Characteristics of atmospheric ammonia and its relationship with vehicle emissions in a megacity in China

Atmospheric Environment 182 (2018) 97–104

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Characteristics of atmospheric ammonia and its relationship with vehicle emissions in a megacity in China

T

Ruyu Wanga, Xingnan Yea,b,∗, Yuxuan Liua, Haowen Lia, Xin Yanga,b, Jianmin Chena,b,∗∗, Wei Gaoc, Zi Yina a

Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention, Department of Environmental Science and Engineering, Fudan University, Shanghai 200433, China b Institute of Atmospheric Sciences, Fudan University, Shanghai 200433, China c Shanghai Key Laboratory of Meteorology and Health, Shanghai 200030, China

G R A P H I C A L A B S T R A C T

A R T I C LE I N FO

A B S T R A C T

Keywords: Ammonia Seasonal variability Traffic emissions Shanghai

Atmospheric ammonia plays an important role in haze formation in East China. In this study, long-term measurements of NH3 concentrations were implemented at urban, suburban, and tunnel sites in Shanghai, the largest city in East China. The average monthly ammonia concentrations at the urban site varied from 3.7 ppb to 14.5 ppb and exhibited the highest levels in summer and lowest levels in winter, indicating that the biological emissions and agriculture in the surrounding areas are important contributors. The suburban NH3 levels were significantly higher in autumn compared to those at the urban site, indicating the important contribution of agricultural activities. Regardless of the season, the difference of NH3 concentrations between the tunnel and urban sites remained almost constant. On average, the tunnel NH3 level was three times higher than that of the nearby urban site, indicating strong vehicle NH3 emissions in the tunnel. The tunnel NH3 levels on weekdays were comparable to those on weekends, a result that was in agreement with the daily average traffic volume. It was estimated that the vehicle emissions contributed 12.6–24.6% of the atmospheric NH3 in the urban area and 3.8–7.5% for the whole area of Shanghai. Our results suggest that vehicle NH3 emissions should be considered, although agricultural emissions are still more important for mitigating severe haze pollution during wintertime in the megacities of China.

∗ Corresponding author. Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention, Department of Environmental Science and Engineering, Fudan University, Shanghai 200433, China. ∗∗ Corresponding author. Shanghai Key Laboratory of Atmospheric Particle Pollution and Prevention, Department of Environmental Science and Engineering, Fudan University, Shanghai 200433, China. E-mail addresses: [email protected] (X. Ye), [email protected] (J. Chen).

https://doi.org/10.1016/j.atmosenv.2018.03.047 Received 30 November 2017; Received in revised form 16 February 2018; Accepted 20 March 2018 Available online 22 March 2018 1352-2310/ © 2018 Elsevier Ltd. All rights reserved.

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1. Introduction

in Beijing during 2013 was expected to be comparable to agricultural emissions (Pan et al., 2016). Despite these results, the contribution of traffic sources to urban NH3 concentrations is still an area of debate. For example, Reche et al. (2012) believed that the urban design of the streets should be the determining factor influencing urban NH3 concentrations. Similarly, Teng et al. (2017) suggested that emissions from local green space inside the urban areas and evaporation of ammonia and ammonium in dew droplets were major sources of atmospheric NH3 in urban area of Qingdao for that the temporal variation of NH3 concentrations was synchronized with them. In past decades, China suffered from severe fine particulate pollution. Existing in the main forms of (NH4)2SO4 and NH4NO3, the secondary sulfate, nitrate, and ammonium aerosols accounted for a large fraction of PM2.5 mass in all haze episodes (Huang et al., 2012a, 2014). It was established that the increase of atmospheric NH3 concentrations might play a vital role in the enhancement of fine particulate nitrate and PM2.5 levels (Chu et al., 2016; Wen et al., 2015; Ye et al., 2011a). In a recent study, Wang et al. (2016) revealed that the ammonia-rich atmosphere was a crucial factor for the periodic occurrence of PM2.5 episodes in Beijing. Moreover, Fu et al. (2017) emphasized that the effort of SO2 and NOx emissions reduction on mitigating haze pollution in East China was partly offset by the increasing ammonia concentrations. Wu et al. (2016) highlighted that reduction of NH3 emissions should be more efficient than other particle precursors at mitigating PM2.5 pollution in China. With increasing concern over PM2.5 pollution, it is crucial to understand the concentrations and sources of urban NH3. However, ammonia has not been regulated in the national ambient air quality standards of China (GB 3095–2012), and few observations on the urban NH3 concentrations were reported. With a population of 24 million and a total land area of 6400 km2, Shanghai is the biggest megacity and the economic center of the Yangtze River Delta. Chang et al. (2016) estimated that vehicle emissions accounted for 12% of the total NH3 emissions in the urban area of Shanghai. Wang et al. (2015) found strong positive correlations between the conversion rate of ammonia and particulate sulfate and nitrate mass over the city. In this study, long-term measurements of ambient NH3 were performed at urban, suburban, and tunnel sites in Shanghai. The objective of this study is thus threefold: (1) to characterize seasonal variability of NH3 concentration in the city, (2) to identify whether vehicle emissions is an important source of urban NH3, and (3) to assess the contribution of vehicle emissions on NH3 budget in urban area and the whole area of Shanghai. These data will help researchers gain a better understanding of ambient NH3 levels and the importance of vehicle emissions in this area.

Ammonia (NH3) is the dominant alkaline gas and the third most abundant nitrogen-containing species after nitrogen gas and nitrous oxide in the atmosphere. Classical nucleation theories and laboratory studies demonstrate that the presence of NH3 at ppt levels significantly enhances the nucleation rates in the H2SO4-H2O system (Zhang et al., 2012), indicating that NH3 is a critical factor in the nucleation of new particles. The important role of NH3 in new particle formation events has also been established by field observations. Smith et al. (2005) found that ammonium and sulfate were the only constituents of the 6–15-nm particles during nucleation events in Atlanta. Yue et al. (2010) reported that the neutralization of H2SO4 by NH3 contributed approximately 50% to the initial growth of freshly formed particles. In the atmosphere, ammonia neutralizes readily with acidic species from SO2 and NOx precursors and has long been recognized as an important contributor to the formation of secondary sulfate and nitrate aerosols, which are major components of PM2.5 around the world (Seinfeld and Pandis, 2006). Moreover, the heterogeneous oxidation of SO2 is favored in the ammonia-rich atmosphere, promoting persistent sulfate formation (Wang et al., 2016). The increase of secondary inorganic aerosols and PM2.5 mass not only degrades atmospheric visibility but also imposes strong impacts on human health (Pui et al., 2014). In addition, both wet and dry deposition of NH3 and particulate ammonium salts have many effects on the ecosystem, accounting for the soil acidification and water eutrophication (Behera et al., 2013). Agriculture is undoubtedly the dominant NH3 source on the global scale, contributing over 80% of global NH3 emissions (Behera et al., 2013). In contrast, the contribution of on-road transportation to the global NH3 is negligible (∼1% of the total contribution). In China, the total NH3 emissions was nearly one million tons in 2012, with the largest contribution being from livestock (52% of contribution) and the second largest source being from applied fertilizer (29%). The traffic sector accounted for 4% of the total emissions, followed by biomass burning (3%) and the chemical industry (3%) (Kang et al., 2016). Due to the strong emissions from livestock and applied fertilizer, higher concentrations of atmospheric NH3 were reported in rural areas (Shen et al., 2011; Xu et al., 2016). However, many studies have found that ambient NH3 concentrations in many urban areas were comparable to or even higher than those in the nearby rural areas, indicating that nonagricultural emissions may contribute greatly to urban NH3 sources (Cao et al., 2009; Meng et al., 2011). The estimation of non-agricultural NH3 emissions in urban areas is still challenging, since biomass burning, domestic sewage, garbage collection systems, road traffic, industrial emissions, and other factors are contributors of local NH3 emissions and the relative emission strengths of individual NH3 sources remain unclear (Behera et al., 2013). For example, Zhang et al. (2017) reported that sewage treatment accounted for ∼4% of the ammonia emissions in the urban area of the Pearl River Delta (PRD), whereas the estimated contribution was as high as 34% based on the NH3 emission inventories compiled by Zheng et al. (2012). Since the introduction of three-way catalytic converters (TWCs), gasoline vehicles generate more NH3 by over-reducing NOx in vehicular exhaust (Suarez-Bertoa and Astorga, 2016). The adoption of selective catalytic reduction (SCR) by the addition of urea or NH3 to diesel exhaust to reduce NOx emissions is another potential NH3 source from road traffic. With the extensive equipment of TWCs or SCR, road traffic is expected to contribute more NH3 to the urban atmosphere. Fraser and Cass (1998) reported that the contribution of vehicle sources increased from 2% to 15% of the total NH3 emissions since the installment of TWCs in Los Angeles. Similarly, vehicle emissions have become a significant contributor of atmospheric NH3 in Houston (Zhang and Ying, 2010). In the United Kingdom, the transport sector was estimated to be the primary source of non-agricultural NH3 emissions, followed by sewage emissions (Sutton et al., 2000). With the development of urbanization, the emission of NH3 from traffic sources

2. Experimental 2.1. Sampling sites Ambient NH3 samples were collected using Ogawa passive samplers (Ogawa USA, Pompano Beach, Florida) at an urban site, a suburban site, and a tunnel site in Shanghai. The sampling locations are shown in Fig. 1. The urban site was located on the main campus of Fudan University (31.3°N, 121.5°E). The campus can be considered as a representative urban site for Shanghai due to many dwelling quarters and commercial blocks in the surrounding area. The tunnel NH3 samples were collected over the open section of the Handan Road Tunnel, which is at a distance of approximately 300 m from the urban site. To improve ventilation effects, the 760 m-long tunnel consists of two buried sections at both ends and a 200 m-long open section (fencelike area in the map) at the central tunnel. Thus, the tunnel samples may reflect NH3 emissions from on-road vehicle sources. The suburban site was located in the town of Zhoupu, which is approximately 30 km to the southeast of the main campus of Fudan University. There are stretches of residential houses and farmlands near this suburban site. The main crops in these farmlands are rice, vegetables, and more. 98

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Fig. 1. Locations for sampling ambient ammonia over Shanghai. (●) urban site; (◆) suburban site.

where WNH3 represents the collected amounts of NH3 (ng) during the exposure time t (min). α represents the temperature-dependent concentration conversion coefficient and is equal to 43.8 ppb min/ng at 25 °C (α = 43.8 × (293/(293 + t/∘C ))1.83 ).

2.2. NH3 passive sampling Passive sampling is widely used to monitor ambient NH3 concentrations, and the Ogawa passive sampler has been validated previously (Meng et al., 2011). The Ogawa sampler consists of a solid Teflon cylinder with two open but unconnected chambers. Each chamber of the sampler is assembled with a diffuser end-cap, two stainless steel screens, and a Teflon disk. The sampling and analysis procedures were performed following the manufacturer's protocols (http://www.ogawausa.com). The exposure and collection time recommended by the manufacturer is 1–7 days. All sampler components were carefully rinsed with pure water and dried before each use. A citric acid-impregnated collection pad (14 mm in diameter) was mounted between the inner and outer stainless steel screens in the chamber to convert the diffusive NH3 into ammonium citrate. After exposure, the samples were sealed in glass vials and stored at −18 °C for chemical analysis. A weekly sampling period was implemented at the suburban site. Both five-day weekday and two-day weekend sampling periods were implemented at the urban and tunnel sites during the sampling in 2014. A two-day sampling period was implemented at the urban site during the sampling in August 2011 and in November and December 2012.

3. Results and discussion Fig. 2 shows the monthly distribution of NH3 concentrations at the urban site. The monthly average NH3 concentration varied from the maximum of 14.5 ppb in August to the minimum of 3.7 ppb in December, with the annual mean value of 6.4 ppb. From a seasonal perspective, the average NH3 concentrations were 6.3 ppb, 10.5 ppb, 5.6 ppb, and 3.7 ppb in spring, summer, autumn, and winter, respectively. In contrast to the NH3 concentrations in spring, which were comparable to those in autumn, the average NH3 concentration in summer months was approximately three times higher than that in winter. Shi et al. (2014) conducted online measurements with a Monitor for AeRosols and Gases (MARGA) at the urban site in autumn of 2012 and found that the average NH3 concentration was 6.6 μg m−3,

2.3. Chemical analysis The NH3 collection pad was put into a 10-ml vial and extracted with 8 ml of deionized water (18 MΩ cm). The ammonium concentration in the collection extract was determined by an Ion Chromatograph (Metrohm 883 basic IC plus, Switzerland) equipped with a Metrosep C4-100/4.0 column. The eluent was 1.0 mmol l−1 HNO3 + 1.0 mmol l−1 2,6-pyridine dicarboxylic acid. The detection limit for NH4+ was 2.8 μg l−1, corresponding to the ambient NH3 concentration of 0.25 ppb for a two-day sample. Three blanks were analyzed in the same procedures for each campaign to reduce the system error. The ambient NH3 (in ppb) was calculated according to the following equation:

[NH3] = α ×

WNH3 t

Fig. 2. The monthly variation of atmospheric ammonia concentrations in the urban site.

(1) 99

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Fig. 4. Temporal variations of atmospheric ammonia concentration in the suburban site. Fig. 3. The correlation between the concentration of atmospheric ammonia and ambient temperature in the urban site.

ammonium increased from 2.5 to 7.8 μg m−3 (Tao et al., 2016). It should be noted that the temperature-dependent feature was insignificant for any season in this study. The most plausible explanation is that long-range transport of ammonia, atmospheric diffusion conditions, rainfall, and other meteorological factors exert stronger impacts on daily average NH3 concentration than temperatures in the same season. Fig. 4 shows the temporal variation of NH3 concentrations at the suburban site. The concentrations of NH3 were approximately 10 ppb in September and October but varied in the range of 4–7 ppb from November to January. As described in the experimental section, the suburban site was surrounded by a large amount of farmland. The high levels of atmospheric NH3 in autumn indicated the importance of agricultural activities (e.g. rice cultivation) as an NH3 source over suburban areas. The concentration of NH3 declined sharply in November, although December was generally regarded as the beginning of winter in Shanghai. This feature was mostly attributed to the decrease in agricultural activities during wintertime. The decrease of agricultural NH3 concentration during cold seasons was also reported previously (Xu et al., 2016). The suburban NH3 level in Shanghai was lower than the rural NH3 levels in northern China, due to different agricultural structures (e.g. the main crop is rice in southern China whereas they are wheat and maize in northern China) and meteorological conditions (Cao et al., 2009; Meng et al., 2011; Wen et al., 2015; Xu et al., 2016). Notably, the autumn NH3 levels at the suburban site were significantly higher compared to the urban site, whereas they were at the same levels during cold months. Table 1 compares NH3 levels between urban and suburban/rural sites in different regions. The results from active DOAS and MARGA measurements showed that the annual mean NH3 at the urban site was significant lower than that at the rural site (Wang et al., 2015), which was similar to our results during autumn and winter. The mean NH3 concentration in July exceeded 30 ppb at the rural site, which significantly increased the annual average NH3 concentration in their study. The phenomenon of NH3 concentration at rural sites remaining considerably higher than that at urban sites, was also found in the IndoGangetic plain and in Georgia (Saylor et al., 2010; Singh and Kulshrestha, 2014). The high NH3 level at the rural site on the IndoGangetic plain was attributed to fertilizer application and biomass burning, while poultry husbandry was considered an important NH3 source at the rural site in Georgia. In contrast, the concentrations of NH3 at urban sites were reported to be clearly higher than those at rural sites in Beijing and Rome (Meng et al., 2011; Perrino et al., 2002). The different trends between NH3 levels at the urban and rural sites can be attributed to the difference in NH3 sources and their emission strength. To explain the higher NH3 concentrations at urban sites, traffic emissions were considered as an important contributor of atmospheric NH3

which is highly consistent with our results. The mean annual NH3 concentration at the urban site was also comparable to the observations by method of active Differential Optical Absorption Spectroscopy (DOAS) from 2013 to 2014 (Wang et al., 2015). The difference is that the winter NH3 level was relatively high in the study by Wang et al. (2015), possibly due to frequent occurrence of pollution episodes in the winter of 2013. The urban NH3 concentrations observed in this study were lower than those determined by the same passive sampling method in Beijing and Xi'an, two megacities in northern China (Cao et al., 2009; Meng et al., 2011). This feature is consistent with the fact that the acid rain zones are mainly located to the south along the Yangtze River (Hao et al., 2001). However, the NH3 level in Shanghai was considerably higher than those observed in some other cities such as Madrid (Reche et al., 2015) and Houston (Gong et al., 2011), indicating high NH3 level in China compared to light-polluted area in Europe and North America. The characteristics of seasonal variation is an indicator of NH3 sources and sinks because natural emissions increase under warmer conditions, whereas industrial and traffic emissions are insensitive to temperature. Fig. 3 illustrates the variation of NH3 concentration as a function of ambient temperature. The temperature dataset was obtained from the Hongqiao Airport meteorological site in Shanghai (http://www.wunderground.com). Generally, the concentration of atmospheric NH3 increased exponentially with ambient temperature (R2 = 0.57, p < 0.001), indicating that temperature was a key parameter influencing ambient NH3 concentrations in the urban area of Shanghai. With frequent agricultural activities and fertilizer application, ammonia emissions in summer are expected to increase significantly in surrounding agricultural areas. On the other hand, microbial activities and the volatilization of NH3 from domestic garbage, city sewage, and vegetation are favored under warm conditions. Considering the short lifetime of atmospheric NH3, the presence of the highest NH3 level in summer was mostly attributable to biological sources and regional agricultural activities. Similar temperature-dependent trends were also found in previous studies (Xu et al., 2016; You et al., 2014). In contrast, microbial activities are expected to be suppressed under cold months, resulting in the decrease of NH3 emissions from biological sources. Moreover, the formation of secondary aerosol NH4NO3 is highly sensitive to temperature because of its low thermodynamic stability. The gas-to-solid phase conversion is favored under cold conditions in contrast to high evaporation of NH4NO3 in summer. The effect of nitrate formation on the concentration of NH3 was supported by the fact that the fine particle nitrate increased from 0.7 μg m−3 in summer to 13.9 μg m−3 in winter, while the fine particle 100

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Table 1 Comparison of atmospheric NH3 concentrations between urban and suburban/rural areas in different regions. Location

Period

NH3 concentration (ppb) Urban

Suburban/ Rural

Reference

Shanghai, China Shanghai, China Xi'an, China

2014.08–2015.01

3.7–5.6

4.8–8.2

This study

2013–2014

6.2

12.4

2006.04–2007.04

18.6

20.3

Beijing, China Georgia, U.S.

2008.02–2010.07(urban) 2007.01–2010.07(rural) 2007.07–2007.12

22.8

10.2

1.35

3.32

Rome, Italy

2001.05–2002.3

5.3

3.5

Toronto, Canada IndoGangetic plain, India

2003.07–2011.09(urban) 2006.06–2007.03(rural) 2012.10–2013.09

2.3–3.0

0–4

52.8

65.5

(Wang et al., 2015) (Cao et al., 2009) (Meng et al., 2011) Saylor et al. (2010) Perrino et al. (2002) (Hu et al., 2014) Singh and Kulshrestha (2014)

in previous studies. The monthly variation of tunnel NH3 concentrations is shown in Fig. 5a. The monthly average concentration of tunnel NH3 varied in the narrow range of 15.3–20.0 ppb, with an average of 17.4 ± 3.9 ppb. The tunnel's NH3 level in this study was comparable to the Parque do Povo tunnel in the urban area of São Paulo (Vieira et al., 2016) The considerable higher NH3 concentration at the well-ventilated tunnel compared to the urban site indicated the importance of vehicle emissions as a strong NH3 source. Recently, the application of 15N-Stable isotope technology provided strong support for the contribution of vehicle emissions to atmospheric NH3 during hazy days in the urban areas of Beijing (Pan et al., 2016). However, the tunnel NH3 level measured in this study was significantly lower compared to the Zhujiang tunnel in Guangzhou (Liu et al., 2014). The possible explanation is that the tunnel was well-ventilated in this study whereas the ventilation fans in the Zhujiang tunnel were turned off in their study. Unlike seasonal variations at the urban site, the tunnel NH3 level remained relative stable. To explain the characteristics of tunnel NH3 concentration, a simplified reactor model is considered here. Assuming a well-mixed condition in the tunnel, the flux of NH3 should satisfy the following equation:

Q× Cambient + S− Q× Ctunnel = V×

dCtunnel dt

Fig. 5. The monthly variation of atmospheric ammonia concentrations in the tunnel site (a), the difference and ratio of ammonia concentration between tunnel and urban sites (b).

11.8 ± 1.3 ppb. The small standard deviation of 1.3 ppb matched well with the reactor model, supporting our assumption that the tunnel was well-ventilated. The standard deviation increased from 1.3 ppb of the difference to 3.9 ppb of the tunnel NH3 concentration, indicating the influence of ambient air on the tunnel NH3 concentration. The tunnel NH3 was relatively stable because the vehicle emissions contribute greatly to NH3 budget in the tunnel compared to ambient air. The ratio of NH3 concentrations between the tunnel and urban sites (RT/U) varied in the range of 2.45–4.59, with an average of 3.23. The ratio was absent in summer, because the tunnel NH3 was not measured during this season. Vieira et al. (2016) found that the concentrations of NH3 inside the tunnel were also three times higher than those outside in the urban area of São Paulo. Löflund et al. (2002) reported a three-fold increase in the NH3 concentration in a tunnel in Australia after the application of the three-way catalysts. In addition, Perrino et al. (2002) reported that the NH3 concentrations at the traffic sites were about five times the background level. Tanner (2009) reported that the roadside NH3 levels were seven times higher than that of residential air. The significantly elevated levels of tunnel NH3 compared to the atmosphere outside provided further support for the importance of vehicle emissions as an NH3 source. Fig. 6 shows the box plots of NH3 concentrations on weekdays and weekends at the tunnel site in different seasons. The median NH3

(2)

where, Cambient, Ctunnel, Q, S, and V represent the concentration of the urban ambient NH3, the concentration of the tunnel NH3, the flow rate through the tunnel, the NH3 source from vehicles, and the volume of the tunnel, respectively. In a steady state, the difference between the tunnel and ambient NH3 concentrations is ruled by the following equation:

Ctunnel − Cambient =

S Q

(3)

Therefore, the tunnel NH3 concentration may vary with the ambient NH3 level whereas the difference between the tunnel and ambient NH3 concentrations should remain relatively constant if the traffic density and the flow rate through the tunnel do not vary significantly. The difference between tunnel and urban NH3 concentrations and the ratio of tunnel and urban NH3 concentrations are shown in Fig. 5b. The difference between monthly tunnel and urban NH3 concentrations varied in the narrow range of 10.3–13.9 ppb, with an average of 101

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Table 2 The contributions of vehicle emissions to atmospheric NH3 in Shanghai and other regions. Year

Location

w%

References

2014

Shanghai, China

This study

2014 2006 1993 2006 2012 2006

Shanghai, China Shanghai, China Los Angeles, USA PRD, China PRD, China Beijing, China

12.6–24.6% (urban) 3.8–7.5% (regional) 12% (urban) 4.7% (regional) 15% (regional) 2.5% (regional) 18.8% (regional) 5.2% (regional)

Chang et al. (2016) Huang et al. (2012b) Fraser and Cass (1998) Zheng et al. (2012) Liu et al. (2014) Huang et al. (2012b)

trend on weekends. The traffic density increased slowly on weekend mornings but remained elevated for a longer time and maintained the maximum flow from 11:00 to 18:00. It is notable that the daily average traffic volume was comparable (84270 vehicles per day for weekends, 84343 vehicles per day for weekdays), which provided a reasonable explanation for the almost identical NH3 levels between weekdays and weekends. Table 2 compares vehicle contributions to atmospheric NH3 in Shanghai and other regions. As described in the experimental section, the Handan tunnel contains a 200 m-long open section. This special design provided a unique opportunity to determine the NH3 emissions from vehicle sources. In this study, we estimated the contribution of onroad traffic to atmospheric NH3 in the urban area of Shanghai by comparing the difference between the NH3 concentration above the open section of the tunnel and the concentration at the urban site:

Fig. 6. Box plots showing tunnel NH3 concentrations for weekdays and weekends. The whiskers represent the outliers, the two borders of box display the 25th and 75th percentile, and the band in each box denotes the median.

w% =

(Ctunnel − Cambient ) S × road × 100% Cambient Surban

(4)

where w% is the contribution of vehicle emissions to the urban ammonia. Sroad and Surban are the total roadway area and the urban area, respectively. The latest statistics showed that the urban area and the total roadway area in the urban area reached 998 and 74.1 km2, respectively (China Urban Rural Statistics Yearbook, 2015; Shanghai Comprehensive Transportation Yearbook, 2015). At seasonal average, the contribution of vehicle emission to ambient NH3 in the urban area was estimated to be 12.6%, 15.3%, and 24.6% in spring, autumn, and winter, respectively. These results highlighted the significant contribution of vehicle emissions to the atmospheric NH3 in the urban area, particularly during wintertime. Wang et al. (2016) reported that both NH3 and NOx were important factors influencing sulfate formation on deliquesced fine particles. Ye et al. (2011b) reported that most accumulation mode particles were deliquesced under haze conditions in winter. The high contribution of vehicle NH3 suggests that the reduction of traffic emissions may be one effective measure to mitigate haze pollution during wintertime in China. Considering that the ratio of roadways and land area for the whole region is 70% less than that of the urban area, the contributions of vehicle NH3 decreased to 3.8–7.5% when the gross land and roadway areas of Shanghai were input. The estimated contributions of vehicle emissions to the atmospheric NH3 in the urban area in this study is comparable to the previous study by Chang et al. (2016). The contribution of vehicle emissions to the atmospheric NH3 in the whole area of Shanghai is also consistent with the comprehensive NH3 emission inventory (Huang et al., 2012b). Although the vehicle source is an important contributor on the regional NH3 budget, the other sources accounting for approximately 92% of total atmospheric NH3 during wintertime dominates the formation of NH4NO3. It is crucial to accurately quantify the contributions of local emissions and long-range transport of NH3, agricultural and non-agricultural sources. The contribution of vehicle emissions on the NH3 budget in the area of Shanghai is considerably lower than that in the PRD region reported by Liu et al. (2014), but higher than that reported by Zheng et al. (2012). It is notable that the contribution of vehicle ammonia in the PRD region is very different throughout the literature

Fig. 7. The diurnal variations of hourly traffic density at nearby the tunnel site.

concentrations on weekdays were 16.6 ppb, 16.1 ppb, and 16.1 ppb in spring, autumn, and winter, respectively. Considering that the biological sources are highly temperature-dependent, the almost identical NH3 levels in different seasons further supported vehicle emissions as the dominant contributor of NH3 in the tunnel. The median NH3 concentrations on weekends were 16.7 ppb in spring, 16.3 ppb in autumn, and 18.1 ppb in winter. It is notable that there were not significant differences between weekday and weekend NH3 concentrations in spring and autumn and that the weekend concentrations were somewhat higher than those on weekdays in winter. Löflund et al. (2002) reported that the concentration of tunnel NH3 was higher on Sunday than that on Monday and attributed it to the higher density of gasoline vehicles on weekends. Pandolfi et al. (2012) found a slight decrease of NH3 concentrations on weekends, consistent with the decrease in other traffic-related pollutants. One of possible explanation for the higher weekend NH3 concentration during wintertime is that there were more private gasoline cars on road. To explain the similar NH3 concentrations observed in this study, the diurnal variations of traffic density in the tunnel on weekdays and weekends are illustrated in Fig. 7. The statistics on traffic volume in 2015 were used in this study because the dataset of 2014 was not available. On weekdays, the diurnal variation of traffic volume displayed a bimodal distribution, with two peaks in the morning and evening rush hours and a significant decrease during business time. The traffic density distribution displayed a different

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because the emission factor of vehicle NH3 determined by Liu et al. (2014) is 230 mg km−1 in contrast to the maximum of 63.2 mg km−1 used by Zheng et al. (2012). The large uncertainty of ammonia emission factor indicates that more studies should be considered in the future. The contribution of vehicle NH3 in the area around Los Angeles estimated with an emission factor of 61 mg km−1 was also higher than that of Shanghai (Fraser and Cass, 1998), indicating that much more attention should be paid to the contribution of vehicle NH3 in Chinese megacities undergoing motorization.

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