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Summertime distributions of peroxyacetyl nitrate (PAN) and peroxypropionyl nitrate (PPN) in Beijing: Understanding the sources and major sink of PAN

Peroxyacetyl nitrate (PAN), a major secondary pollutant in the atmosphere, has received much concern for its particular importance in atmospheric chemistry and adverse effects on human and plants. Atmospheric PAN and PPN were measured by using a gas chromatograph equipped with electron capture detector (GC-ECD) from June to September 2010, and the source and major sink for PAN were firstly studied in Beijing. The distinct diurnal variations of PAN and PPN with maximum in the afternoon were observed, and the mean and maximum values were 2.61 +/- 2.57 ppbv = 839) and 12.5 ppbv for PAN and 0.52 +/- 0.38 ppbv (N = 152) and 2.16 ppbv for PPN during the measuring period, respectively. Good correlation (R = 0.85) between PAN and PPN with a slope (Delta PPN/Delta PAN) of 0.134 indicated that anthropogenic volatile organic compounds (AVOCs) dominated the photochemical formation of PANs in Beijing. Further, we found acetaldehyde was the predominant carbonyl precursor of PAN with the contribution of 59.7% to the total peroxyacetyl (PA) radical. Methyl glyoxal, methacrolein, acetone, methyl vinyl ketone, and biacetyl contributed 7.1%, 8.8%, 19.7%, 3.4%, and 13% to total PA radical, respectively. Anti-correlation between PAN concentrations and the NO/NO2 ratios was found during the whole investigating period. In addition, the amount of PAN lost by thermal decomposition (TPAN) accounted for remarkable fractions of PAN observed under high temperature during both daytime and nighttime. (C) 2014 Elsevier Ltd. All rights reserved.


Introduction
Peroxyacetyl nitrate (PAN) and its homolog compounds (hereafter related to PANs, RC(O)OONO 2 ) are typically produced from photochemical processes involving nitrogen oxides (NO x,NO and NO 2 ) and volatile organic compounds (VOCs) (Aikin et al., 1982;Lonneman et al., 1976;Singh and Hanst, 1981). PANs are thermally unstable in the lower troposphere, but are quite stable in the upper troposphere and can be transported over long distances from the polluted continental regions into the remote troposphere (Honrath et al., 1996). The release of NO 2 from the thermal decomposition of PANs has been considered as an important NO x source for remote troposphere where they may contribute to the photochemical production of O 3 (Singh, 1987). Additionally, PANs have been recognized as lachrymators (Altshuller, 1978), mutagens (Shepson et al., 1986), and phytotoxins (Taylor, 1969). Although the atmospheric concentration of PANs in polluted areas is usually about one magnitude lower than that of O 3 , the phytotoxic effect of PANs is about 1e2 magnitude greater than that of O 3 .
PAN, the most abundant species among PANs, has been extensively measured in various regions since the first recognition of PAN in a photochemical smog (Stephens et al., 1956). The mixing ratios of atmospheric PAN range from tens pptv (parts per trillion by volume) in remote areas (Bottenheim and Gallant, 1989;Moore and Remedios, 2010;Singh and Salas, 1983) to several ppbv (parts per billion by volume) in cities (Duane et al., 2002;Grosjean et al., 2002;Lee et al., 2008;McFadyen and Cape, 2005). In China, photochemical pollution has gradually provoked much concern with extensive studies of O 3 and VOCs (Shao et al., 2006(Shao et al., , 2009Wang et al., 2003Wang et al., , 2006Xu et al., 2008). However, the published researches related to PANs in rural and urban atmosphere of China are relatively sparse, and a few groups implemented sporadic observations (Liu et al., 2010;Wang et al., 2010;Wang and Zhang, 2007;Yang et al., 2009;Zhang and Tang, 1994). In the megacity of Beijing, only few PAN measurements have been conducted in summertime, and the remarkably high levels of PAN (up to 14 ppbv) observed surpass those measured over other urban regions in recent years (Liu et al., 2010;Wang and Zhang, 2007). However, previous studies only focused on PAN's concentration levels and diurnal variations, its source and major sink were not clearly understood. To the best of our knowledge, only Liu et al. (2010) employed a 1-D version of a chemical transport model to study the dominant source for atmospheric PAN in Beijing (Liu et al., 2010).
In this study, atmospheric PAN and PPN were measured from June to September 2010 in Beijing. In terms of the source of PAN, we discussed the relationships among PAN, PPN, and NO x and further estimated the peroxyacetyl (PA) radical production from acetaldehyde, acetone, methacrolein (MACR), biacetyl, methyl glyoxal (MGLY), and methyl vinyl ketone (MVK), respectively. In addition, the PAN lost by thermal decomposition (TPAN) for clear days chosen from each month was calculated to investigate its thermal decomposition as major sink in the atmospheric chemical process of PAN.

Site description
In this study, field sampling was conducted on the sixth floor of the Environmental Technological building (20 m above the ground level) in the Research Center for Eco-Environmental Sciences (RCEES), which is located in the north of Beijing city (39.8 N, 116.5 E) between the 4th and 5th ring roads. The surroundings of RCEES are several institutes, campuses, residential apartments, and two side roads with moderate traffic.

Experimental methods
The detail information about PAN measurement and calibration were described in our previous study and summarized briefly here (Zhang et al., 2012a,b). Ambient air was sampled into the GC loop (0.5 mL) by a mini-pump (NMP830KNDC, KNF Inc.) via a Teflon tube (3 mm OD, 6 m length) with a flow rate of~1.5 L min À1 . Each sampling time was 20 s and the pressure was equilibrated for 5 s with the pump power off, and then the air sample was injected into GC-ECD via a six-port valve. Capillary column of DB-1 (30 m length, 0.32 mm ID, 1.0 mm film thickness) was adopted as separation column. The temperatures of the separation column and the ECD were set to be 20 C and 38 C, respectively. The detection limits (three times the signal to noise) of PAN and PPN were 22 and 36 pptv, respectively. The calibration of instrument for PAN was achieved by Model 42i NO x analyzer (Thermo Scientific, USA), according to the in-situ formation of PAN via photochemical reaction of a gaseous mixture of excess acetone with NO in a 10 L Teflon bag under irradiation of 254 nm low-pressure mercury lamps. NO could be completely converted to PAN in presence of excess acetone after 5 min irradiation. Then the formed PAN concentration in the gaseous mixture was quantified by NO x analyzer and simultaneously detected by GC-ECD. The overall uncertainty for measuring atmospheric PAN was estimated to be ±15% based on the combined uncertainties of the GC precision and the uncertainties of NO x calibration system. The calibration was conducted every two days and the results showed no significant drift within one week. Atmospheric O 3 and NO x were measured using ozone analyzer (Model 49i, Thermo Scientific) and NO x analyzer (Model 42i, Thermo Scientific), respectively. The sampling frequency was half hour or 1 h, mostly conducting from 09:00 am to 21:00 pm (LT, Beijing time). The sampling days were randomly selected for 7 days in June, 13 days in July, 7 days in August, and 9 days in September.
The VOCs measurements including acetaldehyde, acetone, MVK, and MACR were also concurrently conducted by our group from June to August 2010 at this sampling site (Zhang et al., 2012a(Zhang et al., ,b, 2014. The analysis procedure was mainly based on the EPA TO-11A method (EPA, 1999), and the details have been described in our previous publications (Pang and Mu, 2006). Carbonyls were sampled at least 4 days per month, with carbonyls' samples collected at 2 h interval.

Ambient concentrations of PAN and PPN
PAN could be detected in nearly all samples during the whole period, whereas PPN was frequently below the detection limit of the GC-ECD except August. In June 2010, we stopped the integrator after the elution of the PAN peak due to without the calibration of PPN, so no PPN measurements are available for this period. Fig. 1 shows the time series of the mixing ratios of ambient PAN and PPN. During the investigating period the mean and maximum values were 2.61 ± 2.57 ppbv (N ¼ 839) and 12.5 ppbv for PAN and 0.52 ± 0.38 ppbv (N ¼ 152) and 2.16 ppbv for PPN, respectively. Apparently, the concentrations of PAN in July were much lower in comparison with other months, which were possibly ascribed to the extremely synoptic situation. During July there were sixteen rainy days, six clear days with extremely high temperature (T > 37 C). For these clear days in July, the average temperature was up to 32 C, higher than 28 C in June, 29 C in August, and 26 C in September. It's well known that the rainy days often represent weak photochemical reactivity which limits PAN formation, and the extremely high air temperature during the clear days greatly accelerated PAN thermal decomposition as discussed in the Section 3.5. In addition, O 3 and NO x were both found to be generally lower in July than other months. It was ascribed to the weak atmospheric photochemical activity which inhibits the formation of O 3 and the significant effect of wet deposition on the removal of NO x during the continuous rainy days in July, respectively. But for the primary pollution of CO, it was higher in July than other months probably due to the weak diffusion under the stable synoptic condition.
The mean diurnal profiles of PAN and PPN measured by this study in August are illustrated in Fig. 2 for comparison with other studies. The mean daily maximum PAN concentration of 5.5 ppbv observed in this study was in good agreement with the value (5 ppbv) reported by Liu et al. (2010) in August 2007. Compared with other studies in Beijing, the maximum PAN value (7.0 ppbv) observed by this study was higher than 2.5 ppbv observed in August 2005 (Wang and Zhang, 2007), but lower than 11.2 ppbv in August 2006 (Yang et al., 2009) and 14.0 ppbv in August 2007 (Liu et al., 2010). It's possibly ascribed to the interannual differences of the PAN precursors and the synoptic situation in Beijing.
It is interesting to note that evident elevation of PAN concentration after sunset (around 20:00) was frequently observed (Figs. 2 and 3), even PAN maximum appearing after sunset on 6 September (Fig. 3). The elevation of PAN after sunset was coincident with the increases of CO and NO x which were mainly from the primary emission of vehicles. There were no evidences that PAN could be directly emitted from exhaust of vehicles, e.g., no elevation of PAN was observed when CO and NO x significantly increased during early morning hours. The elevation of PAN concentration after sunset in Beijing was also observed by Wang and Zhang (2007) and they concluded the elevation was due to long distance transportation. Considering the similar sources and sinks for PAN and O 3 (for example, good correlation (R ¼ 0.76, N ¼ 69) between PAN and O 3 was found in September), if long distance transportation made contribution to the PAN elevation, O 3 would have the similar behavior. However, steady decrease of O 3 after sunset was   observed. Thus we speculated that dark reactions or other sources might contribute to the elevated PAN based on the fact that photochemical reactions ceased after sunset. The special diurnal variations of PAN and O 3 on 6 September provided some clues for disclosing the reason for PAN elevation after sunset. On 6 September PAN and O 3 both exhibited evident difference from other typical clear days, e.g., instead of quick decrease of PAN after around 15:00 for most typical clear days, PAN concentration on 6 September continually increased from 8:00 to 21:00; high O 3 concentration after its maximum on the day lasted much longer time than those on most typical clear days. On 6 September the elevated PAN after sunset was negative correlated (R ¼ À0.83) with the decreased O 3 (consumed by NO and/or combined with NO 2 ): [PAN] ppb ¼ (À0.06 ± 0.03)*[O 3 ] ppb þ (14.6 ± 1.90) ppb. We inferred that the much higher O 3 concentration from later afternoon to sunset (around 20:00) on 6 September than those on most other clear days would facilitate conversion of NO to NO 2 and hence greatly restrain PAN thermal decomposition rate (Grosjean et al., 2001). On the other hand, the combination of O 3 and NO 2 both with the relative high concentrations just after sunset possibly favored the formation of NO 3 radicals which would accumulate due to no photolysis. NO 3 is reactive species towards various VOCs, and has been considered as the dominant species triggering nighttime atmospheric chemistry, resulting in formation of various organic radicals including peroxyacetyl (PA) radical which reacts with NO 2 to form PAN. The significant amount of PAN thermal decomposition after sunset would also make great contribution to OH radical which further initiates VOCs oxidation to partially compensate PAN thermal loss and CH 3 radical which leads to formaldehyde in turn. It's well confirmed by the findings of Duane et al. (2002) that isoprene, methacrolein (MACR), methyl glyoxal (MGLY), and methyl vinyl ketone (MVK) both reached their respective peak after sunset (ca. 20:00) accompanying with high concentration of formaldehyde in summer in Insubria, Northern Italy. In addition, inferring from the results of model simulation, Liu et al. (2010) expected that large gradient of PAN vertical distribution might exist in Beijing, the downward transport of PAN produced aloft in the boundary layer was also suspected to make great contribution to the evident elevation of PAN after sunset.

Relationship between PAN and PPN
PAN and PPN are expected to be highly correlated due to their similar photochemical process in the atmosphere (Singh, 1987). In this study, good correlation (R ¼ 0.85, N ¼ 152, p < 0.0001) between PAN and concurrently detectable PPN was found: PPN However, their precursors are not identical. The precursors of PAN, leading to the peroxyacetyl (PA) radical, come from both human activity and biogenic emissions, while the precursors of PPN, leading to the peroxypropionyl radical, are predominantly anthropogenic in origin (Grosjean et al., 2001;Roberts et al., 2004). Thus, the ratio of PPN to PAN provides an indicator to evaluate the relative importance of anthropogenic and biogenic VOCs in PANs formation. With extensive measurements, Roberts et al. (2002Roberts et al. ( , 2004 concluded that the formation of PANs was dominated by anthropogenic sources when the slope (DPPN/DPAN) was between 13.5% and 17.2% and was dominated by biogenic sources when the slope was less than 10% (Roberts et al., 2004(Roberts et al., , 2002. The slope of 13.4% reported here, which was comparable with the reported values of 0.135e0.172 in polluted urban areas (Roberts et al., 2002(Roberts et al., , 2004, indicated that anthropogenic VOCs (AVOCs) were dominating the photochemical production of PANs during summer in Beijing.

Relationship between PAN and NO x
As shown in the reaction Equations R1eR3, PAN formation depends on complex atmospheric chemical reactions of VOCs and NO x , and the thermal decomposition rate of PAN depends on the ratio of NO to NO 2 (NO/NO 2 ) in addition to air temperature. Therefore, atmospheric PAN concentration was correlated with NO x concentration and the NO/NO 2 ratio. Owing to the same responds of PAN and NO 2 on the analyzer (Gerboles et al., 2003;Zhang et al., 2012a,b), we defined NO 0 2 and NO 0 x representing the actual concentrations of NO 2 and NO x after subtracting PAN concentration from the NO 2 and NO x values displayed in the NO x analyzer, respectively. Very weak positive correlation (R ¼ 0.07, N ¼ 537, P ¼ 0.101) between PAN and NO 0 x was found during our investigated period, which was in contrast with the result observed in clean marine air of the Pacific where PAN concentration was found to increase with increasing NO x (Singh et al., 1985). In the remote areas, NO x mostly presents as NO 2 , whereas in the polluted city of Beijing, large amount NO emission from vehicles would greatly favor PAN thermal decomposition, leading to the weak correlation. Fig. 4 showed PAN concentrations as a function of NO=NO 0 2 in Beijing. Apparently, almost no high PAN concentrations appeared at high NO=NO 0 2 ratios, further confirming that NO favors PAN thermal decomposition. The relationship between PAN and NO=NO 0 2 ratio obtained by this study was in good agreement with previous studies in other cities (McFadyen and Cape, 2005;Schrimpf et al., 1998).

Sources of atmospheric PAN
To the best of our knowledge, two types of estimation method have been implied to investigate the sources of atmospheric PAN: the first one is based on estimating the formation potentials of various hydrocarbons to PAN formation using Master Chemical Mechanism (MCM) model (Altshuller, 1993) and SAPRC 97 model (Grosjean et al., 2002), the second is recently confined to estimate the direct contribution of acetaldehyde, biacetyl, MACR, MVK, acetone, and MGLY to PA production using steady-state model (Cleary et al., 2007;LaFranchi et al., 2009). Specifically for the second estimation, PAN formation occurs upon the production of PA radical derived from the oxidation or the photolysis of acetaldehyde, acetone, MGLY, MACR, MVK, and biacetyl (CH 3 COCOCH 3 ) (Altshuller, 1993). The dominant pathways leading to PA radical from the six carbonyls in the troposphere are summarized in Fig. 5 ( LaFranchi et al., 2009). In this study, based on the fact that its lifetime is very short (t~10 s) relative to oVOCs (t~1e2 h) (LaFranchi et al., 2009), we likewise considered PA radical in steady-state and partly followed the method simplified by LaFranchi et al. (2009) to roughly estimate the direct contributions of six compounds to PA production. Atmospheric PAN formation is determined by the formation rates of PA radicals which depend on the concentrations of the carbonyls and OH radicals, the rate constants of their reactions with OH or photolysis, and the branching ratios of the reactions for formation of PA radicals. The average concentrations of the carbonyls, the rate constants of OH reactions and photolysis, and the branch ratios of PA formation adopted in this study are summarized in Table 1. Note that during the investigating period, the data of acetaldehyde and acetone were detected in all samples (N ¼ 69) with average concentrations of 4.93 and 11.1 ppbv, the MACR (co-elute with butanal) data were often obtained with the average concentration of 0.84 ppbv. MVK was sometimes below detection limit and averaged to 0.30 ppbv with limited data (N ¼ 21). MGLY was estimated to be 0.47 ppbv according to the observed ratio (0.62) of MVK to MGLY (Shao et al., 2009). To date, biacetyl has not been measured in Beijing. In this study we roughly estimated biacetyl concentration as 0.09 ppbv approximately as that of glyoxal which was inferred from the reported ratio of 0.32 of glyoxal to MVK in the summer of Beijing (Shao et al., 2009).
The PA formation rates estimated for the six carbonyls as well as their proportion to the total are shown in Fig. 6. It is evident that acetaldehyde is the dominant source for PA formation with the contribution of 59.7% among the six carbonyls in Beijing, which was in agreement with the conclusion that acetaldehyde is often the dominant PA radical source over urban areas (Kondo et al., 2008;Roberts et al., 2001). The photolysis of acetone is the second largest contribution to the total PA radical with the production rate of ca. 3.0 Â 10 6 molecules cm À3 s À1 . MACR and MGLY contributed to the formation of PA radical with the rate of ca. 1.35 Â 10 6 molecules cm À3 s À1 and 1.09 Â 10 6 molecules cm À3 s À1 , respectively. Although the photolysis rate of biacetyl is as fast as 4.4 Â 10 À5 s À1 , its contribution to the PA radical is very small due to its very low concentration in the atmosphere. In this study, considering the combination of systematic and random errors, we estimated the uncertainty (~15%) as much as a factor of 3 relative standard deviation in detecting these carbonyls (Pang and Mu, 2006), and calculated individual variations of acetaldehyde (17.91%), acetone (5.91%), MACR (2.64%), MGLY (2.13%), MVK (1.02%), and biacetyl (0.39%) in the total PA production rate. Therefore, the total uncertainty was estimated to be~30% by summing individual variations above in this study.

Major sink of PAN: thermal decomposition
Laboratory investigations have indicated that thermal decomposition is the predominant removal pathway of PAN in the daytime (Roberts and Bertman, 1992;Senum et al., 1986;Grosjean et al., 2001), because its reaction rate with OH radical (Wallington et al., 1984) and its photolysis are slow (Senum et al., 1984) as well as the low wet deposition rate (Gaffney et al., 1987). From the reactions (R2, R-2, and R3) illustrated in Section 3.3, the PAN lost by thermal decomposition (TPAN) is given by Equation (1).
(1) k À2 and k 3 are the reaction rate constants for reactions R-2 and R3, respectively.
[PAN], [NO], and ½NO 0 2 represent the concentrations. k 3 /k 2 of 1.95 ± 0.28 has been found to be independent of temperature from 10 C to 40 C under standard atmospheric pressure (Tuazon et al., 1991), and thus the Equation (1) could be simplified as Equation (2) by replacing k 3 with k À2 .
Because the rate constant k À2 is temperature dependent, k À2 (s À1 ) ¼ 2.52 Â 10 16 e (À13,573/T) (Tuazon et al., 1991), the thermal decomposition rate of atmospheric PAN depends both on temperature and the ratio of ½NO 0 2 =½NO. The amount of PAN lost by thermal decomposition (designated as TPAN) could be integrated over 1 h from the measured PAN concentration and the decomposition rate of PAN derived from Equation (2) at each temperature and the ½NO 0 2 =½NO ratio encountered during field measurement.
Positive correlation (Slope{(TPAN/PAN)/ T} ¼ 0.080 ± 0.006, Intercept ¼ À23.7 ± 1.92, R ¼ 0.95, SD ¼ 0.176, P < 0.0001) between TPAN/PAN and air temperature (T) was found in the temperature range of 293e313 K in this study. Fig. 7 illustrates the diurnal variations of TPAN and PAN on 24 June, 5 July, 12 August, and 4 September as the typical clear days in each month. As clearly seen in Fig. 7B, TPAN greatly exceeded the observed PAN from morning to late afternoon owing to the extremely high air temperatures (above 35 C) throughout the day. The relatively high TPAN, the abnormal trend (PAN concentrations decreased from morning to noontime), and the extremely low values of PAN (less than 0.6 ppbv) on the day in July indicated that the newly formed PAN via photochemical reactions might be totally consumed by the thermal decomposition. Because the air temperature on the days in June, August, and September in the morning (all lower than 30 C) was much less than that on the day in July, the TPAN accounted for smaller fraction of the total (PAN þ TPAN) than that on the day in July, and accumulation of atmospheric PAN formed via photochemical reactions from morning to afternoon was observed. In the late afternoon, photochemical reactivity decreased (mainly due to decrease of OH radicals), while thermal decomposition of PAN still kept high rate, atmospheric PAN decreased accordingly. Although the photochemical reactivity (implying PAN formation potential) in September became much less than that during summertime (from June to August), the peak values of PAN observed in September were comparable to (or even greater than) those observed in June and August (Fig. 7). The TPAN in each hour (with only the exception on 16:00) on the day in September was lower than PAN (Fig. 7D), whereas the TPAN after noontime in the summer months was always greater than PAN (Fig. 7AeC). It is evident that the relatively high PAN peak values observed in September were Table 1 Summary of the reactions for the production of PA radical with the observed and inferred concentrations and the rate constants used in the analysis.

Carbonyls
Concentrations ( a OH and a photolysis are the branching ratios for PA radical formation via the reaction with OH radical and the photolysis, respectively (Atkinson et al., 2006). a Data cited our previous work (Zhang et al., 2012a(Zhang et al., ,b, 2014. b Data inferred from (Shao et al., 2009). c OH concentration was assumed to be constant and estimated to be 5 Â 10 6 molecules cm À3 during summertime in Beijing (Liu et al., 2012). d Recommended by Atkinson et al. (2006) at 298 K. e Recommended by Darnall et al. (1979). f The maximum photolysis rates were calculated at noon on 12 August in Beijing based on the actinic fluxes (Finlayson-Pitts and Pitts, 1986) and the absorption cross sections of acetaldehyde (Martinez et al., 1992), MGLY (Meller et al., 1991), MACR (Gierczak et al., 1997), MVK (Gierczak et al., 1997), acetone (Meyrahn et al., 1986), and biacetyl (Horowitz et al., 2001). reasonably ascribed to the remarkable decrease of PAN consumed by thermal decomposition.

Conclusion
In our study, high pollution level of PAN and PPN were observed in Beijing during summertime. The results of correlation analysis between PAN and PPN indicated that AVOCs dominated the photochemical formation of PANs in Beijing. Further, acetaldehyde was found to be the predominant precursor of PAN with the contribution of 59.7% to the total peroxyacetyl (PA) radical. As expected, the amount of PAN lost by thermal decomposition (TPAN) accounted for remarkable fractions of PAN observed under high temperature during both daytime and nighttime.