1. Introduction
China’s rapid economic growth and urbanization greatly increased energy consumption and air pollutant emissions since economic reforms in 1978 [
1]. Environmental monitoring campaigns carried out in urban centers highlighted O
3 values that are of concern for human health [
2,
3]. Elevated surface ozone (O
3) is a growing environmental concern in China [
1,
2,
3,
4,
5]. Serious O
3 episodes are frequent in China’s most economically vibrant and densely populated regions, such as the Beijing–Tianjin–Hebei (BTH) area and the Yangtze River Delta (YRD) and Pearl River Delta (PRD) regions [
5]. Studies showed substantial detrimental effects of elevated ground-level O
3, including declining forest ecosystem services, crop yield loss, and associated premature deaths [
4,
6,
7].
Tropospheric O
3 is formed by a series of complex nonlinear photochemical reactions between O
3 precursors, namely methane, carbon monoxide, volatile organic compounds (VOCs), and nitrogen oxides (NO
x = NO + NO
2) [
8]. Because NO
x is mainly anthropogenic, control of O
3 is usually through NO
x regulation [
9]. In China’s 12th five-year plan (2011–2015), the goal was set to reduce national NO
x emissions by 10% relative to the 2010 level [
10]. Both satellite retrievals and emission inventories show success in reducing NO
x emissions nationwide [
11,
12,
13,
14,
15]. Compared with other large NO
x emission regions such as BTH and YRD, the decline in NO
x emissions in the PRD region started earlier in 2005 [
11,
14].
Shenzhen is a major megacity in the PRD region and a twin megalopolis to Hong Kong. Shenzhen was established as China’s first Special Economic Zone in 1980 and developed from a small village into a large metropolitan area with a population over 10 million in three decades. Accompanied by rapid economic growth, the air quality in Shenzhen deteriorated [
16]. Epidemiologic studies showed significant associations between O
3 levels and increased mortality and respiratory diseases in Shenzhen [
17,
18]. To tackle air pollution issues, the governments of Guangdong Province and Hong Kong worked closely to reduce emissions of SO
2, NO
x, and VOCs throughout the region [
19]. In addition to national and regional environmental control efforts, Shenzhen also implements energy reforms for public transportation and private motor vehicles. It is the world’s first city that electrified 100% of public buses and will electrify all taxis by the end of 2020 [
20]. The decline in NO
x emissions is fastest in Shenzhen compared with other megacities like Shanghai, Beijing, and Guangzhou [
11].
NO
x regulation showed effectiveness in reducing peak O
3 levels over the long term and elevated mean O
3 levels at the same time [
21,
22,
23,
24]. However, due to the nonlinear relationship between NO
x and O
3, NO
x control might worsen the O
3 problems in urban areas in the short term [
19,
25]. This is mainly caused by the lessened titration effect of NO in urban areas, where the O
3 photochemical chemistry is mainly VOC-sensitive [
21,
22,
23,
24]. Simulation results showed that peak O
3 levels could be effectively reduced in the PRD region when cutting down NO
x and VOC emissions [
15,
19,
25]. Recent observations showed that annual mean O
3 levels were increasing at both urban and suburban sites in the PRD region [
24]. However, the changes in peak O
3, which is closely related to human health, are rarely shown by observation data.
Apart from emissions of O
3 precursors, meteorological factors are also important in affecting O
3 levels both directly and indirectly. Meteorological factors can directly impact the formation of O
3 because the reactions are sensitive to the changes in sunlight and temperature [
5,
26,
27]. Meteorological factors such as wind speed, surface pressure, and precipitation also impact the accumulation, dilution, and deposition of O
3 [
5,
27,
28,
29]. The prevailing wind direction that changes with the season is closely related to air pollutant transport from near and distant areas [
5,
27,
28]. Actually, the relationship between air pollutants and meteorological factors varies greatly by geographical location and season [
5]. The investigation of the relationship between O
3 and meteorological parameters in Shenzhen will help understand the pollution situation in the PRD region, as well the large cities that are close to Shenzhen, namely Hong Kong and Macau.
The characteristics of O
3 at urban, suburban, and rural sites depend on the O
3 photochemical sensitivity of the region. One previous study reported higher 1-h and 8-h O
3 in central areas of Shenzhen, which is uncommon for most large cities [
17]. Higher urban O
3 is a possible pattern for some cities, e.g., Atlanta and Houston, where large sources of VOCs originate from biological and anthropogenic emissions [
30,
31]. For most populous areas in the PRD region, the photochemical regime was indicated as VOC-sensitive, with a mixed or transitional regime in Shenzhen [
32]. Studying the spatial O
3 pattern between urban and rural areas will help determine the controlling precursors related to the O
3 production rate, which is critical for O
3 control and management.
The comprehensive monitoring of O3 and NOx by the Shenzhen Meteorological Bureau was carried out at four air monitoring sites, starting in 2010. This is three years earlier than the introduction of CAAQS GB 3095–2012 in Shenzhen in 2013, allowing us to analyze the spatial and interannual variations of air pollutants. In this study, we investigate the characteristics and trends of O3 at urban, suburban, and rural sites of Shenzhen based on seven years of observations of O3 and NOx from 2011 to 2017. Our objectives were to (1) assess and compare O3 risks to human and plants at urban, suburban, and rural sites; (2) investigate weekly and monthly O3 and NOx variations and their influences; and (3) evaluate O3 and NOx trends during the 2011–2017 period.
4. Discussion
The common autumn (October) O
3 maxima and summer (June) O
3 minima at all representative sites are consistent with seasonal characteristics of O
3 air concentrations in subtropical Asia, including western India [
43], central and southwestern Taiwan [
44,
45], the PRD region [
17,
46], Hong Kong [
28], and the South China Sea area [
47]. All those areas are geographically close to each other and have similar atmospheric background conditions. Solar radiation, precipitation, and relative humidity are the most significant impact meteorological factors that affect O
3 daily variations. The formation of O
3 is favored by sunlight and high temperature [
5,
26,
27,
48]. The strong positive correlation in the winter between O
3 and both of these meteorological factors is likely due to the cold season, which makes them limiting factors for O
3 production. Surface pressure showed the opposite relationship in the summer and other seasons. O
3 production favors high-pressure systems with the subsidence of air [
28,
47]. This explains the observed positive relationship between O
3 and surface pressure in all other seasons except summer. This negative relationship in the summer was probably controlled by a synoptic-scale feature with the co-product of the negative correlation between temperature and pressure [
48]. The negative correlation between O
3 and relative humidity and precipitation explains the summer O
3 trough. With an increase in relative humidity, the cloud abundance usually increases at the same time, and the photochemical processes of O
3 production will decrease [
27]. Rainy weather is also not conducive for O
3 production, and heavy rain will remove O
3 from the air [
46].
In contrast, wind speed is the least relevant factor and is only negatively related to O
3 levels in the winter. Higher wind speed will increase the turbulence of the air and cause dispersion of O
3 [
28]. The negative relationship between O
3 and wind speed in winter might be influenced by the wind direction in winter which is an important factor that affects O
3 transport [
5]. Wind direction is found more important for influencing air pollutant levels. Shenzhen has at least two different sources of air pollution. One is a more regional emission source from the west and is active throughout the year. Emissions from the shipping business on the Pearl River Estuary and industrial activities in Guangzhou and Foshan to the west of Shenzhen might contribute to the continuous regional sources [
49]. The high concentrations associated with the northeasterly wind in the winter confirms that the impact of air pollutant transport from northern China [
5]. Interestingly, there are no substantial air pollutants associated with the southerly winds in all seasons, which means that contribution of air pollutants from big cities to the south of Shenzhen (i.e., Hong Kong) are less important in impacting air pollutants in Shenzhen.
The difference in NO
x air concentrations is significant between weekdays and the weekend at the urban site. This is probably caused by a weekly routine of human activities. Traffic volume is increased in the city center during workdays. Similar to a previous study result for the PRD region, we also found no statistically significant difference between weekdays and the weekend at all sites in Shenzhen [
46]. However, the O
3 weekend effect was significant (
p = 0.062) at the urban site, with O
3 levels being 1.19 ppb higher on Sunday than on weekdays. O
3 is elevated mainly by less on-road vehicular emissions on the weekend with NO
x reduction and a lowered O
3 titration effect [
24,
46,
50]. In our study, we found the difference in O
3 and NO
x air concentrations between weekday and Saturday were less than those between Saturday and Sunday. A carryover of heavy Friday evening emissions is likely explained the diminished weekday and Saturday O
3 difference [
50]. Overtime work on Saturday might also be an important factor. Individual mobility studies show that a Saturday morning peak of passengers is significant, indicating a large proportion of citizens working on that day [
51]. Moreover, heavy-duty trucks that contribute to strong emissions of NO
x are more active on Saturday because of fewer traffic restrictions [
52].
Different from a previous study that showed that central areas of Shenzhen had higher 1-h and 8-h O
3 [
17], we showed results similar to most large cities whereby the urban site has lower O
3 compared with other sites [
18,
23,
26]. The reliability of the previous study is questionable because only one-year observations were used and no statistical comparison was performed. Shenzhen had a lower maximum 1-h, 8-h, and cumulative O
3, as well as exceedances of O
3, compared with suburban and rural sites. The annual mean levels that are lower (
p > 0.05) at SY than at ZZL may be caused by the low nighttime O
3 at SY, possibly related to continuous emissions of NO
x from industrial activities [
53,
54,
55]. The characteristics of the O
3 relationship between urban, suburban, and rural sites can be explained by O
3 photochemical sensitivity as indicated by the O
3 diurnal pattern. The sharp decline in O
3 at the urban site in the early morning is mainly caused by the strong titration effect resulting from high levels of NO from traffic emissions when sunlight is limited for O
3 production [
53]. The slower O
3 increasing rate at the urban site than at the suburban sites suggests O
3 production might be limited by VOC availability [
21,
27].
The faster daytime O
3 increasing rates at the suburban sites suggest a more suitable photochemical environment. The highest O
3 daily peak and fastest O
3 production rate at SY were possibly due to the abundance of both precursors. We previously reported high levels of NO
2 in the surrounding area of SY, which were even higher than NO
2 levels at the urban center [
55]. Additionally, larger sources of VOCs are from vegetation emissions [
21,
23,
30]. Moreover, industrial and construction activities are also active in suburban areas; both are major sources of anthropogenic VOC emissions [
54]. According to research on O
3 at a coastal site in South China, the highest O
3 levels at the rural coastal site are probably due to weaker NO titration and a high O
3 production rate because of stronger oxidative capacity in the air [
56]. VOC data need to be measured as well to understand the O
3 dynamics.
The overall NO
x levels decreased significantly from 2011 to 2017. Winter has the highest NO
x levels and the fastest decline rate. NO
x is high in winter because of more emissions and longer NO
x lifetime during the cold season [
26]. In northern China, emissions from coal for heating are substantial; however, as a southern city, Shenzhen does not have central heating in the winter. As can be seen from the pollution rose map, the air pollutant levels in Shenzhen during the wintertime are greatly impacted by prevailing northeast winds which bring polluted air masses from northern China [
5,
28,
47]. The fastest decline in winter at both sites is possibly attributed to the effectiveness of NO
x control in northern China [
15].
Different from previous simulation results, both daily maximum 1-h and 8-h O
3 showed overall increases from 2011 to 2017 at all sites [
15,
25]. There is a clear linkage between the degree of urbanization and the direction of the O
3 trends. The increases in O
3 are profound in the winter at more urbanized sites. This rise can be attributed to increases in O
3 from the transport of continental air masses by northeasterly wind and a lowered O
3 titration effect by NO due to the NO
x reduction [
23,
57]. A small decline in O
3 occurred mainly at less urbanized sites in the autumn, the highest O
3 season [
22]. This implies that the photochemical sensitivity might shift from being VOC-sensitive to NO
x-sensitive in the autumn when photochemical reactions are the most active [
19,
57]. The increasing changes in 1-h O
3 at all sites indicate that the O
3 sensitivity in the afternoon might not be limited by NO
x availability in Shenzhen [
57].
Great challenges remain in controlling O
3 in Shenzhen. Generally speaking, the O
3 chemistry is mainly VOC-sensitive. As NO
x decreases, O
3 levels increase at all sites, with a faster increasing rate at more urbanized areas, posing greater human health risks in areas with high population density. Our results showed a slight decline of O
3 in autumn, which means that controlling NO
x emissions in this period would be the most effective way of lowering the O
3 levels in the less urbanized areas. For the more urbanized sites, emphasis should be placed on VOC reduction. The national VOC reduction strategy is specified in the 13th five-year plan (2015–2020). The goal was set to reduce major industrial VOC emissions by 10% compared with the 2015 level [
58]. In Southern China, VOC emissions are dominated by biogenic sources [
32]. More work is needed for spatial variability and the O
3 formation potential of VOCs for more efficient O
3 control management.
5. Conclusions
In this study, four stations, including one urban, one rural, and two suburban sites in Shenzhen, were investigated for characteristics and trends of O3 air pollution from 2011 to 2017. O3 matched seasonal O3 characteristics of subtropical Asia, i.e., a maximum in the autumn (October) and a minimum in the summer (June). Solar radiation, precipitation, and relative humidity are the most important meteorological factors that affect O3 daily variations.
Wind speed is the least relevant factor, but wind direction is important in impacting the presence of high O3 air concentrations. Shenzhen has at least two different sources of air pollution. One is a more regional emission source from the west and is active throughout the year. Another source is active mainly in the winter, where the high concentrations are associated with the northeasterly wind indicating the impact of air pollutant transport from northern China. The impact of the southerly wind is minimal compared with other wind directions.
O3 pollution is serious in Shenzhen, with about 28–53% of days per year exceeding the WHO ambient O3 guideline, and about 1-3 times above the limits for vegetation protection. Both the human population and vegetation are exposed to high O3, and more exceedances occur at the suburban and rural sites than at the urban site.
The O3 weekend effect is significant (p = 0.062) at the urban site with O3 levels of 1.19 ppb higher on Sunday than on weekdays. A carryover of heavy Friday evening emissions and overtime work on Saturday might be the primary reason for a diminishing weekday and Saturday O3 difference.
Both 1-h and 8-h O3 are increasing at all sites, opposite to the trend of NOx levels. The rapid decline in NOx in the winter might be related to the transport of cleaner air masses from northern China. The increase in O3 is fastest in the winter at more urbanized sites. This can be attributed to the transport of elevated O3 from northern China and a reduced O3 titration effect with NOx reduction. A slight decline in O3 occurs at less urbanized sites in autumn. This might be caused by photochemical sensitivity shifting from being VOC-sensitive to NOx-sensitive during this period.
Our results show faster O3 increases in more urbanized areas, which brings greater human health risks to areas with high population density. Great challenges remain in controlling O3 in Shenzhen with the reduction of NOx. VOC reduction should be emphasized in more urbanized areas, while NOx reduction in the autumn would be the most effective way of controlling O3 levels. The characteristics and trends of VOCs need to be studied as well to understand O3 dynamics for more efficient O3 control.