from 1.5 and 0.6 µg m-3 on clean days to 5.9 and 2.0 µg m-3 on M-pollution days and to 6.6 and 2.5 µg m-3 on H-pollution days during autumn and early winter, respectively. The increased rates of LSOA were much higher than that of RSOA; thus, the mass fraction of LSOA increased dramatically from clean days to M-pollution and H-pollution days in autumn and early winter (i.e., 26 % to 40 % and 50 % during autumn and 33 % to 37 % and 42 % during early winter), whereas the mass fraction of RSOA decreased from clean days to M-pollution and H-pollution days (i.e., 25 % to 23 % and 13 % during autumn and 14 % to 10 % and 5 % during early winter). These observations suggest that locally formed SOA had more important contributions than regional sources in haze pollution during autumn and early winter, implying a different contribution of secondary aerosol in different seasons.

The clean and pollution episodes occurred in “sawtooth cycles”, in which meteorological conditions, regional transport, local emissions and secondary formation intertwine and play different roles in the evolution of PM pollution. To get a better insight into aerosol sources and atmospheric processes, seven clean episodes (average PM1 concentration < 20 µg m-3), seven M-pollution episodes (40 µg m-3 < average PM1 concentration < 80 µg m-3) and five H-pollution episodes (average PM1 concentration > 100 µg m-3) were selected for further analysis. As shown in Fig. 5, the pollution episodes were generally associated with higher RH and lower wind speeds (< 1 m s-1) than in clean episodes in autumn and early winter, with RH usually higher than 60 % in pollution episodes (both M-pollution and H-pollution) and lower than 45 % in clean episodes. Specifically, an M-pollution (M1; 47.6 µg m-3) episode in late summer had similar RH and wind speed to the adjacent clean period (C1; 14.1 µg m-3). However, the contribution of organic species decreased from 68 % in C1 to 61 % in M1, but the mass fraction of secondary inorganic species (particularly sulfate) increased from 23 % in C1 to 33 % in M1. This phenomenon may result from enhanced photochemical formation of secondary species in M1 due to higher oxidation capacity as M1 had higher O3 concentration (54.1 ppb) than C1 (31.0 ppb). In autumn, the mass concentrations of organics increased from 4.8 to 6.3 during C2–C5 to 21.2–27.8 µg m-3 during M2–M6, while the contributions decreased from 56 %–71 % to 39 %–55 %, and the corresponding contributions of secondary inorganic species increased from 17 %–29 % during C2–C5 to 36 %–52 % during M2–M6 with mass concentrations increasing from 1.6–2.9 to 16.7–33.1 µg m-3. The contributions of secondary organic species to OA also increased from 50 %–61 % to 55 %–73 % with mass concentrations increasing from 2.7–3.6 to 14.1–19.4 µg m-3. This indicates a notable production and accumulation of secondary aerosol during pollution events. Compared to M-pollution episodes, there was no further increase in the contribution of secondary inorganic species during H1–H3 (42 %–47 %) although the mass concentrations increased to 45.3–56.6 µg m-3 due to the systematic concentration growths of all species from M-pollution to H-pollution. Secondary organic species also had similar contributions to OA during H1–H3 (52 %–75 %) to that during M2–M6 (55 %–73 %) although the mass concentrations increased from 14.1–19.4 to 25.6–38.5 µg m-3. Further analysis shows that the RH during H1–H3 (71.7 %–81.6 %) is lower than that during M2–M6 (74.1 %–91.8 %), which indicates that the stronger aqueous-phase chemistry during M2–M6 may lead to the efficient formation of secondary species, and the mass concentration growths of secondary species were faster than that of other species in PM1; thus, the mass fraction of secondary species in M2–M6 was higher or similar to that in H1–H3. A similar phenomenon was also found in early winter. The contributions of secondary species increased from clean episodes (C6 and C7) to pollution episodes (M7, H4 and H5), while the contributions of secondary species were similar in M7, H4 and H5 because of similar RH. These PM evolution characteristics observed here highlight the importance of meteorological conditions for driving particulate pollution (Li et al., 2017) and imply different formation mechanisms of PM pollution during different seasons.

To further elucidate the formation mechanisms of secondary aerosol, the sulfur oxidation ratio (FSO4) (Sun et al., 2006) was calculated according to Eq. (1): 1FSO4=n[SO4]nSO4+n[SO2], where n[SO4] and n[SO2] are the molar concentrations of sulfate and SO2, respectively. Figure 6a–c plots FSO4 versus Ox (= O3 + NO2) concentration which is a tracer to indicate photochemical processing during late summer, autumn and early winter, respectively. During late summer, positive correlations between FSO4 and Ox with similar slopes and correlation coefficients in RH < 65 % and RH > 65 % were observed, suggesting the important role of photochemical oxidation during late summer irrespective of the RH range. During autumn and early winter, at RH < 65 % sulfate was also formed by photochemical oxidation because of the positive correlations between FSO4 and Ox, while there was no correlation between FSO4 and Ox at RH > 65 %, indicating that other processes (e.g., aqueous-phase reactions) may contribute to the sulfate formation in high-RH conditions. This is supported by the relationships between FSO4 and ALWC at RH > 65 % and low atmospheric oxidative capacities of Ox < 60 ppb (Fig. 6d–f). There were positive correlations between FSO4 and ALWC during all three seasons in high-RH conditions, indicating the contribution of aqueous-phase processing to the sulfate formation. Meanwhile, we found that FSO4 was up to ∼0.6 with Ox while it was only up to ∼0.3 with ALWC during late summer, suggesting the more important role of photochemical oxidation for the sulfate formation during late summer. By contrast, during early winter the increase in FSO4 with ALWC (from ∼0.05 to ∼0.5) was more efficient than that with Ox (from ∼0.05 to ∼0.2), indicating that aqueous-phase reactions were more responsible during early winter. During autumn, FSO4 was up to about 0.4–0.5 both with Ox and ALWC, suggesting that for sulfate formation during autumn both photochemical oxidation and aqueous-phase reaction had important contributions. It should be noted that at the typical atmospheric level of OH radicals, the lifetime of SO2 from the reaction with OH is about 1 week (Seinfeld and Pandis, 2016; Zhang et al., 2015), and the bivariate polar plots of Ox in late summer also showed a regional source (Fig. S6). Thus, SO2 oxidation into sulfate may proceed during long-range transport in late summer (Rodhe et al., 1981), consistent with our results in Fig. 3.

We further investigated the formation mechanisms of SOA during different seasons. Figure 7 shows the effects of ALWC and Ox on the mass concentrations and mass fractions of LSOA and RSOA during different seasons. During late summer, the ALWC ranged from 2.1 to 53.6 µg m-3; both the mass concentrations of LSOA and RSOA increased as ALWC increased when ALWC was higher than ∼25–35 µg m-3. In comparison, the ALWC concentrations during autumn and early winter were much higher than that during late summer, and the increasing trends of SOA were much obvious than that during late summer. The mass concentrations of LSOA and RSOA increased from 7.3 to 33.3 and 3.5 to 11.5 µg m-3, respectively, when ALWC increased from 12.3 to 519.6 µg m-3, and the mass fraction of SOA increased from 30 % to 38 % during autumn. In comparison, during winter, the mass concentration of LSOA increased from 5.6 to 37.9 µg m-3 when ALWC increased from 9.7 to 436.6 µg m-3 with the mass fraction of LSOA increasing from 37 % to 42 %. RSOA displayed no clear increase trend with ALWC as it played a minor contribution during early winter. These variations indicated the influence of aqueous-phase processes on the formation of SOA especially during autumn and early winter with higher ALWC. Variations in the mass concentrations and fractions of LSOA and RSOA as functions of Ox during different seasons are also shown in Fig. 7. The mass concentrations of SOA increased clearly with the increase in Ox concentration during all three seasons, and the mass fraction of SOA also increased from 64 % to 76 % during late summer and increased from 59 % to 80 % during autumn as Ox increased from 30 to 120 ppb. Similar to that of ALWC, the increasing rates of LSOA and RSOA as functions of Ox were substantially different among different seasons. In late summer, both LSOA and RSOA presented linear increases with the increase in Ox. As a comparison, LSOA showed higher increase rates with Ox than that of RSOA during autumn and early winter as LSOA played a dominant role in the haze formation during autumn and early winter. These results clearly indicate that both photochemical processing and aqueous-phase reactions played important roles in the formation of SOA during all three seasons.