Field campaigns were performed in Wangdu, Beijing, and Mt. Tai in sequence during the winter–early spring of 2017–2018 (Table 1). The locations of the three sites are shown in Supplement Fig. S1. The sites were selected for investigation of ClNO2 in urban, rural, and mountain areas of the NCP. The winter indoor-heating period lasts from early November to 15 March of the following year (Ran et al., 2016), and thus the observations were made mostly during the heating period during which coal is intensively used. Detailed descriptions of the measurement sites are available in previous studies (Tham et al., 2016; Z. Wang et al., 2017; Xia et al., 2019), and a brief introduction is given here.

Our observations at the Wangdu site were part of the Campaign of Oxidation Potential Research for air Pollution in winter (COPPER). The Wangdu site is located in Dongbaituo Village, Hebei Province. Local villagers use coal stoves for cooking and heating during winter. National road G4 and provincial road S335 are 1 and 3 km to the west of the sampling site, respectively. Many heavy-duty trucks passed through G4 and S335 every night during the study period, emitting a large amount of NOx and particulate matter. Therefore, the site experienced heavy pollution from coal burning and road traffic (Peng et al., 2020).

The Beijing site is located at the Chinese Research Academy of Environmental Science (CRAES), which is 15 km northeast of the city center. The sampling site is surrounded by intra-city roads, commercial buildings, and residential buildings with few industrial facilities. When the prevailing wind originates from the north (i.e., remote mountainous regions), the site is upwind from the Beijing downtown area and thus is less polluted. However, when the wind originates from the south, the site receives pollutants from Beijing's urban areas in the NCP (Xia et al., 2019).

Mt. Tai is located approximately 40 km south of Jinan City (population: 8.9 million) and 15 km north of Tai'an City (population: 5.6 million) (Wen et al., 2018). Measurements were taken at Mt. Tai meteorological station (1534 m a.s.l.). The site is isolated from the anthropogenic emissions of tourist areas and is not affected by local emissions. The observation period, i.e., March to April, was in early spring in the NCP. However, considering the low temperature (4.6 ± 6.3∘) observed on top of Mt. Tai, this study considered the observation period to be winter to early spring.

N2O5 and ClNO2 were simultaneously measured by a chemical ionization mass spectrometer with a quadrupole mass analyzer (Q-CIMS; THS Instruments). The principles and calibrations of the Q-CIMS measurements are available in previous studies (Wang et al., 2016; Tham et al., 2016; Z. Wang et al., 2017; Xia et al., 2019). Briefly, the primary ions used in the Q-CIMS were iodide (I-) and its water clusters, which were generated using CH3I with an inline ionizer (210Po). The iodide adducts, namely IN2O5- and IClNO2-, were then detected by the mass spectrometer. An example of the mass spectrum is shown in Fig. S2. The integration time of the signals recorded by the Q-CIMS is shown in Table S1. The isotopic ratios of I35ClNO2- and I37ClNO2- in the ambient data were used to confirm the identity of ClNO2 (Fig. S3). Gas-phase mixtures of NO2 and O3 produced N2O5 for N2O5 calibration. The synthetic N2O5 was converted to ClNO2 by passage through a humidified NaCl slurry for ClNO2 calibration. The dependence of the N2O5 sensitivities on ambient RH was tested once in each campaign and used to calibrate the N2O5 data (Fig. S4a). ClNO2 sensitivities were found to be not affected by RH (Fig. S4b). Single-concentration calibrations of N2O5 and ClNO2 were performed every 1–2 d, which showed stable sensitivities of N2O5 and ClNO2 (Text S1 and Fig. S5). And the linearity of the N2O5 and ClNO2 signals to concentrations was checked via multi-concentration calibrations in the Mt. Tai campaign (Fig. S6). Backgrounds of N2O5 and ClNO2 were determined by passing the ambient air through glass wool once a day at different time. The background signals of N2O5 (3.3–7.7 pptv) and ClNO2 (1.0–7.5 pptv) were stable and independent of the time of the day (Fig. S7). The detection limits of N2O5 and ClNO2 were 6.9–7.3 and 3.8–5.3 pptv, which is defined here as 3 times the standard deviation of the background signals in 5 min (Table S2 in the Supplement). A virtual-impactor design (Peng et al., 2020) was adopted, and the sampling tube was replaced daily to minimize inlet artifacts.

The trace gases, particle number size distribution (PNSD), and ionic composition of aerosols and other species were simultaneously measured (Table S3). Online non-methane hydrocarbons were measured by gas chromatography–flame ionization detection–mass spectrometry (GC-FID/MS; Chromatotec Group) at the Beijing site (Zhang et al., 2017) and Wangdu site (Zhang et al., 2020). At Mt. Tai, we used canisters to collect air samples, which were analyzed using GC-FID/MS. In Wangdu and Mt. Tai, oxygenated volatile organic compound (OVOC) samples were collected on DNPH-coated sorbent cartridges followed by post-campaign analysis using high-performance liquid chromatography. The ionic compositions of PM2.5 (e.g., NH4+, NO3-, SO42-, and Cl-) were quantified by the Monitor for AeRosols and GAses in ambient air (MARGA, Metrohm) at the Beijing and Mt. Tai sites (Wen et al., 2018). An aerosol chemical speciation monitor (ACSM, Aerodyne Research Inc.) was utilized at the Wangdu site to monitor the non-refractory components of these ions in PM2.5. The concentrations of the NO3-, SO42-, and NH4+ measured simultaneously by the MARGA and ACSM were in good agreement, whereas the concentration of Cl- measured by the ACSM was slightly lower than that measured by the MARGA, which was possibly due to the significant proportion of refractory chloride, e.g., NaCl, present in the aerosols (Xia et al., 2020). We assumed that the particles sampled by a wide-range particle spectrometer (WPS) were spherical in shape and calculated the aerosol surface area density (Sa) and volume density (Va). A parameterization was adopted to consider the hygroscopic growth factor (GF) of aerosol sizes, as follows: GF=a×b+11-RH1/3 (Lewis, 2008), where the parameters a and b were derived as 0.582 and 8.460, respectively, in a previous field study over the NCP (Achtert et al., 2009).

Some analytical metrics were calculated from the observation data. P(NO3) was calculated using Eq. (1), where k1 represents the rate constant of Reaction (R9) (Atkinson and Lloyd, 1984). 1P(NO3)=k1×[O3]×[NO2] k(NO3) during the night was calculated using the measured mixing ratios of NO and non-methane hydrocarbons that were measured by GC (Sect. 2.3). As most OVOCs react with NO3 at much slower rates compared to those with hydrocarbons, especially alkenes (Atkinson and Arey, 2003), the OVOCs were not included in the calculation of k(NO3). Nonetheless, the k(NO3) might be slightly underestimated here. 2k(NO3)=∑ki[VOCi]+kNO+NO3[NO], where ki is the rate constant for a specific VOC + NO3 reaction and is adopted from Atkinson and Arey (2003), and kNO+NO3 represents the rate constant for Reaction (R11) (DeMore et al., 1997). The ambient concentrations of NO3 were estimated by assuming that NO3 and N2O5 were in dynamic equilibrium (DeMore et al., 1997). 3NO3=[N2O5]NO2Keq The loss rates of NO3 due to NO and VOCs were then calculated by kNO+NO3[NO][NO3] and ∑ki[VOCi][NO3], respectively.

The loss rate coefficient of N2O5 on the aerosol surface (k(N2O5)) is expressed as follows. 4k(N2O5)=0.25×c(N2O5)×Sa×γ(N2O5), where c(N2O5) represents the average molecular velocity of N2O5. The rate constants (k1, ki, and kNO+NO3) and equilibrium constant (Keq) are calculated as temperature-dependent parameters.

γ(N2O5) and φ(ClNO2) were estimated using steady-state analysis in applicable cases (Brown et al., 2006). This method assumes a steady state of N2O5, which means that the production rate of N2O5 is equal to its loss rate. We adopted the criteria described by Xia et al. (2020) to select the cases, namely low concentrations of NO, an increasing trend of ClNO2 concentrations, and stable air masses. Equation (5) was then established by plotting τN2O5-1×NO2×Keq against 0.25×Sa×CN2O5×NO2×Keq, with γ(N2O5) as the slope and k(NO3) as the intercept in the linear regression (Brown et al., 2003). Here, the derived γ(N2O5) was accepted when the regression had R2 > 0.5 and k(NO3) > 0. 5τN2O5-1×Keq×NO2≈0.25×CN2O5×Sa×Keq×NO2×γN2O5+kNO3 φ(ClNO2) was then calculated using the following equation: 6φ(ClNO2)=d[ClNO2]/dtkN2O5[N2O5], where d[ClNO2] / dt and [N2O5] represent the increasing rate of ClNO2 production and the average concentration of N2O5, respectively, within the selected cases.

An observation-based chemical box model was utilized to simulate the concentrations of Cl and ROx radicals and the production and loss pathways of O3. The detailed model description is available in Peng et al. (2020). Based on Master Chemical Mechanism (MCM) v3.3.1 (Jenkin et al., 2015), Peng et al. (2020) modified the chemical mechanisms to include up-to-date gas-phase chlorine and bromine chemistry. The observed N2O5, ClNO2, NOx, HONO, O3, jNO2, and related species were constrained in the model for every 10 min of model time, after interpolating or averaging the data (Table S4). The mixing ratios of non-methane hydrocarbons (NMHCs) and OVOCs (Sect. 2.3) were constrained every hour. As OVOCs were not measured in Beijing in this study, we adopted the concentrations of OVOCs measured in previous studies in winter Beijing (Gu et al., 2019; Qian et al., 2019). We also assumed the CH4 mixing ratio to be 2000 ppbv, which was the mean value in summer at Wangdu (Tan et al., 2017), for our Wangdu and Beijing sites. We acknowledge that the adoption of the summer CH4 concentrations for our winter studies may underestimate the CH4 level, but the effect on the ROx budget is expected to be insignificant due to much smaller contribution of CH4 to ROx than NMHCs and OVOCs in polluted environments. The photolysis frequencies of ClNO2, O3, and other species were simulated according to the solar zenith angle using the tropospheric ultraviolet and visible (TUV) radiation model and scaled by the observed jNO2 values. Numerical experiments were conducted by constraining (Case 1) and not constraining ClNO2 data (Case 2) at each site. The differences in the radical concentrations and O3 budgets between Cases 1 and 2 represented the effect of ClNO2. For example, the increase in ROx (%) due to ClNO2 was calculated by (ROx_w – ROx_wo) / ROx_wo, where ROx_w represents the concentration of ROx in Case 1 with ClNO2 constrained in the model, and ROx_wo represents the concentration of ROx in Case 2 without ClNO2 constrained.

The time series of N2O5 and ClNO2 levels in the three campaigns are displayed in Fig. 1. Overall, elevated levels of N2O5 and ClNO2 were observed with different patterns at each site. The ground sites (Wangdu and Beijing) were characterized by high NOx levels (83.2 ± 81.3 ppbv and 35.6 ± 27.3 ppbv, respectively) and low O3 levels (8.5 ± 8.8 ppbv and 17.3 ± 11.4 ppbv, respectively), whereas the mountain site, Mt. Tai, was marked by relatively lower NOx levels (2.4 ± 2.0 ppbv) and higher O3 levels (64.6 ± 14.7 ppbv) (Fig. S8). The campaign-averaged mixing ratios of ClNO2 were similar at the ground sites (71 ± 132 pptv and 76 ± 103 pptv in Wangdu and Beijing, respectively) and were significantly lower than that at Mt. Tai (179 ± 247 pptv). The nocturnal ratio of ClNO2 / N2O5 at each site displayed large day-to-day variability, which was positively dependent on the ambient RH (Fig. S9) and, to a lesser extent, positively correlated with Sa (figure not shown).

The campaign-averaged diurnal patterns of the mixing ratios of N2O5, ClNO2, and related species are depicted in Fig. 2. ClNO2 levels typically exhibited a daily cycle, peaking at night and decreasing during the day. The diurnal pattern of ClNO2 at the Wangdu site in winter was an exception, with elevated concentrations (10th–90th percentiles) around midday (12:00–14:00 local time, LT), which resulted from a noontime peak in ClNO2 concentrations during a few days at Wangdu. The detailed observation results from each site are separately introduced as follows.

The nocturnal production of ClNO2 was insignificant in Wangdu despite the presence of abundant Cl- (3.3 ± 3.2 µg m-3 throughout the observation), which likely originated from the intensive residential coal combustion in the area (Peng et al., 2020). The Wangdu site experienced high mass concentrations of PM2.5 (a maximum of approximately 450 µg m-3) and very large mixing ratios of NO (a maximum of approximately 350 ppbv). The wind rose analysis showed that the high concentrations of NO originated from the west of the sampling site where two major roads were located. Numerous heavy-duty trucks on these roads were responsible for high NO concentrations. The presence of abundant NO inhibited N2O5 formation by consuming O3 and NO3 at the Wangdu site. When the ambient concentrations of NO substantially decreased, e.g., on 10 December, the N2O5 mixing ratios increased to 1 ppbv. The mixing ratios of ClNO2 were mostly low (< 200 pptv) during the night. The relationship between nighttime levels of ClNO2 and grouped NO and NOx concentrations is shown in Fig. 3. ClNO2 showed higher levels when the NO mixing ratios were below 10 ppbv and NOx mixing ratios ranged from 10–20 ppbv (Fig. 3a, d). However, significant daytime peaks in ClNO2 mixing ratios were observed on 14 and 28 December, reaching approximately 0.8 and 1.3 ppbv, respectively. The daytime peaks in ClNO2 concentrations at the three sites are discussed in detail in Sect. 3.3. For comparison, the ambient mixing ratios of NO in the summer campaign at Wangdu were much lower (mostly 0–10 ppbv), and O3 mixing ratios were much higher (i.e., exceeded 90 ppbv on most days), which favored the production of N2O5 and ClNO2 (Tham et al., 2016).

The winter Beijing observations showed that there was significant production of N2O5 but limited conversion of N2O5 to ClNO2 in dry conditions. The observation period in Beijing was divided into polluted days (24 h PM2.5 > 75 µg m-3; China's Grade II air quality standard for PM2.5) and clean days (24 h PM2.5 < 35 µg m-3; Grade I standard). The polluted periods were characterized by simultaneous high levels of PM2.5 and NO, e.g., on 19 January. The clean periods were marked by relatively high mixing ratios of O3 and low levels of PM2.5 and NOx, e.g., from 8 to 11 January. Both polluted and clean conditions were unfavorable for ClNO2 formation owing to the high concentrations of NO on the polluted days and the low concentrations of NO2 and aerosols on the clean days. Moreover, the RH observed in Beijing was typically below 40 %, which indicated relatively slow heterogeneous loss of N2O5 and slow ClNO2 formation. Consequently, N2O5 mixing ratios frequently accumulated to elevated levels, exceeding 0.4 ppbv on 10 of the 26 observation nights, and the mixing ratio of ClNO2 was mostly below 0.4 ppbv. Nighttime levels of ClNO2 in winter Beijing were higher when NO mixing ratios ranged from 0–10 ppbv and NOx mixing ratios ranged 20–50 ppbv (Fig. 3b, d). The highest mixing ratios of ClNO2 were observed (up to approximately 0.8 ppbv) when the site occasionally intercepted air masses with a higher RH (approximately 75 %), e.g., on the night of 22 January. This result is similar to the previous observation in Beijing (Xia et al., 2019), in which the ratio of ClNO2 to N2O5 increased significantly from late spring with a low RH (10 %–30 %) to early summer with a higher RH (20 %–80 %). The overall mixing ratios of ClNO2 in the present Beijing study in winter were also significantly lower than those reported in summer (maximum of 1.4 to 2.9 ppbv) in other studies (Le Breton et al., 2018; Zhou et al., 2018).

Elevated mixing ratios of ClNO2 (i.e., above 0.5 ppbv) were frequently recorded at the Mt. Tai station in winter. Nighttime levels of ClNO2 were slightly higher when NO levels were below 0.5 ppbv (Fig. 3c) and showed a positive correlation with NOx levels (Fig. 3f). High concentrations of PM2.5 (34.5 ± 27.3 µg m-3) and high RH (63.6 ± 27.1 %) favored the ClNO2 formation at Mt. Tai. The maximum level of ClNO2 (approximately 1.7 ppbv) was observed just before midnight on 18 March, which was slightly lower than the highest concentration observed at Mt. Tai in the summer of 2014 (Z. Wang et al., 2017). The elevated concentrations of ClNO2 observed in the previous summer study at Mt. Tai were due to emissions from distinct coal-fired power plants, whereas this winter study found that coal burning had less effect on concentrations of ClNO2. The campaign-averaged levels of SO2 and particulate SO42- were 1.6 ± 1.6 ppbv and 3.6 ± 2.9 µg m-3, respectively, during the winter observations, which were significantly lower than those observed in the summer campaigns (2.9 ± 3.7 ppbv and 14.8 ± 9.0 µg m-3, respectively). The decreases in SO2 and sulfate were attributed to strengthened emission control for coal-fired power generation during 2014–2018 and also less transport of emissions from the ground to the Mt. Tai site (1534 m a.s.l.) in late winter and early spring compared with that in summer.

We compared the observed winter concentrations of ClNO2 with those reported in previous studies in Asia, North America, and Europe (Fig. 4). The highest winter concentrations of ClNO2 to date were observed in southern China, with a maximum level of 4.7 ppbv at a mountain top in Hong Kong in aged urban/industrial plumes from the Pearl River Delta (PRD) (Wang et al., 2016) and 8.3 ppbv during a severe pollution episode within the PRD (Yun et al., 2018). The high-concentration ClNO2 events in southern China were due to concurrent high levels of PM2.5 and O3 (e.g., 400 µg m-3 and 160 ppbv found by Yun et al., 2018), which contrast the high concentrations of PM2.5 and low concentrations of O3 over northern China during the cold winter. The winter mixing ratios of ClNO2 in the United States and Europe range from approximately 0.3 ppbv in urban California (Mielke et al., 2016) and urban Manchester (Priestley et al., 2018), respectively, to 1.3 ppbv in the outflow of coastal urban areas (Riedel et al., 2013; Haskins et al., 2019). In general, the winter concentrations of ClNO2 over northern China were comparable to or slightly higher than those observed in the United States and Europe.

To gain insight into the processes controlling the variability in concentrations of ClNO2, nocturnal P(NO3) and NO3 loss pathways were compared using Eqs. (1–5) in Sect. 2.4. The average P(NO3) was comparable at the three sites in winter, ranging from 0.15 to 0.25 ppbv h-1, and these rates were significantly lower than the respective summer values (Fig. 5a). The lower P(NO3) in winter was caused by both lower k1 and lower [NO2] × [O3] in winter (see Eq. 1). The lower k1 in winter is caused by lower temperature in winter, while the lower [NO2] × [O3] in winter is mainly caused by less photochemical production of O3 and more NO that consumes the available O3 in winter (Table S5). Nighttime NO3 removal through NO3 and N2O5 was estimated by comparing k(NO3) × [NO3] (Eqs. 2–3) and k(N2O5) × [N2O5] (Eqs. 4–5). The average γ(N2O5) values derived from each campaign (Table S6 and Fig. S10) were used in Eq. (4). The nighttime NO3 loss via NO titration and VOC oxidation was greater than the N2O5 heterogeneous loss in all the winter and summer campaigns (Fig. 5b). These were the campaign average results. In contrast, the N2O5 loss was greater than the NO3 loss in selected cases in summer at Mt. Tai (Wang et al., 2017c). To determine the nocturnal loss of NO3, we further compared the N2O5 / NO3 ratio and γ(N2O5) at the three sites.

The thermal decomposition of N2O5 was suppressed in winter and resulted in high ratios of N2O5 / NO3 (Fig. 6a; up to approximately 1000), which favored N2O5 loss over NO3 loss. However, the γ(N2O5) in winter was systematically lower than that in summer (Fig. 6b), which indicated slower N2O5 loss in winter. A previous field study in winter Beijing also reported small values of γ(N2O5), ranging from < 0.001 to 0.02 (Wang et al., 2020). This result differs from previous laboratory studies, which reported larger γ(N2O5) on (NH4)2SO4 aerosols at lower temperatures (Hallquist et al., 2003; Griffiths and Anthony Cox, 2009). It is possible that other factors, such as RH and aerosol composition (aside from (NH4)2SO4), had a large influence on γ(N2O5). The limited number (two to four) of γ(N2O5) values obtained in each winter campaign (Table S6) may have also caused a bias in the estimation of the overall γ(N2O5). The opposite effects – a higher N2O5 / NO3 ratio and lower γ(N2O5) in winter – offset each other in Wangdu (Fig. 6b) but favored N2O5 loss in Beijing and NO3 loss at Mt. Tai compared with those in the respective summer campaigns. The higher concentrations of ClNO2 at Mt. Tai during the winter campaigns may be attributable to higher φ(ClNO2) values in Mt. Tai (Fig. 6c).

In the winter campaigns, high concentrations of ClNO2 were sustained after sunrise. Distinct peaks in ClNO2 concentrations were observed on 3–4 d in each campaign, as shown in Fig. 7 displaying one case at each site. Other daytime cases from the three sites are shown in Figs. S11–13. The validity of the daytime peaks was checked by performing isotopic analysis of ClNO2, background detection, and on-site calibration. The signals of I35ClNO2- and I37ClNO2- were well correlated (R2 > 0.99) during daytime peaks in ClNO2 concentrations (Fig. S3a–c) and calibrations (Fig. S3d–f). The ratio of I37ClNO2- to I35ClNO2- (0.32–0.35) was consistent with the natural isotopic ratio of 37Cl to 35Cl. The background signals of ClNO2 were checked when its daytime peaks in concentrations were observed, and no increase in the background was found. We also checked the signal of primary ions (IH2O-) and found no abnormal changes when ClNO2 concentrations showed daytime peaks. These results confirmed that the daytime peaks in ClNO2 concentrations were real atmospheric phenomena.

The daytime ClNO2 episodes usually occurred from 10:00 to 11:00 LT at each site. The highest daytime mixing ratio of ClNO2 was 1.3 ppbv (5 min average) observed at 11:30 on 28 December 2017 in Wangdu. In comparison, the daytime ClNO2 concentration observed in the previous summer study at Wangdu (Tham et al., 2016) reached a maximum in the early morning (08:00 LT) and declined to several parts per trillion by volume at 11:00. Attenuated solar radiation was observed during the days with daytime peaks in ClNO2 concentrations. For example, the daily maximum rates of jNO2 (1 min average) for the Wangdu case shown in Fig. 7a (2.5 × 10-3 s-1) was significantly lower than the highest rate observed during this campaign (6.0 × 10-3 s-1). The attenuated solar radiation reduced the photolysis of ClNO2, which allowed it to persist for a longer period during the day. The chemical data showed contrasting features during the daytime peaks in ClNO2 concentrations at the three sites. At Wangdu, ClNO2 concentrations showed a sharp peak while the concentrations of other pollutants decreased (Fig. 7a); in Beijing, the daytime peak in ClNO2 concentrations appeared with little simultaneous change in the NO3-, NOx, and O3 levels after sunrise (Fig. S12a). In two cases, daytime peaks of ClNO2 concentrations at Mt. Tai (Figs. 7c and S13c) occurred together with significant increases in NO3-, NOx, and PM2.5 levels, whereas O3 concentrations decreased after sunrise and resumed previous levels.

The daytime peaks in ClNO2 concentrations were likely caused by the transport of air masses to the respective sites. In situ production of ClNO2 was limited during the days on which significant daytime ClNO2 occurred, because the mixing ratios of N2O5 were near the detection limit of the instrument (several parts per trillion by volume). The photochemical lifetime of ClNO2 at 10:00 am LT was estimated to be 1–2 h, based on the inverse of jClNO2, which allowed the transport of ClNO2 produced elsewhere to the observation sites. As daytime peaks in ClNO2 concentrations appeared at both the ground and mountain sites, the high-ClNO2 region may exist in the residual layer above the nocturnal mixing layers. At sunrise, ClNO2-rich air masses may be transported downward to the ground sites (Wangdu and Beijing) and upward to the mountaintop site (Mt. Tai). The downward transport of ClNO2 at Wangdu in summer has been illustrated by Tham et al. (2016), and the upward transport to the top of Mt. Tai has also been implicated by the increasing daytime concentrations of O3 and other pollutants (e.g., Gao et al., 2005; Zhou et al., 2009; Jiang et al., 2020). Measurements in the residual layers are needed to further investigate the transport of ClNO2 within the entire boundary layer.

We used the box model (Sect. 2.5) to show the impact of ClNO2 on photochemical oxidation at the three sites (Fig. 7a–c). In campaign-averaged conditions, the impact of ClNO2 was minor, owing to the low daytime concentrations of ClNO2. The daytime-averaged P(Cl) (06:00–18:00 LT) from ClNO2 photolysis was in the range of 0.03–0.06 ppbv h-1, with the peak values of 0.07–0.12 ppbv h-1, and the photolysis of ClNO2 enhanced the daytime ROx concentrations by 1.3 %–3.8 % and net O3 production by 1.3 %–6.2 % at the three sites (figures not shown). Such impacts were lower than those during summer at Wangdu (Tham et al., 2016).

However, the impact of ClNO2 increased considerably in the cases of daytime-peak concentrations, as shown in Fig. 8. The daytime-averaged P(Cl) values from ClNO2 photolysis were 0.15 ± 0.13 (maximum of 0.46), 0.11 ± 0.09 (maximum of 0.32), and 0.19 ± 0.20 (maximum of 0.74) ppbv h-1 at Wangdu, Beijing, and Mt. Tai, respectively (Fig. 8a–c). The winter P(Cl) peak in Wangdu (Fig. 8a, 0.46 ppbv h-1) was twice the summer average value (0.24 ppbv h-1) (Tham et al., 2016). The P(Cl) during the daytime peaks of ClNO2 in this study is significantly higher than that in Riedel et al. (2012) (maximum ∼ 0.08 ppbv h-1) but slightly lower than that in Haskins et al. (2019) (maximum ∼ 1.3 ppbv h-1). P(Cl) from other sources (e.g., the HCl + OH reaction) was minor (8.8 %–14.5 %) during these cases. The relative importance of ClNO2 in primary radical production varied among these sites. ClNO2 had a minor contribution in Beijing but became increasingly important in Wangdu and Mt. Tai (Fig. 8b, c). HONO photolysis was the most important source of OH at the two ground sites, whereas O3 was also important at Mt. Tai.

The liberated Cl (mostly from ClNO2 photolysis) accounted for 28.5 %–57.7 % of the daytime (06:00–18:00 LT) oxidation of alkanes, 6.1 %–13.7 % of that of alkenes, 5.3 %–14.2 % of that of aromatics, and 4.6 %–6.0 % of that of aldehydes in the cases of high levels of daytime ClNO2. The Cl + VOC reactions enhanced the production of OH, HO2, and RO2 by up to 15 %–22 %, 24 %–31 %, and 36 %–52 %, respectively (Fig. 8d–f). The photolysis of ClNO2 increased the daytime net O3 production by 5.4 ppbv (18 %), 2.8 ppbv (17 %), and 2.6 ppbv (13 %) at Wangdu, Beijing, and Mt. Tai, respectively (Fig. 8g–i). These results indicate the considerable impact of daytime ClNO2 on the atmospheric oxidative capacity and production of secondary pollutants.

The impact of Cl in the NCP is likely larger than the result shown above. Our model calculations considered photolysis of ClNO2 (and HCl + OH) as the source of Cl, but not other photolabile Cl-containing gases. However, in the Wangdu field campaign, we frequently observed elevated daytime concentrations of bromine chloride (BrCl) and molecular chlorine (Cl2), which dominated the Cl production (Peng et al., 2020). In addition, our ClNO2 measurements were conducted at polluted ground-level sites and at a high mountain site (1534 m a.s.l.), which are not in the nocturnal residual layer where strong ClNO2 production is expected to occur (Zhang et al., 2017). It is thus highly desirable to measure ClNO2 in the residual layer in future studies to comprehensively assess the role of ClNO2 in the lower part of the atmosphere.