Elucidating Decade-Long Trends and Diurnal Patterns in Aerosol Acidity in Shanghai

Abstract

Aerosol acidity is a critical factor affecting atmospheric chemistry. Here, we present a study on annual, monthly, and daily variations in PM_2.5 pH in Shanghai during 2010–2020. With the effective control of SO₂ emissions, the NO₂/SO₂ ratio increased from 1.26 in 2010 to 5.07 in 2020 and the NO₃⁻/SO₄²⁻ ratio increased from 0.68 to 1.49. Aerosol pH decreased from 3.27 in 2010 to 2.93 in 2020, regardless of great achievement in reducing industrial SO₂ and NOx emissions. These findings suggest that aerosol acidity might not be significantly reduced in response to the control of SO₂ and NOx emissions. The monthly variation in pH values exhibited a V-shape trend, mainly attributable to aerosol compositions and temperature. Atmospheric NH₃ plays the decisive role in buffering particle acidity, whereas Ca²⁺ and K⁺ are important acidity buffers, and the distinct pH decline during 2010–2016 was associated with the reduction of Ca²⁺ and K⁺ while both temperature and SO₄²⁻ were important drivers in winter. Sensitivity tests show that pH increases with the increasing relative humidity in summer while it is not sensitive to relative humidity in winter due to proportional increases in ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ and aerosol liquid water content (ALWC). Our results suggest that reducing NOx emissions in Shanghai will not significantly affect PM_2.5 acidity in winter.

1. Introduction

Aerosol acidity is a critically important indicator affecting human health and atmospheric chemistry. Aerosol acidity can affect chemical reactions; for example, the main oxidation pathway of SO₂ is related to the acidity of the aerosol [1,2]. With the increase in acidity, the solubility of transition metal ions (such as Fe²⁺ and Mn²⁺) increases in atmosphere aerosol, the number of cations involved in the oxidation reaction increases, and then the ability to catalyze the oxidation of SO₂ is enhanced [3]. On the contrary, in neutral and alkaline atmospheres, the solubility of metal ions is low, the ability of metal ions to catalyze the oxidation of SO₂ decreases, and the oxidation of SO₂ by O₃ and NOx becomes the main way [4]. In addition, when the solubility of metal ions increases with the acidity of aerosols, the toxicity of aerosols is enhanced, which then affects the ecosystem [5,6,7]. Recently, significant correlations were confirmed between the pH value, water-soluble Fe, the concentration causing 20% inhibition of cell viability (IC20), and the concentration of exposure substance corresponding to a 1.5-fold increase in reactive oxygen species generation relative to control (EC1.5), indicating the strong impact of acidity on aerosol toxicity by increasing toxic equivalent concentrations of metals [8]. In addition, aerosol acidity can influence the acid-catalyzed heterogeneous reactions in the atmosphere, leading to a potential multiplication of secondary organic aerosol (SOA) mass [9,10]. Furthermore, aerosol acidity controls the phase distribution of semi-volatile compounds, such as HNO₃/NH₄NO₃ and HCOOH/HCOONH₄ systems [11,12].

Over the past decade, the Chinese government has implemented strict air pollution control strategies, such as the Air Pollution Prevention and Control Action Plan (2013–2017) and the three-year action plan for winning the blue-sky defense battle (2018–2020), leading to a significant reduction in PM_2.5 and great changes in aerosol chemical compositions. In the Yangtze River Delta (YRD), the proportion of sulfate in PM_2.5 was reduced from 18% to 14% between 2011 and 2018, while the proportion of nitrate increased from 24% to 29% [13,14]. Similarly, the concentration of PM_2.5 in Beijing significantly decreased from 88 μg m⁻³ in 2011 to 26 μg m⁻³ in 2020, and the mass ratio of NO₃⁻/SO₄²⁻ in PM_2.5 increased from 0.88 to 1.70 during the same period [15].

The acidity of atmospheric particles has been extensively studied before. Based on model predictions that the aerosols were moderately acidic, Liu et al. (2017) questioned the role of the aqueous oxidization of SO₂ by NO₂ in sulfate productions in haze events in China [16]. For the highly acidic aerosols in Canada, aerosol pH has different responses to the changes in chemical composition in different seasons [17]. An observation in a northern city in China showed that aerosol pH values ranged between 0.33 and 13.6, and were highly dependent on the source contributions of water-soluble ions such as coal combustion, mineral dust, and vehicle exhaust [18]. Similarly, Sharma et al. (2022) highlighted the important role of SO₄²⁻, NH₃, and K⁺ in determining aerosol pH [19]. Fu et al. (2022) suggested that seasonal pH changes were mainly determined by aerosol compositions in Shanghai and reducing NH₃ emissions by 20% could not effectively mitigate winter PM_2.5 pollution but significantly increased particle acidity [20].

Changes in the aerosol chemical composition may change aerosol acidity. A model simulation showed that PM_2.5 pH increased from 4.4 to 5.4 during haze episodes in Beijing when the molar ratio of NO₃⁻/SO₄²⁻ increased from 1 to 5 [21]. In contrast, another observation in Beijing showed that PM_2.5 pH in winter and autumn decreased significantly with elevated TNO₃ [22]. Model simulations showed that aerosol acidity strongly decreased over Europe and North America in recent decades while it increased over Asia [1]. This assumption was supported by observations in Guangzhou [23] and Beijing [24]. However, Zhou et al. (2022) reported that aerosol acidity increased in Shanghai in recent decades [25].

An in-depth analysis of aerosol pH changes is crucial for understanding long-term trends in aerosol pH and dominant drivers of pH variations, helping to predict pH in the future and formulate air pollution control strategies. This study investigated annual, monthly, and daily variations in PM_2.5 pH in Shanghai during 2010–2020. The main objectives of this study are (1) to elucidate the long-term trends in aerosol pH in highly urbanized areas over ten years; and (2) to explore the role of meteorological conditions and chemical compositions in driving pH changes.

2. Methods

2.1. Measurement Site and Instrumentation

The observation in this study was carried out in the downtown area of Shanghai during 2010–2021. The samples from 2010 to 2013 were collected on the roof of the No. 4 Teaching Building in the main campus of Fudan University (121.50° E, 31.30° N), and the samples from 2014 to 2021 were collected in the Pudong New Area Environmental Monitoring Station (121.54° E, 31.23° N), less than 10 km away from Fudan University (Figure S1). The two sites are surrounded by commercial and residential buildings, representing Shanghai’s densely populated urban areas. There may be a small difference in aerosol composition between the two measurement sites.

A Monitor for Aerosols and Gases in Air (MARGA, ADI 2080, Metrohm Applikon B.V., Barendrecht, The Netherlands) with a time resolution of 1 h was used to determine aerosol SO₄²⁻, NO₃⁻, Cl⁻, NH₄⁺, Na⁺, K⁺, Ca²⁺, and Mg²⁺, and their gas precursors NH₃, HCl, and HNO₃. Due to the lack of MARGA data in 2012–2013, filter-based sampling with a time resolution of 24 h was used instead. Data were not available for 2011. Detailed information on MARGA was reported previously [20]. In brief, the water-soluble gases were absorbed by a wet rotating denuder while the water-soluble ions in the aerosols were extracted using a steam jet aerosol collector. The collected gas and aerosol samples were analyzed online using a dual-channel ion chromatograph. Technical specifications for QA/QC followed local standards in the YRD region (DB31/T 310006-2021), which regulate that the correlation coefficients (r) for all targeted ions are larger than 0.995 [20]. The method detection limit of all components was 0.10 μg per cubic meter of air or better, except for K⁺ (0.16 μg m⁻³), Mg²⁺ (0.12 μg m⁻³), and Ca²⁺ (0.21 μg m⁻³) [26]. For filter-based sampling, QA/QC parameters for the determination of inorganic ions were previously reported [27]. Hourly meteorological data (temperature, relative humidity (RH), wind speed, and wind direction) were released by the China Meteorological Administration website. Hourly concentrations of PM_2.5, PM₁₀, SO₂, NO₂, CO, and O₃ were provided by Pudong Environmental Monitoring Center.

2.2. ISORROPIA II

ISORROPIA II is a computational and efficient aerosol thermodynamic equilibrium model, which has been widely used to calculate the thermodynamic equilibrium in the aerosol NH₄⁺-SO₄²⁻-NO₃⁻-Cl⁻-Na⁺-Ca²⁺-K⁺-Mg²⁺-H₂O system and the corresponding gas precursor [28]. ISORROPIA Ⅱ operates in two modes: forward mode and reverse mode. In the forward mode, the input variables are temperature, RH, and the concentrations of TNH₃ (NH₃ and NH₄⁺), TNO₃ (HNO₃ and NO₃⁻), TCl (HCl and Cl⁻), SO₄²⁻, Na⁺, Ca²⁺, K⁺, and Mg²⁺. In the reverse mode, the input variables are ambient temperature, RH, and the concentrations of SO₄²⁻, NO₃⁻, Cl⁻, NH₄⁺, Na⁺, Ca²⁺, K⁺, and Mg²⁺. The output of both modes is aerosol liquid water content (ALWC), hydrogen ion content, and the concentration of species in the gas and aerosol phases [29]. The forward model predicts aerosol composition and gas-particle partitioning much more accurately than the reverse model, because the forward model uses gas and aerosol inputs, effectively limiting the impact of measurement errors [30]. Therefore, our study chose to run the forward mode and assumed that the aerosol was metastable with no solid precipitation. Aerosol pH is defined as follows:

$$pH=-{{\mathit{log}}_{10}{\gamma }_{H^{+}}H}_{aq}^{+}=-{\mathit{log}}_{10}\left({\displaystyle \frac{{{\gamma }_{H^{+}}H}_{air}^{+}}{ALWC}}\right)$$

(1)

where ${\mathsf{\gamma }}_{{\mathrm{H}}^{+}}$ refers to the activity coefficient of hydrogen ion (assumed to be 1). $H_{aq}^{+}$(mol L⁻¹) and $H_{air}^{+}$(ng m⁻³) are the concentrations of H⁺ in aqueous particles and ambient air, respectively. ALWC (μg m⁻³) is the water uptake by inorganic species.

Overall, nitrate in the particle phase was underestimated when RH was lower than 40% since urban particles possibly dehydrate at RH below 40% [20]. When RH is above 95%, ALWC may grow exponentially with RH, resulting in large uncertainties of pH [31]. Therefore, samples collected at RH > 95% and RH < 40% were discarded in the study.

3. Results and Discussion

3.1. Long-Term Variations in the Nitrate-to-Sulfate Ratio

Figure 1 shows the decade changes in SNA (sulfate, nitrate, and ammonium) and corresponding gas precursors from 2010 to 2020. The annual variation in atmospheric SO₂ showed a distinct downward trend, from an average of 28.8 ± 10.8 μg m⁻³ in 2010 to 7.1 ± 2.8 μg m⁻³ in 2020. In contrast, the long-term trend in atmospheric NO₂ was relatively flat, with an average of 38.7 μg m⁻³ in 2010 and 36.1 μg m⁻³ in 2020. As a result, the atmospheric ratio of NO₂/SO₂ increased from 1.26 in 2010 to 5.07 in 2020. Like their gas precursors, the sum concentration of sulfate plus nitrate decreased from approximately 20 μg m⁻³ to less than 12 μg m⁻³, a decline of 40%. The lower SNA concentration in 2010 relative to the following years might be attributed to different sampling locations. The mass ratio of nitrate to sulfate (NO₃⁻/SO₄²⁻) increased from 0.68 to 1.49, with a ratio below 1.0 before 2015. Ye et al. (2021) reported that NO₃⁻/SO₄²⁻ in rainwater increased from approximately 0.3 to above 1.0 during the same period [32], consistent with the aerosol NO₃⁻/SO₄²⁻ ratio. The lower ratio of NO₃⁻/SO₄²⁻ in rainwater might be associated with the in-cloud reactions of SO₂. It is worth noting that the ratio of NO₃⁻/SO₄²⁻ was always less than that of NO₂/SO₂ due to the longer atmospheric life of SO₂. These variations can be attributable to the non-proportional reduction of SO₂ and NO_x emissions. As illustrated in Figure S2, industrial SO₂ emissions in Shanghai were reduced from 2.2 × 10⁵ tons in 2010 to 5.2 × 10³ tons in 2020, indicating that the industrial desulfurization and clean energy substitution strategies have achieved great success. Since the reduction in industrial NOx emissions was almost offset by the increasing number of on-road vehicles, NOx pollution shifted from industry-dominated to vehicle-dominated in 2016. Generally, the total emissions of acidic pollutants were significantly reduced. Meanwhile, the concentration of Ca²⁺ dropped by 76% during the observation period, indicating the significant achievement in the control of soil dust and the weakened capacity to neutralize atmospheric acidity. These findings suggest that aerosol acidity might not be significantly reduced in response to the control of SO₂ and NO_x emissions.

3.2. Variation Trend in Aerosol pH

3.2.1. Annual Variations in Aerosol pH

Figure 2 shows the long-term trend in aerosol pH in Shanghai over the decade from 2010 to 2020. The aerosol acidity showed an overall enhanced trend, with the annual average pH decreasing from 3.27 in 2010 to 2.93 in 2020. We should point out that the potential contributions of secondary organic aerosols to water uptake and organic acids to H⁺ are not considered in this study, contributing to uncertainty in aerosol pH assessment. In contrast to the atmospheric sulfate and nitrate decrease by 40% over the decade, aerosol acidity increased by 0.34 units pH, supporting the idea that aerosol acidity responded nonlinearly to SO₂ and NO_x emission control. Similar to our results, an annual pH decline rate of around 0.04 units was reported previously [25]. The small discrepancies between the two studies could be attributed to different sampling locations and statistical methods. The pH of rainwater increased by 0.8 units during the same period [32], possibly because rain droplets can scavenge alkaline coarse particles. The median pH for each year was generally higher than the mean, indicating that the distribution of hourly pH was skewed to the lower pH direction. The changing trend in aerosol pH displays a two-stage shape, with a distinct decrease during 2010–2016 and a weaker variation from 2016 to 2020. Meanwhile, the concentration of Ca²⁺ decreased from 0.85 μg m⁻³ to 0.13 μg m⁻³ from 2010 to 2016 while it remained around 0.1 μg m⁻³ from 2016 to 2020, indicating that the nonvolatile cations were important drivers affecting the decade aerosol pH variations.

Table 1. Concentrations of major inorganic ions and aerosol pH in different measurement sites worldwide (unit: μg m⁻³).

No.	Location	SO42−	NO3−	NH4+	NH3	pH	Reference
1	Shanghai, CN	8.3 ± 6.6	8.5 ± 10.8	5.6 ± 5.7	5.4 ± 3.7	3.1 ± 0.6	This study
2	Nanjing, CN	23.2 ± 0.7	16.9 ± 1.7	5.3 ± 0.6	/	3.3 ± 0.1	[35]
3	Wuhan, CN	10.1 ± 6.1	10.2 ± 9.2	8.0 ± 5.9	9.9 ± 5.4	3.0 ± 1.0	[36]
4	Guangzhou, CN	9.9 ± 5.9	4.0 ± 4.3	3.8 ± 3.0	8.7 ± 5.8	2.5 ± 0.7	[37]
5	Beijing, CN	7.7 ± 7.5	13.6 ± 16.1	7.4 ± 7.8	18.3 ± 7.8	4.2 ± 0.4	[22]
6	Tianjin, CN	8.9 ± 1.3	10.7 ± 2.8	6.9 ± 1.8	/	4.9 ± 1.4	[18]
7	Alabama, US	1.7 ± 1.2	0.1 ± 0.1	0.5 ± 0.3	0.4 ± 0.3	1.0 ± 0.5	[38]
8	Pasadena, US	1.9 ± 0.7	3.7 ± 1.5	1.8 ± 0.7	0.8 ± 0.6	2.7 ± 0.3	[11]
9	Crete, GR	2.3 ± 1.6	0.1 ± 0.1	0.8 ± 0.6	/	1.3 ± 1.1	[39]
10	Harrow, CA	7.2 ± 1.1	0.5 ± 0.4	2.0 ± 0.2	1.7 ± 0.2	2.5 ± 1.5	[40]
11	Indo-Gangetic Plain, India	18.6 ± 5.8	1.1 ± 0.8	20.3 ± 5.2	/	3.0 ± 0.2	[19]
12	Cabauw, NL	2.2 ± 0.3	4.3 ± 1.2	1.8 ± 0.3	7.4 ± 2.8	3.6 ± 0.3	[41]
13	Po Valley, Italy	1.8 ± 1.1	8.2 ± 5.9	3.3 ± 2.0	/	3.7 ± 0.6	[42]

CN: China; US: United States; CA: Canada; GR: Greece; NL: Netherlands.

summarizes aerosol pH in different measurement sites over the world. The overall aerosol pH in Shanghai (3.1 ± 0.6) was comparable to Nanjing (3.3 ± 0.1) and Wuhan (3.0 ± 1.0) in the middle and lower reaches of the Yangtze River, but significantly lower than that of Beijing (4.2 ± 0.4) and Tianjin (4.9 ± 1.4) in North China. Intriguingly, these aerosols with high loadings of sulfate and nitrate were weakly acidic while the aerosols were highly acidic (pH ≈ 1.0) in less-polluted Alabama and Crete where sulfate and nitrate loadings were much lower, indicating that aerosol acidity was not necessarily consistent with the mass loadings of sulfate and nitrate. As shown in Table 1, aerosol pH was positively correlated with the concentration of atmospheric NH₃, suggesting that ammonia was the major driving factor affecting global particle acidity distributions. Wang et al. (2016) attributed the difference between highly acidic London smog in 1952 and less acidic Beijing aerosols in 2012 to NH₃ levels [33]. The lower pH in Guangzhou than in Shanghai might be attributed to higher temperatures. These findings indicated that aerosol pH was much more sensitive to the amount of atmospheric NH₃ available for neutralizing acidic sulfate and nitrate. In addition, nonvolatile cations such as Na⁺, K⁺, Ca²⁺, and Mg²⁺ could affect aerosol acidity when their concentrations were significant relative to anions [34]. To validate the central role of NH₃, aerosol pH in Pasadena and Beijing was re-calculated by inputting the average NH₃ concentration in Shanghai (5.4 μg m⁻³). The re-calculated pH increased from 2.7 to 4.0 in Pasadena as the NH₃ concentration increased from 0.8 to 5.4 μg m⁻³ and decreased from 4.2 to 3.5 in Beijing as the NH₃ concentration decreased from 18.3 to 5.4 μg m⁻³, confirming that both ammonia and alkaline nonvolatile cations were important buffers of aerosol acidity.

3.2.2. Seasonal Variations in Aerosol pH

Figure 3 illustrates the long-trend in seasonal pH values from 2010 to 2020. The monthly variation in pH values exhibited a V-shape trend, decreasing from February and increasing from September. Aerosol pH in August was the lowest (with an average of less than 2.5) and the fluctuation range was the largest. From a seasonal perspective, aerosol pH followed the order of winter > spring > autumn > summer, indicating that ambient temperature was an important driver for pH variation. In this study, the seasons from spring to winter are defined as the period from March to May, June to August, September to November, and December to February, respectively. Similar seasonal pH trends were reported previously [17,22,36,38,43]. As indicated by Equation (1), aerosol pH depends on $H_{air}^{+}$and ALWC, which are functions of pollutant concentrations, temperature, and RH. Elevated ALWC by increasing RH can dilute $H_{aq}^{+}$ and increase pH as well as promoting nitrate formation via nighttime N₂O₅ hydrolysis. The approximately one-unit difference in pH between winter and summer was mainly attributed to aerosol compositions because a doubling ALWC in winter increases aerosol pH by only 0.3 units [20]. Although temperature is not a parameter in Equation (1), it greatly affects the photooxidation rate of SO₂ and NOx and the gas-particle partitioning of semi-volatile NH₄NO₃ and NH₄Cl, with an indirect effect on aerosol acidity. During 2010–2015, the seasonal average pH showed a steady downward trend in winter while remaining stable in summer and autumn, resulting in a decline in the seasonal pH difference from 1.14 to 0.72. Interestingly, the seasonal difference in pH returned to approximately 1.0 in 2016–2020. In addition, the seasonal average pH was unexpectedly high in the summer and autumn of 2015. The mechanisms driving seasonal pH variations will be discussed in the next sections.

3.2.3. Diurnal Variation in Aerosol pH

Figure 4 illustrates the diurnal variation in aerosol pH along with ALWC and the mass ratio of NO₃⁻/SO₄²⁻. Overall, aerosol pH began to decrease in the early morning, reaching the minimum values at noon, followed by a continuous increase in the afternoon and high values at night, further supporting the temperature dependence on aerosol acidity. In summer, the average pH values at night were approximately 0.4–0.7 units larger than those at noon. With the strongest acidity, the aerosol noon minimum dropped to nearly 2.1 in 2016. Similar to summer, aerosol pH in winter showed a V-shaped diurnal trend but the diurnal range was much narrower. In 2020, the daily variation in aerosol pH narrowed to below 0.3 units. Contrary to our results, the diurnal variation in aerosol pH over the southeastern United States was approximately one unit larger in winter than in summer [38], indicating that the diurnal pH trends are highly related to the regional pollution background.

Both ALWC and NO₃⁻/SO₄²⁻ followed a similar diurnal pattern as pH, indicating that they provided an additive effect on pH variation. As ALWC is a function of RH and the total amount of hygroscopic aerosols, ALWC generally peaked in the early morning and rapidly decreased after sunrise. ALWC at noon was reduced by 50–70% relative to the early morning, corresponding to a pH increase of 0.3–0.5 units, indicating that ALWC was the main driver for the strong diurnal pH variation. The flat diurnal pattern of the NO₃⁻/SO₄²⁻ ratio in 2010 indicates an insignificant effect of chemical composition on the diurnal trend in pH, providing a reasonable explanation for the narrowest diurnal variation in pH.

From the perspective of aerosol composition, sulfate and nitrate are regarded as the major driving factors of diurnal patterns in pH [38]. When sulfate and nitrate are neutralized primarily by ammonium, the concentration of H⁺ is controlled primarily by the thermodynamic equilibrium between particle-phase NH₄⁺ and gaseous NH₃. As (NH₄)₂SO₄ is less volatile than NH₄NO₃, the elevated sulfate leads to a much greater increase in ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$than that of TNO₃. NH₄NO₃ tended to evaporate with increasing temperature and decreasing ALWC during the daytime. In contrast, uptake on humid particles was favored at night due to its dissociation being highly sensitive to temperature and RH changes, resulting in a V-shaped diurnal pattern. The NO₃⁻/SO₄²⁻ ratio in winter was significantly larger than in summer, which can be partly responsible for the weaker acidity of winter aerosols.

3.3. Quantitative Analysis of pH Drivers

3.3.1. Effect of Alkaline Buffers on Interannual pH Variations

As aforementioned, ammonia is the major alkaline buffer for aerosol pH globally while nonvolatile cations exhibit a non-negligible impact on aerosol pH when their concentrations become significant. In this study, temperature, relative humidity, and concentrations of chemical composition were averaged for each summer and winter (Table S1). Figure 5 illustrates the aerosol pH predicted by the average concentrations and the buffer effect of NH₃ and nonvolatile cations on summer and winter aerosols. The aerosol pH values predicted by the averaged concentrations of chemical composition and meteorological parameters are close to those averaged by hourly pH values, indicating that they can represent the typical pH trends. The inserted images represent the predicted aerosol pH values by setting the concentration of NH₃ to zero while the other variables are fixed at their average values. It is worth noting that aerosol pH would be below zero in the absence of NH₃, emphasizing the decisive role of NH₃ in buffering particle acidity. Similar to NH₃, aerosol pH is predicted by inputting the concentrations of Ca²⁺ and K⁺ as 0 while other variables are fixed (Table S2). The buffering effects of Ca²⁺ and K⁺, ΔpH, are obtained by comparing the pH decline when the concentrations of Ca²⁺ and K⁺ are set to 0. Ca²⁺ and K⁺ are tracers of crustal dust and biomass burning, respectively. Coal combustion is another important source of K⁺ in many Chinese cities [20]. The buffing effect of Ca²⁺ was more significant than K⁺ before 2013, but K⁺ remained higher in 2014–2015. The combined buffing effect of Ca²⁺ and K⁺ decreased by 0.4 units of pH during the winter of 2010–2016, corresponding to the aerosol pH decline during this period, indicating that the reduction in acidic pollutant emissions was largely offset by the reduction in alkaline buffers. During 2016–2020, the buffing effect of nonvolatile cations could be ignored in summer while the buffing effect of K⁺ remained at approximately 0.4 units pH in winter. Long-range pollutant transport from north China contributes greatly to the increase in K⁺ concentration and PM_2.5 pollution in Shanghai because the prevailing wind direction in Shanghai varies from northwest to northeast in the winter [26,44]. The increase in K⁺ in winter was possibly related to the transportation of biomass burning and coal combustion pollutants under the regime of prevailing winds from the north. In contrast to aerosol pH below zero in the absence of NH₃, the largest buffering effect of nonvolatile cations was less than 0.6 units pH, emphasizing the decisive role of NH₃ in buffering particle acidity.

3.3.2. Effects of Meteorological Parameters and Chemical Composition on Diurnal pH Variations

The contributions of individual driving factors to diurnal pH variations are estimated in Figure 6 and Tables S4–S7. Similar to the above section, the effect of a driving factor on aerosol pH is evaluated by replacing this factor at 10:00, 14:00, 20:00, and 24:00 with the value at 5:00 in the morning. The reference point is set at 5:00 when aerosol pH was generally the highest. In both summer and winter, the temperature during the day was higher than that at night, while the humidity was just the opposite. The increase in temperature can reduce aerosol pH by partitioning aerosol NH₄NO₃ and NH₄Cl into the gas phase, leading to decreases in NO₃⁻/SO₄²⁻ and pH [41]. The effect of RH on aerosol pH is more complicated. On the one hand, elevated RH can enhance ALWC which dilutes the ionic concentration. On the other hand, the increase in ALWC favors more gaseous NH₃ and HNO₃ partitioning into the particle phase. The diurnal pH variation in summer was mainly driven by temperature and RH. Only a 7 °C increase in temperature is required for a 0.5 unit drop in pH, which is lower than that in Canada [17], possibly due to the higher mass loading of secondary inorganic aerosols in our study. As shown in Tables S4 and S6, the narrow fluctuations of dew point temperature (T_d) indicate that the diurnal RH variations were mainly driven by temperature, further highlighting the important role of temperature in the diurnal variations in aerosol pH. The effects of SO₄²⁻, TNO_3, and TNH₃ were much weaker than those of temperature and RH in summer. It is worth noting that both elevated SO₄²⁻ and decreased NO₃⁻ in the daytime contributed to the decreasing aerosol pH. In contrast, the main drivers of diurnal pH variation in winter were temperature and SO₄²⁻. Another significant difference from summer was that the decreasing RH could enhance aerosol pH in winter. It is worth noting that RH played a minor role in diurnal pH patterns in winter. In summary, the diurnal variation in aerosol pH in summer is mainly affected by temperature and RH, while the diurnal variation in aerosol pH in winter is sensitive to both meteorological parameters and aerosol chemical composition.

For an in-depth understanding of the effect of meteorological parameters and chemical composition on aerosol pH, sensitivity tests were performed based on two cases of summer and winter aerosols (Figure 7). The concentration of NH₄⁺ decreased nonlinearly with the increase in temperature because the concentration product of NH₃ and HNO₃ is an exponential function of temperature [4]. Aerosol pH almost linearly decreased with the increase in temperature since the concentration of H⁺ in ambient air $({\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$) almost exponentially increases with the increasing temperature. Although ALWC always increased with the increase in RH, ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ displayed different trends between summer and winter. In summer, ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ first decreased with the increase in RH which favors more gaseous NH₃ partitioning into the aerosol water. In contrast, ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ in winter increased with the increase in RH, leading to a slight pH decrease at RH < 80%. As shown in Figure 7e,f, the concentration of NH₄⁺ increased linearly with the increase of SO₄²⁻ due to the higher affinity of H₂SO₄ to NH₃. However, the hygroscopic growth of (NH₄)₂SO₄ was lower than that of NH₄NO₃, indicating that aerosol pH decreased with the decrease of NO₃⁻/SO₄²⁻. In contrast, the concentration of aerosol NH₄⁺ only slightly increased with the increasing TNO₃ in summer due to most of the TNO₃ partition to the gas phase. Aerosol pH increased with the increase in TNO₃ due to the increase in NO₃⁻/SO₄²⁻. In winter, aerosol pH increased slightly with the increase in TNO₃ for NO₃⁻/SO₄²⁻ < 1.5. However, aerosol pH slightly decreased with the increase in TNO₃ for higher NO₃⁻/SO₄²⁻. An observation in Beijing showed that PM_2.5 pH increased with increasing NO₃⁻/SO₄²⁻ [21], while another observation found that PM_2.5 pH decreased with increasing TNO₃ [22]. This finding suggests that the impact of NO₃⁻/SO₄²⁻ on pH depends on the pollution background, providing a reasonable explanation for the different trends observed in Beijing. Our results suggest that reducing NOx emissions in Shanghai will not significantly affect PM_2.5 acidity in winter.

4. Conclusions

This study investigated the long-term trends in PM_2.5 pH in response to emission control in Shanghai. The annual average ratio of NO₃⁻/SO₄²⁻ increased from 0.68 in 2010 to 1.49 in 2020, attributable to the significant reduction in SO₂ emissions and the less effective control of NOx emissions. PM_2.5 acidity showed a slightly increasing trend since the reduction in acidic emissions was partly offset by the decrease in alkaline nonvolatile cations. The monthly variation in pH values exhibited a V-shape trend, decreasing from February and increasing from September, mainly attributed to aerosol compositions and temperature which controls the partitioning of HNO₃/NH₄NO₃. The diurnal pH pattern showed a V-shaped trend with stronger fluctuation in summer than in winter due to diurnal variations in ALWC and NO₃⁻/SO₄²⁻. Atmospheric NH₃ plays a decisive role in buffering particle acidity, providing a plausible explanation on moderately acidic aerosols in the Yangtze River Delta, highly polluted with NOx emissions. Ca²⁺ and K⁺ were important buffers of particle acidity and the reduction in Ca²⁺ and K⁺ was responsible for the pH decline during 2010–2016.

The diurnal pH variations in summer were mainly affected by temperature and RH. The diurnal RH variations were mainly driven by temperature, underlying the decisive role of temperature in the diurnal variations in aerosol pH. In contrast, the dominant drivers of diurnal pH variation in winter were attributed to temperature and SO₄²⁻. Sensitivity tests showed that aerosol pH almost linearly decreased with the increase in temperature since the concentration of ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ almost exponentially increased with the increase in temperature. pH increased with the increasing RH in summer while it was not sensitive to RH in winter due to proportional increases in ${\mathrm{H}}_{\mathrm{a}\mathrm{i}\mathrm{r}}^{+}$ and ALWC. pH was not very sensitive to TNO₃ in winter, indicating that reducing NOx emissions in Shanghai will not significantly affect PM_2.5 acidity in winter.