Different physicochemical behaviors of nitrate and ammonium during transport: a case study on Mt. Hua, China

To understand the chemical evolution of aerosols in the transport process, the chemistry of PM_2.5 and nitrogen isotope compositions on the mountainside of Mt. Hua (∼1120 m above sea level, a.s.l.) in inland China during the 2016 summertime were investigated and compared with parallel observations collected at surface sampling site (∼400 m a.s.l.). The PM_2.5 exhibited a high level at the mountain foot site (MF; average 76.0±44.1 µg m⁻³) and could be transported aloft by anabatic valley winds, leading to the gradual accumulation of daytime PM_2.5 with a noon peak at the mountainside sampling site (MS). As the predominant ion species, sulfate exhibited nearly identical mass concentrations at both sites, but its PM_2.5 mass fraction was moderately enhanced by ∼4 % at the MS site. The ammonium variations were similar to the sulfate variations, the chemical forms of both of which mainly existed as ammonium bisulfate (NH₄HSO₄) and ammonium sulfate ((NH₄)₂SO₄) at the MF and MS sites, respectively. Unlike sulfate and ammonium, nitrate mainly existed as ammonium nitrate (NH₄NO₃) in fine particles and exhibited decreasing mass concentration and proportion trends with increasing elevation. This finding was ascribed to NH₄NO₃ volatilization, in which gaseous HNO₃ from semi-volatile NH₄NO₃ subsequently reacted with dust particles to form nonvolatile salts, resulting in significant nitrate shifts from fine particles into coarse particles. Such scavenging of fine-particle nitrate led to an enrichment in the daytime ¹⁵N of nitrate at the MS site compared with to the MF site. In contrast to nitrate, at the MS site, the ¹⁵N in ammonium depleted during the daytime. Considering the lack of any significant change in ammonia (NH₃) sources during the vertical transport process, this ¹⁵N depletion in ammonium was mainly the result of unidirectional reactions, indicating that additional NH₃ would partition into particulate phases and further neutralize ${\mathrm{HSO}}_{\mathrm{4}}^{-}$ to form ${\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}$. This process would reduce the aerosol acidity, with a higher pH (3.4±2.2) at the MS site and lower ones (2.9±2.0) at the MF site. Our work provides more insight into physicochemical behaviors of semi-volatile nitrate and ammonium, which will facilitate the improvement in the model for a better simulation of aerosol composition and properties.

1 Introduction

Atmospheric particulate matter measuring equal to or less than 2.5 µm in aerodynamic diameter (PM_2.5) is a worldwide air pollution burden that can deteriorate the urban air quality and induce adverse human health effects that contribute to lowered life expectancies (Shiraiwa et al., 2017; Lelieveld et al., 2015; Fuzzi et al., 2015; Wang et al., 2016). Recent studies have disclosed that the mechanisms underlying these effects are profoundly dependent on particle properties, e.g., the size, concentration, mixing state, and chemical compositions of particles (Li et al., 2016; Liu et al., 2021; Guo et al., 2014). Thus, since 2013, China has issued strict emission directives to mitigate haze pollution. Consequently, the annual PM_2.5 concentration in China fell by approximately one-third from 2013–2017 (Zheng et al., 2018). These controls notwithstanding, the PM_2.5 levels in most cities in China still exceed the least-stringent target of the World Health Organization (2021; 35 µg m⁻³), especially in rural areas and small cities (Lv et al., 2022; Li et al., 2023).

Near-surface particulate matter (PM) can also be transported to the upper air, and this process critically impacts radiative forcing, cloud precipitation, and the regional climate by scattering/absorbing solar radiation and by influencing aerosol–cloud interactions (Van Donkelaar et al., 2016; Andreae and Ramanathan, 2013; Fan et al., 2018). Past assessments of these effects have been characterized by large uncertainties (Carslaw et al., 2013); for example, Bond et al. (2013) found that black carbon climate forcing varied from +0.17 to +2.1 W m⁻² with a 90 % uncertainty. Such massive uncertainties are mainly due to our limited knowledge regarding the spatiotemporal distribution, abundance, and compositions of airborne PM (Seinfeld, 2016; Raes et al., 2000). In addition, aerosols may undergo aging during the vertical transport process, causing increasingly complex compositions and changes in aerosol properties. Despite these factors, to date, vertical observations remain comparatively scarce compared to surface measurements. Therefore, to obtain an improved understanding of the fundamental chemical and dynamical processes governing haze development, more field observations of upper-layer aerosols are necessary, as these measurements could provide updated kinetic and mechanistic parameters that could serve to improve model simulations.

Currently, various monitoring approaches have been developed and applied to measure vertical aerosols, e.g., satellite remote sensing and in situ lidar methods; these approaches can be used to obtain the pollution concentration profiles (Van Donkelaar et al., 2016; Reid et al., 2017). To accurately measure chemical compositions, aircraft and unmanned aerial vehicles (UAVs) equipped with a variety of instruments can be utilized in short-term sampling campaigns (Lambey and Prasad, 2021; Zhang et al., 2017), but these tools are unsuitable for long-term continued observations due to their high operational costs. In cases of near-surface vertical urban atmosphere observations, techniques involving tethered balloons, meteorological towers, and skyscrapers are usually adopted (Zhou et al., 2020; Xu et al., 2018; Fan et al., 2021). However, the vertical application range of these methods is limited to only ∼500 m, thus hardly meeting the requirements of research conducted above the boundary layer. Therefore, high-elevation mountain sites have long been regarded as suitable places for long-term research on the upper-layer aerosol (including its composition, chemical–physical properties, formation processes, etc.; Dzepina et al., 2015; Zhou et al., 2021; Wang et al., 2013), which are conducive to better understanding of the haze episodes in the lower troposphere. Although the fixed observation position is the key drawback of this monitoring approach, it has still been widely selected for use in various vertical observation campaigns, e.g., in past studies conducted in Salt Lake Valley (Baasandorj et al., 2017), in Terni Valley (Ferrero et al., 2012), and on Mt. Tai (Meng et al., 2018; Wang et al., 2011).

Mt. Hua adjoins the Guanzhong basin of inland China, where haze pollution has been a persistent environmental problem (Wu et al., 2020a, 2021; Wang et al., 2016). In our previous studies conducted at the mountaintop of Mt. Hua, we found that air quality was significantly affected by surface pollution, and distinctive differences were found in the aerosol compositions and size distributions at the mountaintop compared to those measured at lower elevations near ground level (Wang et al., 2013; Li et al., 2013). With the implementation of strict emission controls, the atmospheric environment in this region has changed dramatically from the SO₂/sulfate-dominated previous environment to the current NO_x/nitrate-dominated environment (Baasandorj et al., 2017; Wu et al., 2020c). However, the fundamental chemical and dynamical processes driving this PM_2.5-loading explosion are unclear under the current atmospheric state with increasing O₃ and NH₃ levels. To better rationalize these processes, in this work, 4 h integrated aerosol samples were synchronously collected on the mountainside and at the lower-elevation land surface, and the chemical components and stable nitrogen isotope compositions of nitrate and ammonium were analyzed in the collected PM_2.5 samples. We compared the chemical compositions and diurnal cycles between the two sampling sites and then discussed the changes in the chemical forms of secondary inorganic ions during their vertical transport from lower to higher elevations. Our study revealed that nitrate and ammonium exhibited distinct physicochemical behaviors during the aerosol-aging process.

2 Experiment

2.1 Sample collection

In this campaign, the PM_2.5 samples were synchronously collected at two locations in the Mt. Hua area during the period from 27 August to 17 September 2016. One sampling site was located on a building belonging to the Huashan Meteorological Bureau (34^∘32^′ N, 110^∘5^′ E; 400 m above sea level, a.s.l.) at the foot of Mt. Hua. Surrounded by several traffic arteries and dense residential and commercial buildings, as shown in Fig. 1b, this site is an ideal urban station for studying the impacts of anthropogenic activities on local air quality and is referred to hereafter as the MF site. The mountainous sampling site (34^∘29^′ N, 110^∘3^′ E; 1120 m a.s.l.) was located approximately 8 km from the city site horizontally (Fig. 1c) at an elevation of 720 m above the average Huashan town level of ∼400 m a.s.l. This site was situated on a mountainside that experiences little anthropogenic activity due to its steep terrain and is abbreviated hereafter as the MS site. Furthermore, this location adjoins one of the larger valleys of Mt. Hua; therefore, the measurements taken at this location were strongly affected by the lower-elevation air pollutants transported upwards by the valley winds. At both measurement sites, the PM_2.5 aerosol samples with a 4 h interval in were collected onto prebaked (at 450 ^∘C for 6 h) quartz filters using high-volume (1.13 m³ min⁻¹) air samplers (Tisch Environmental, Inc., USA). All air samplers were installed on the roofs of buildings, approximately 15 m above the local ground surface. Furthermore, size-resolved aerosol sampling was synchronously conducted at two sites during summertime (10–22 August, 2020). These samples, with nine size bins (cutoff points were 0.43, 0.65, 1.1, 2.1, 3.3, 4.7, 5.8, and 9.0 µm, respectively), were collected using an Anderson sampler at an airflow rate of 28.3 L min⁻¹ for ∼72 h. After sampling, the filter samples were stored in a freezer (at −18 ^∘C) prior to analysis.

Figure 1 (a) Location of the study sites in China, (b) topographic view of Mt. Hua, with the sampling sites marked, and (c) vertical views of the two sampling sites and the horizontal distance between them. The maps are the reproductions from © Mapbox (https://account.mapbox.com/, last access: 31 December 2021), with added marks for our study locations.

The hourly PM_2.5, NO_x, and O₃ mass concentrations were detected at the mountainside sampling site using an E-BAM, a chemiluminescence analyzer (Thermo Fisher Scientific, model 42i, USA) and a UV photometric analyzer (Thermo Fisher Scientific, model 49i, USA), respectively. At the MF site, only PM_2.5 was monitored, using another E-BAM, while the data of the other species were downloaded from the Weinan Ecological Environment Bureau (http://sthjj.weinan.gov.cn/, last access: 8 July 2021). Meteorological data characterizing both sampling sites throughout the whole campaign were obtained from the Shaanxi Provincial Meteorological Bureau (http://sn.cma.gov.cn/, last access: 8 July 2021).

2.2 Chemical analysis

Four punches (1.5 cm diameter) of each aerosol sample were extracted into 10 mL Milli-Q pure water (18.2 MΩ) under sonication for 30 min. Subsequently, the extracts were filtered with 0.45 µm syringe filters and water-soluble ions were detected (Na⁺, ${\mathrm{NH}}_{\mathrm{4}}^{+}$, K⁺, Mg²⁺, Ca²⁺, ${\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}$, ${\mathrm{NO}}_{\mathrm{2}}^{-}$, ${\mathrm{NO}}_{\mathrm{3}}^{-}$, and Cl⁻) by using ion chromatography. The detection limit for these nine ions was <0.01 µg mL⁻¹. A DRI model 2001 thermal–optical carbon analyzer was used, following the IMPROVE_A temperature protocol to analyze the organic carbon (OC) and elemental carbon (EC) in each PM_2.5 filter sample (in 0.526 cm² punches). For more details regarding the utilized methods, readers can refer to our previous studies (Wu et al., 2020a).

To quantify the stable nitrogen isotope compositions of nitrate (δ¹⁵N-${\mathrm{NO}}_{\mathrm{3}}^{-}$) and ammonium (δ¹⁵N-${\mathrm{NH}}_{\mathrm{4}}^{+}$) in PM_2.5 samples, the filter samples were pretreated, as described for the water-soluble ion analysis. The ammonium in the extracts (approximately half of the resulting solution) was oxidized by hypobromite (BrO⁻) to nitrite (${\mathrm{NO}}_{\mathrm{2}}^{-}$), which was subsequently reduced by hydroxylamine (NH₂OH) in a strongly acidic environment. The above product (N₂O) was then analyzed by a commercially available purge-and-cryogenic-trap system coupled to an isotope ratio mass spectrometer (PT-IRMS). A bacterial method (Pseudomonas aureofaciens, a denitrifying bacterium without N₂O reductase activity) was used to convert the sample ${\mathrm{NO}}_{\mathrm{3}}^{-}$ into N₂O, which was ultimately quantified through PT-IRMS. As revealed in previous studies (Fang et al., 2011), the presence of ${\mathrm{NO}}_{\mathrm{2}}^{-}$ in aerosols may interfere with the denitrifier method when measuring δ¹⁵N. Nonetheless, ${\mathrm{NO}}_{\mathrm{2}}^{-}$ generally composed tiny portions in most of our samples and, on average, contributed <1.0 % to ${\mathrm{NO}}_{\mathrm{3}}^{-}+{\mathrm{NO}}_{\mathrm{2}}^{-}$. Thus, we believe that the proportion of ${\mathrm{NO}}_{\mathrm{2}}^{-}$ in the sample was too small to affect the resulting δ¹⁵N measurements, based on the discussions reported by Wankel et al. (2010). More details regarding the analytical artifact and quality control protocols can be found elsewhere (Wu et al., 2021; Liu et al., 2014).

2.3 Concentration-weighted trajectory (CWT) analysis

CWT is a powerful tool used here to reveal the potential spatial sources responsible for the high PM_2.5 loadings measured on Mt. Hua; this method has been used previously in similar studies (Wu et al., 2020c, b). In this study, the CWT analysis was conducted using the Igor-Pro-based tool coupled with hourly PM_2.5 concentrations and 12 h air mass backward trajectories that were simulated by using the Hybrid Single-Particle Lagrangian Integrated Trajectory (HYSPLIT) model (Petit et al., 2017).

2.4 Theoretical calculations of the partial pressures of NH₃ and HNO₃ and the dissociation constant of NH₄NO₃

To obtain the product of the partial pressures of NH₃ and HNO₃, the NH₄NO₃ deliquescence relative humidity (DRH) was first calculated using Eq. (1). The average DRH of NH₄NO₃ between the two sites was 65.0±2.9 %, which is slightly lower than the atmospheric RH (66.0±19.3 %). As the works by Wexler and Seinfeld (1991) and Tang and Munkelwitz (1993) revealed, aerosols are multicomponent mixtures, meaning the aerosol DRH is always lower than the DRH of the individual salts in the particles. Thus, the actual DRH of the aerosols observed in this study would be lower than the calculated DRH of NH₄NO₃. Based on these analyses, the particles would be deliquescent most of the time, but for simplification, we always assumed that NH₄NO₃ was in an aqueous state, corresponding to the following dissociation Reaction (R1):

$$\begin{array}{}\text{(1)}& \ln \left(\text{DRH}\right)={\displaystyle \frac{\mathrm{723.7}}{T}}+\mathrm{1.6954}\end{array}$$

$$\begin{array}{}\text{(R1)}& {\mathrm{NH}}_{\mathrm{3}}\left(\mathrm{g}\right)+{\mathrm{HNO}}_{\mathrm{3}}\left(\mathrm{g}\right)\leftrightarrows {\mathrm{NH}}_{\mathrm{4}}^{+}+{\mathrm{NO}}_{\mathrm{3}}^{-}.\end{array}$$

According to the approach illustrated in the referenced work (Seinfeld, 2016), the equilibrium constant of the dissociation reaction can be described as Eq. (2).

$$\begin{array}{}\text{(2)}& {\displaystyle }{K}_{\mathrm{AN}}={\displaystyle \frac{{\mathit{\gamma }}_{{\mathrm{NH}}_{\mathrm{4}}{\mathrm{NO}}_{\mathrm{3}}}^{\mathrm{2}}{m}_{{\mathrm{NH}}_{\mathrm{4}}^{+}}{m}_{{\mathrm{NO}}_{\mathrm{3}}^{-}}}{P_{{\mathrm{HNO}}_{\mathrm{3}}}{P}_{{\mathrm{NH}}_{\mathrm{3}}}}}\text{(3)}& {\displaystyle }\begin{array}{rl}{K}_{\mathrm{AN}}=\mathrm{4}\times {\mathrm{10}}^{\mathrm{17}}\exp \mathit{\{}& \mathrm{64.7}\left({\displaystyle \frac{\mathrm{298}}{T}}-\mathrm{1}\right)+\mathrm{11.51}\\ & \left[\mathrm{1}+\ln \left({\displaystyle \frac{\mathrm{298}}{T}}\right)-{\displaystyle \frac{\mathrm{298}}{T}}\right]\mathit{\}}\end{array}\text{(4)}& {\displaystyle }\ln \left(K_{\mathrm{p}}\right)=\mathrm{118.7}-{\displaystyle \frac{\mathrm{24}\mathrm{084}}{T}}-\mathrm{6.025}\ln \left(T\right),\end{array}$$

where K_AN (mol² (kg² atm²)⁻¹) is the equilibrium constant of Reaction (R1) (this value is temperature-dependent and can be calculated by Eq. 3), ${\mathit{\gamma }}_{{\mathrm{NH}}_{\mathrm{4}}{\mathrm{NO}}_{\mathrm{3}}}$ is the binary activity coefficient for NH₄NO₃ (${\mathit{\gamma }}_{{\mathrm{NH}}_{\mathrm{4}}{\mathrm{NO}}_{\mathrm{3}}}={\mathit{\gamma }}_{{\mathrm{NH}}_{\mathrm{4}}}{\mathit{\gamma }}_{{\mathrm{NO}}_{\mathrm{3}}}$), and $m_{{\mathrm{NH}}_{\mathrm{4}}^{+}}$ and $m_{{\mathrm{NO}}_{\mathrm{3}}^{-}}$ are the molalities of ${\mathrm{NH}}_{\mathrm{4}}^{+}$ and ${\mathrm{NO}}_{\mathrm{3}}^{-}$, respectively. To calculate ${\mathit{\gamma }}_{{\mathrm{NH}}_{\mathrm{4}}{\mathrm{NO}}_{\mathrm{3}}}$ and $m_{{\mathrm{NH}}_{\mathrm{4}}^{+}}{m}_{{\mathrm{NO}}_{\mathrm{3}}^{-}}$, the activity coefficients of the corresponding ions and the aerosol water content were assessed using the E-AIM (IV) model (http://www.aim.env.uea.ac.uk/aim/model4/model4a.php, last access: 2 November 2021). Combining Eqs. (2) and (3), we obtained the product of the partial pressures of NH₃ and HNO₃ ($P_{{\mathrm{HNO}}_{\mathrm{3}}}{P}_{{\mathrm{NH}}_{\mathrm{3}}}$), obtaining an average of $\sim \mathrm{15.2}\pm \mathrm{26.0}$ ppb² (where ppb is parts per billion) at the MF site. This value was within the range of values (1.0 to 37.7 ppb²) measured by the in situ Gas and Aerosol Composition monitor (IGAC) in the summer of 2017 in Xi'an, a metropolitan city located in the Guanzhong basin of inland China that has suffered from serious haze pollution (Wu et al., 2020b). Thus, we believe that $P_{{\mathrm{HNO}}_{\mathrm{3}}}{P}_{{\mathrm{NH}}_{\mathrm{3}}}$ variations can be assessed using the above method (to a certain extent). For simplification, the dissociation constant of dry NH₄NO₃ particles (K_p; ppb²) was applied in this study, which can be calculated as a function of temperature, using Eq. (4), as was revealed by Mozurkewich (1993). Despite not considering the aerosol properties (e.g.,  acidity and mixing state) that may induce the shift of the NH₄NO₃ equilibrium states, this assessment method was also applied in a similar work conducted by Lindaas et al. (2021).

3 Results and discussion

3.1 Overview of PM_2.5 at both sites

3.1.1 Meteorological conditions and temporal variations in PM_2.5 concentrations

The temporal variations in the 4 h PM_2.5 mass concentrations, water-soluble ions, and meteorological factors measured at the two sampling sites are illustrated in Fig. 2, and the comparisons of the above variables are summarized in Table 1. The average temperature (T) and relative humidity (RH) at the MF site were 23.2±4.2 ^∘C and 68.9±18.2 % (Table 1), respectively, and these values were characterized by marked diurnal variations, as shown in Fig. 2a. However, relatively cold and moist weather frequently occurred at the MS site, which exhibited less pronounced diurnal T and RH variations, with variations approximately 8 ^∘C and 6 % lower than the mean values derived at the MF site, respectively. Windy weather (wind speed of 3.2±2.0 m s⁻¹) also prevailed at this sampling site, with gusts above 10.0 m s⁻¹; this condition is conducive to the dissipation of pollutants.

Figure 2 Time series of the temperature (T), relative humidity (RH), boundary layer height (BLH), and mass concentrations of PM_2.5 and the water-soluble ions in PM_2.5 during the observation period at the two sampling sites.

Table 1 Mass concentrations of species in the PM_2.5 samples, pH, and the meteorological conditions at the two sampling sites.

Aerosol liquid water content (ALWC) and pH are predicted by the thermodynamic model (E-AIM (IV)).

Overall, the PM_2.5 concentrations measured at the MF site varied from 22.8 to 245.6 µg m⁻³, with a mean value of 76.0±44.1 µg m⁻³, approximately corresponding to the Grade II level (75 µg m⁻³) of the National Ambient Air Quality Standard in China. Even so, the PM_2.5 levels at Huashan town (i.e., at the MF site) were still higher than those measured in many typical megacities in the summertime, e.g., Xi'an (37 µg m⁻³ in 2017; Wu et al., 2020a) and Beijing (46.3 µg m⁻³ in 2016; Lv et al., 2019). Noticeably, stagnant meteorological conditions with increasing RH (>77 %) and relatively low wind speeds (<2.0 m s⁻¹) occurred during the relatively late stage of observation, leading to a buildup of high PM_2.5 loadings (78.7 to 245.6 µg m⁻³). Such typical haze events last approximately 4 d (12 to 16 September, 2016), indicating that aerosol pollution is still severe in rural towns, despite the notable air quality improvements recorded in most Chinese urban areas. A similar temporal PM_2.5 pattern was seen at the MS site, where the average PM_2.5 concentration (47.0±38.0 µg m⁻³) was only 0.62-fold of that at the MF site and was within the range of that measured at the summit of Mt. Tai (37.9 µg m⁻³ in 2016; Yi et al., 2021) and on Mt. Lushan (55.9 µg m⁻³ in 2011; Li et al., 2015) in the summertime. As shown in Fig. 2d, a multiday episode (mean PM_2.5 of 106.3 µg m⁻³) also appeared at the MS site during the period from 12 to 15 September, corresponding to the days on which high surface pollution was recorded. This was indicative of the potential impacts of surface pollution on air quality in mountainous areas.

3.1.2 Diurnal variation in PM_2.5

As shown in Fig. 2c and d, regular diurnal PM_2.5 variations were seen throughout the whole campaign, especially at the MS site. To reveal the differences in the daily changes in PM_2.5 between the two sampling sites, the mean diurnal cycles of hourly PM_2.5 and the boundary layer height (BLH) are depicted in Fig. 3. At the MF site, the PM_2.5 concentration was moderately enhanced during the nighttime, with a daily maximum (88.2±53.0 µg m⁻³) observed at 06:00 local standard time (LST). After sunrise, PM_2.5 exhibited a decreasing trend until ∼ 15:00 LST, corresponding to thermally driven boundary layer growth. Conversely, the aerosol concentrations at the MS site immediately increased as the boundary layer uplifted in the early morning and peaked at 14:00 LST, when the MS site was located completely within the interior of the boundary layer. Proverbially, anabatic valley winds prevail in mountainous regions during the daytime. Thus, the aerosol-rich air at the MF site may be transported aloft by the prevailing valley breeze, leading to significantly enhanced PM_2.5 levels at the MS site in short time periods. This finding was further verified by the similar diurnal NO₂ pattern identified at the MS site, as illustrated in Fig. S1 in the Supplement. In the forenoon period, continuous enhancement in the NO₂ level was observed at the MS site, with a daily maximum of 14.4±53.0 µg m⁻³ (at 11:00 LST); this maximum was ∼7-fold the early morning NO₂ concentration. However, O₃ exhibited indistinctive variations during this period, and this was indicative of less NO₂ being generated from photochemical reactions. As mentioned above, there are no obvious anthropogenic emission sources around the MS site; therefore, our observations indicate the remarkable transport of pollutants from the lower ground surface to higher elevations during the daytime.

Figure 3 Diurnal variations in PM_2.5 and the boundary layer height (BLH) at the two sampling sites.

Moreover, the PM_2.5 concentrations at the MS site exhibited less nighttime variation, with a modest abatement (Fig. 3b). The nocturnal BLH usually remained below the elevation of the MS site; thus, the surface PM_2.5 may have contributed less to the aerosol levels at the MS site at night. To identify the potential spatial sources of nocturnal PM_2.5 at the MS site, a high-elevation CWT analysis was conducted. As illustrated in Fig. 4, the CWT values in the daylight hours were mostly concentrated over the sampling site, consistent with our discussion above. However, relatively high nighttime CWT loadings were distributed on Mt. Hua and in its surrounding regions, indicating that regional transport may be a major source of PM_2.5 at the MS site at night. Thus, the constituents and variations in nocturnal PM_2.5 at the MS site may mainly be the results of regional features. To verify the feasibility of vertical transport of air parcel, the WRF-Chem (Weather Research and Forecasting model coupled with chemistry) model was applied here to simulate wind filed and the divergence that represents the expansion rate of the air mass in a unit of time. From Fig. 5a, the southerly winds prevailed at mountain foot area during the whole campaign, which would blow the pollutants into the valley. And these pollutants at low-elevation can be transported to the upper layer by the updrafts, as indicated by the positive values of vertical divergence at the MF area that decreased with enhanced elevation (Fig. 5b). Besides that, we also analyzed the organic compounds in PM_2.5 samples, e.g., levoglucosan, BkF, and IP+BghiP, which are major tracers for the emissions from biomass burning, coal combustion, and vehicle exhausts, respectively (Wang et al., 2009; Wu et al., 2020a; Wang et al., 2007). From Fig. S2, the indistinctive divergences of diagnostic ratios and proportion of these organic tracers were found among both sampling sites, suggesting an insignificant change in the corresponding emission sources during the transport.

Figure 4 Concentration-weighted trajectory (CWT) analyses of PM_2.5 in the daytime (08:00–20:00) and nighttime (21:00–07:00) at the MS site.

Figure 5 The distribution of averaged diurnal divergence over the whole campaign, with the corresponding wind field. (a) Horizontal distribution at surface. (b) Longitude–pressure cross-sections at 34^∘29^′ N. Wind speeds were represented by arrows sizes, and the W component of wind vectors was magnified 10 times.

3.2 Characterization of water-soluble ions in PM_2.5

3.2.1 Comparisons of water-soluble ions between the two sites

Figure 6 shows the fractional contributions of the chemical compositions to the PM_2.5 at both sampling sites. As summarized in Table 1, the water-soluble ion level (WSI; 24.0±15.0 µg m⁻³) at the MF site was comparable to that of organic matter (OM; $\text{OM}=\mathrm{1.6}\times \text{OC}$; Wang et al., 2016), with a fractional contribution of ∼31 % to PM_2.5 (Fig. 6). At the MS site, the WSI exhibited lower values (19.5±16.0 µg m⁻³), yet the proportion was moderately enhanced by ∼6 %. Notably, this elevated contribution of WSIs was mostly attributed to sulfate and ammonium. Similar patterns in which the secondary inorganic ions (sulfate, nitrate, and ammonium; SNA) mass fraction increased with altitude within the mixing height have also been observed in Terni Valley (central Italy; Ferrero et al., 2012) and Salt Lake Valley (USA; Baasandorj et al., 2017). Among the SNA components, sulfate was the predominant species, exhibiting slight mass concentration differences between the two sampling sites (10.1±6.4 µg m⁻³ versus 9.0±7.1 µg m⁻³). However, a ∼4 % enhancement in the mass fraction of sulfate was measured at the MS site. Ammonium also exhibited a similar feature, accounting for ∼5 %–7.5 % of the PM_2.5. These sulfate and ammonium mass concentration homogeneities across the two sites were indicative of the further formation of these two ions during transport. Unlike sulfate and ammonium, nitrate and its proportions showed opposite trends, decreasing with elevation; this was consistent with most of the measured components. The above variation features of SNA among two sites were found for most of the time periods in the campaign, except for 12–13 September, which showed a higher SNA concentration at the MS site (Fig. 2e and f). On these 2 d, the MS site remained outside the boundary layers (a.s.l.; ∼550 m), suggesting less of an effect of the surface pollutants on the aerosol upper layers, while the precursor masses (∼12.3 µg m⁻³ for SO₂ and 8.4 µg m⁻³ for NO₂) were insufficient to form so much SNA at the MS site. Thus, the higher SNA aloft on the above 2 d may mostly be driven by regional or long-range transport. This can be verified by the CWT analysis, for which high loadings were mainly distributed in the western and southwestern areas of Mt. Hua (Fig. S3a), and the cities on the air mass transport pathways (e.g., Xi'an and Weinan) also suffered from moderate haze pollution on these 2 d (Fig. S3b). On account of the different sources of PM_2.5 between these 2 d and the remaining periods, the samples during 12–13 September were excluded in the following discussion, whereas the residual SNA data still exhibited similar variations, as mentioned above, and the divergence in nitrate mass concentration and fractional contribution to PM_2.5 among two sites even became more pronounced (Fig. S4). Moreover, distinct nitrate size distributions were also observed between the different sites in the summertime of 2020 (Fig. S5). From Fig. S5, we can note that the nitrate at a low elevation was enriched in the fine mode, with a minor peak in the coarse fraction. However, the high-elevation nitrate exhibited a bimodal pattern, with two equivalent peaks in the fine and coarse fractions, and was well correlated with coarse-mode calcium but poorly correlated with ammonium (R²=0.51, p<0.05). To our knowledge, ammonium nitrate, a major form of fine-mode particulate nitrate, can be easily volatilized and converted into gas-phase NH₃ and HNO₃ (Pakkanen, 1996; Harrison and Pio, 1983). Thus, the gaseous HNO₃ volatilized from fine PM may react with coarse-mode cations (e.g., Ca²⁺, Mg²⁺, and Na⁺) to form nonvolatile salts, leading to that significant nitrate shifts from fine particles to large particles. A similar phenomenon was also found in our previous study conducted at the summit of Mt. Hua (Wang et al., 2013). Nonvolatile sulfate was predominantly found in the fine fraction at both sampling sites, which may support this concept. More evidence for this hypothesis is presented below in Sect. 3.3.

Figure 6 Mass closure of PM_2.5 during the observed period ($\text{OM}=\mathrm{1.6}\times \text{OC}$).

The diurnal cycles of the 4 h sulfate, nitrate, and ammonium are illustrated in Fig. S6. As shown in Fig. S6, the total SNA concentration at the MF site exhibited a morning peak from 08:00–12:00 LST; this variation was quite different from that of PM_2.5. Such a difference between the total SNA and PM_2.5 at the MF site could partially be attributed to the lower sampling resolution and enhanced formation of SNA in the morning. The diurnal total SNA pattern identified at the MS site coincided with the PM_2.5 pattern, exhibiting a daily maximum reaching $\sim \mathrm{21.2}\pm \mathrm{19.9}$ µg m⁻³ (from 12:00–16:00 LST), which is a 1.2-fold increase compared to that measured at the MF site. Among the SNA components, morning peaks of nitrate and ammonium (from 08:00–12:00 LST) were also observed at the MF site. These nitrate and ammonium concentrations at the MF site can contribute to those at the MS site through vertical transport, leading to a significant enhancement in the nitrate and ammonium concentrations aloft, with afternoon peaks during 12:00–16:00 LST. Even so, the maximum nitrate concentration at the MS site (6.5±7.4 µg m⁻³) was still lower than that measured at the MF site (8.9±6.8 µg m⁻³) due to the NH₄NO₃ volatilization under the transport process, while ammonium exhibited the opposite trend. This finding was consistent with the above discussion. Unlike nitrate and ammonium, similar diurnal variations in sulfate were observed between the two sampling sites, with daily maxima observed from 12:00–16:00 LST at both sites. The major sulfate formation pathway during the daytime in summer is the photooxidation of SO₂ with an OH radical, and the formation rate facilitated by this process is much lower than that of the nitrate formation process (Seinfeld, 2016; Rodhe et al., 1981). Thus, sulfate formation may occur continuously during vertical transport, leading to smaller difference in the diurnal cycle of sulfate between the two sites.

3.2.2 Chemical forms of SNA at both sites

As shown in Fig. 6, the water-soluble ions considered here mainly included sulfate, nitrate, and ammonium, which usually exist in the form of ammonium salts (NH₄HSO₄, (NH₄)₂SO₄, NH₄NO₃, and so on). In the H₂SO₄-HNO₃-NH₃ thermodynamic system, H₂SO₄ and HNO₃ are neutralized by NH₃ under ammonia-rich conditions and mainly exist as (NH₄)₂SO₄ and NH₄NO₃ in aerosols. Conversely, H₂SO₄ is converted to ${\mathrm{HSO}}_{\mathrm{4}}^{-}$ in environments with relatively low NH₃ availabilities. Thus, NH₄HSO₄ and NH₄NO₃ may be the dominant aerosol components under such environmental conditions (Rodhe et al., 1981; Seinfeld, 2016). To reveal the major SNA forms at the different sampling sites considered herein, a theoretical ammonium concentration was calculated according to thermodynamic equilibrium with the atmospheric sulfate and nitrate levels. The theoretical ammonium levels were calculated as follows:

$$\begin{array}{}\text{(5)}& {\displaystyle }{{\mathrm{NH}}_{\mathrm{4}}^{+}}_{\mathrm{theory}}=\left({\displaystyle \frac{\left[{\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}\right]}{\mathrm{48}}}+{\displaystyle \frac{\left[{\mathrm{NO}}_{\mathrm{3}}^{-}\right]}{\mathrm{62}}}\right)\times \mathrm{18}\text{(6)}& {\displaystyle }{{\mathrm{NH}}_{\mathrm{4}}^{+}}_{\mathrm{theory}}=\left({\displaystyle \frac{\left[{\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}\right]}{\mathrm{96}}}+{\displaystyle \frac{\left[{\mathrm{NO}}_{\mathrm{3}}^{-}\right]}{\mathrm{62}}}\right)\times \mathrm{18},\end{array}$$

where $\left[{\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}\right]$ and $\left[{\mathrm{NO}}_{\mathrm{3}}^{-}\right]$ represent atmospheric concentrations (µg m⁻³). When (NH₄)₂SO₄ and NH₄NO₃ are the dominant species, the ${{\mathrm{NH}}_{\mathrm{4}}^{+}}_{\mathrm{theory}}$ can be calculated using Eq. (5). In contrast, Eq. (6) suggests that NH₄HSO₄ and NH₄NO₃ are abundantly present in the analyzed aerosols. Figure 7 compares the measured ${\mathrm{NH}}_{\mathrm{4}}^{+}$ concentrations with the theoretical ${\mathrm{NH}}_{\mathrm{4}}^{+}$ concentrations derived by the two equations above. As illustrated in Fig. 7a, the slope of the observational ${\mathrm{NH}}_{\mathrm{4}}^{+}$ values against the theoretical ${\mathrm{NH}}_{\mathrm{4}}^{+}$ values calculated using Eq. (6) was much closer to unit at the MF site than at the MS site, meaning that NH₄HSO₄ and NH₄NO₃ were the major chemical forms of SNA at the MF site. However, the opposite pattern was revealed at the MS site; thus, the upper aerosols were characterized by abundant (NH₄)₂SO₄ and NH₄NO₃. We also found that the diurnal variations in the SNA chemical forms in PM_2.5, which we mainly concerned in this study, were insignificant at both sampling sites. Based on observational data collected during the 2020 summertime, the NH₃ level at the MF site (36.0±68.0 ppb) was ∼9-fold that at the MS site (4.1±2.5 ppb). Under such an abundant NH₃ environment, the S(VI) was in the major form of NH₄HSO₄, but appeared as (NH₄)₂SO₄ in a relatively low NH₃ environment, which was somewhat unexpected. As can be inferred from earlier studies (Seinfeld, 2016; Shi et al., 1999), the NH₃ Henry's law coefficients generally increase in value as the temperature decreases. Therefore, the lower temperatures measured at the MS site would create a more favorable environment for ammonia, thus shifting its partitioning toward the particulate phase. The ${\mathrm{HSO}}_{\mathrm{4}}^{-}$ transported from the MF site would thus be further neutralized to ${\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}$ by this additional ammonium during transport, leading to the significant difference observed in the chemical forms of SNA between the two sites. Moreover, as the chemical component changed from NH₄HSO₄ to (NH₄)₂SO₄, the aerosol acidity moderately decreased, showing a higher bulk PM_2.5 pH (3.4±2.2) at relatively clean upper layer and a lower value (2.9±2.0) at heavily polluted grounds (Table 1). However, previous studies generally recognized that the aerosol would become more acidic when the air parcels were transported from polluted to cleaner/more remote regions (Liu et al., 1996; Nault et al., 2021). As shown in Table 1, the change in aerosol liquid water content (ALWC) has an indistinctive difference among both sampling sites (t test; p>0.05). Thus, we think that such a reduced aerosol acidity with increasing elevation in our study was mainly due to the change in the chemical component, which was caused by the different physicochemical behaviors of the semi-volatile species of nitrate and ammonium during transport. More discussion is included in the following section.

Figure 7 Comparison of the calculated and observed ${\mathrm{NH}}_{\mathrm{4}}^{+}$ concentrations at both sampling sites.

3.3 Physicochemical behaviors of nitrate and ammonium during transport

According to the above discussion, a conceptual model illustrating the physicochemical behaviors of nitrate and ammonium during vertical transport was proposed to explain the chemical composition differences between the two sites. As shown in Fig. 8, surface air parcels containing abundant NH₄HSO₄ and NH₄NO₃ particles can be transported to the upper atmosphere by the prevailing valley winds, and during this process, the volatile NH₄NO₃ is easily converted to gaseous NH₃ and HNO₃. Subsequently, heterogeneous reactions of the gaseous HNO₃ with fugitive dust occur, thus forming nonvolatile salts and resulting in the accumulation of nitrate on the coarse-mode particles. However, as the temperature decreased, the NH₃ that volatilized from the fine particles or was derived from the surface could re-enter the particulate phase through the gas–particle partition. Therefore, (NH₄)₂SO₄ would be formed in the aerosol phase and would gradually replace NH₄HSO₄.

Figure 8 Schematic of the physicochemical behaviors of nitrate and ammonium during the transport process.

To investigate the likelihood of NH₄NO₃ volatilization during the transport process, the dissociation constant of NH₄NO₃ (K_p) and the partial pressures of gas-phase NH₃ and HNO₃ were calculated in this study. More details regarding the calculation steps of the above factors can be found in Sect. 2.4. Based on the thermodynamic principles presented by Stelson and Seinfeld (1982), when the product of the partial pressures of NH₃ and HNO₃ ($P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{{\mathrm{NH}}_{\mathrm{3}}}$) is greater than K_p, then the equilibrium of the system shifts toward the aerosol phase, thus increasing the NH₄NO₃ formation. In contrast, a relatively low $P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{\mathrm{NH}\mathrm{3}}/K_{\mathrm{p}}$ value (<1) suggests that the NH₄NO₃ dissociation is induced and that NH₄NO₃ is transferred to the gas phase. Figure 9 depicts the ratio of the product of the partial pressures of NH₃ and HNO₃ with different ambient temperatures. As shown in Fig. 9, approximately 85 % of the samples collected at both sampling sites were located within the region with $P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{{\mathrm{NH}}_{\mathrm{3}}}/K_{\mathrm{p}}$ of less than 1, demonstrating a common NH₄NO₃ dissociation phenomenon during the observed period. For the samples with $P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{{\mathrm{NH}}_{\mathrm{3}}}/K_{\mathrm{p}}$ ratios <1, the mean value of the MS site ratios was approximately half that of the MF site ratios, indicating that the NH₄NO₃ dissociation may be more likely at higher elevations than that at lower elevations. This finding was inconsistent with the aircraft observations of the wildfire smoke plumes collected by Lindaas et al. (2021), who revealed that $P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{{\mathrm{NH}}_{\mathrm{3}}}/K_{\mathrm{p}}$ exhibited an increasing trend within 1–3 km (a.s.l.). As we already know, the abundant NH₃ and NO_x can be emitted by the wildfire, which would be transported aloft and lead to a higher NH₃ and HNO₃ mixing ratio compared to that at lower elevation. This may drive a higher $P_{{\mathrm{HNO}}_{\mathrm{3}}}\times P_{{\mathrm{NH}}_{\mathrm{3}}}/K_{\mathrm{p}}$ ratio at the upper layers of aircraft observations in the western U.S. (Lindaas et al., 2021).

Figure 9 Temperature dependence of the ratio of the product of the partial pressures of NH₃ and HNO₃ with the dry dissociation constant of NH₄NO₃.

Moreover, the nitrogen isotope compositions of nitrate and ammonium in PM_2.5 that can retain invaluable information regarding physicochemical processes (Wiedenhaus et al., 2021; Elliott et al., 2019) were thus measured to further verify the conceptual model. As previously mentioned, unlike daytime pollutants, nocturnal pollutants exhibited different sources between the two sampling sites. Thus, their nitrogen isotope compositions were more complicated and less comparable. However, for simplicity, only the daytime samples were analyzed here, based on the hypothesis that the sources of the high-elevation pollutants were the same as those of the pollutants collected at the MF site. As shown in Fig. 10, a discrepancy in the δ¹⁵N value of nitrate (δ¹⁵N-${\mathrm{NO}}_{\mathrm{3}}^{-}$) featuring more ¹⁵N-enriched ${\mathrm{NO}}_{\mathrm{3}}^{-}$ was observed at the MS site, with a p value of less than 0.05. This finding can be ascribed to the evaporation of a portion of the particulate NH₄NO₃ due to a dissociation shift in equilibrium; in this shift, the lighter ¹⁴N was preferentially incorporated into the atmosphere, leading to ¹⁵N enrichment in the remaining nitrate. A similar phenomenon was also revealed by Wiedenhaus et al. (2021), who thought that the ammonium nitrate dissociation may be an important reason for the accumulation of ¹⁵N in aerosol particles. Additionally, Freyer et al. (1993) revealed that gas-phase isotopic exchanges between NO and NO₂ result in the enrichment of the heavier ¹⁵N isotope in the more oxidized form and may further affect δ¹⁵N-${\mathrm{NO}}_{\mathrm{3}}^{-}$ through nitrate formation reactions. The above isotopic exchange between NO₂ and NO_x can be roughly described as follows: $[{\mathit{\delta }}^{\mathrm{15}}\mathrm{N}\left({\mathrm{NO}}_{\mathrm{2}}\right)-{\mathit{\delta }}^{\mathrm{15}}\mathrm{N}\left({\mathrm{NO}}_{\mathrm{x}}\right)]=(\mathrm{1}-K)\times (\mathrm{1}-f_{{\mathrm{NO}}_{\mathrm{2}}})$, where K and $f_{{\mathrm{NO}}_{\mathrm{2}}}$ are the temperature-dependent exchange constant and mole fraction of NO₂, respectively. Based on trace gas observations, the $f_{{\mathrm{NO}}_{\mathrm{2}}}$ values of the air aloft were very high due to the frequently undetectable NO concentration, indicating a rather limited isotopic exchange between NO₂ and NO. Therefore, the evaporation of particulate NH₄NO₃ has been a significant factor affecting the measurement of a higher δ¹⁵N-${\mathrm{NO}}_{\mathrm{3}}^{-}$ at the MS site than at the MF site in our observations. According to the above analysis, the ammonium at the MS site should theoretically be more and more enriched in δ¹⁵N with the continuous NH₄NO₃ volatilization. However, our observation of δ¹⁵N-${\mathrm{NH}}_{\mathrm{4}}^{+}$ did not correspond to the above pattern, namely the ammonium at the MS site depleted in δ¹⁵N compared to that at the MF site (p<0.05; Fig. 10). Given the unchanged NH₃ sources, as verified in Sect. 3.1.2, such seemingly unreasonable observations were mainly caused by the gas-to-particle conversion of ammonia. In this process, the reversible phase–equilibrium reactions between NH₃(g) and ${\mathrm{HNO}}_{\mathrm{3}}\left(\mathrm{g}\right)/\mathrm{HCl}\left(\mathrm{g}\right)$ would yield positive enrichment the in δ¹⁵N of aerosol ${\mathrm{NH}}_{\mathrm{4}}^{+}$ (Walters et al., 2019); nevertheless, unidirectional reactions involving NH₃(g) and ${\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}/{\mathrm{HSO}}_{\mathrm{4}}^{-}$ favored ¹⁵N depletion in the particle form, as revealed by Heaton et al. (1997). Thereby, the lower δ¹⁵N-${\mathrm{NH}}_{\mathrm{4}}^{+}$ values at the MS site were mostly driven by those irreversible reactions, rather than the reversible equilibrium ones. This result further confirmed our conjecture that the additional NH₃ would partition into particulate phases and further neutralize the acidic NH₄HSO₄, leading to an increasing pH at the MS site compared to that at the MF site. Taken together, this compelling evidence verifies that fine-mode nitrate and ammonium exhibit distinctly different physicochemical behaviors during their transport.

Figure 10 Nitrate and ammonium δ¹⁵N values at the two sampling sites in the daytime.

4 Conclusions and atmospheric implications

In this study, aerosol samples were collected at 4 h intervals on the mountainside of Mt. Hua, and the OC, EC, water-soluble ions and isotope compositions of nitrate and ammonium were measured and compared with simultaneous observations taken at a lower-elevation site (MF site). The particle mass at the MF site was approximately 1.5-fold that at the MS site, and distinctly different diurnal cycles were observed between the two sampling sites. Based on the BLH variation, we revealed that near-surface PM_2.5 could be transported to the upper layers by the mountain–valley breeze, leading to the gradual accumulation of pollutants on the mountainside during the daytime.

Sulfate, the predominant species found among ions at both sampling sites, exhibited nearly identical mass concentrations at the two sites but had a moderately enhanced mass fraction at the MS site. Such homogeneity was also observed in ammonium, which mainly existed as NH₄HSO₄+NH₄NO₃ and (NH₄)₂SO₄+NH₄NO₃ at the MF and MS sites, respectively. This observation indicated the further formation of ammonium during the transport process. Unlike sulfate and ammonium, nitrate at the MS site exhibited abated trends in both its concentration and proportion, mainly due to the volatilization of NH₄NO₃. With the help of nitrate and ammonium nitrogen isotopes, we proposed a conceptual model to illustrate the different behaviors of nitrate and ammonium during vertical transport; in this model, the semi-volatile NH₄NO₃ in surface air parcels was easily converted into gaseous NH₃ and HNO₃. Subsequently, heterogeneous reactions occurred between the gaseous HNO₃ and fugitive dust, forming nonvolatile salts and leading to a significant nitrate shift from fine particles to coarse particles. In addition, the decreasing temperature was favorable for NH₃ partitioning toward the particle phase, and the addition of ammonium further neutralized ${\mathrm{HSO}}_{\mathrm{4}}^{-}$ to form ${\mathrm{SO}}_{\mathrm{4}}^{\mathrm{2}-}$. This process would reduce the aerosol acidity, with bulk PM_2.5 pH increasing from 2.9±2.0 at the MF site to 3.4±2.2 at the MS site.

Over the past decade, the relative abundance of NH₄NO₃ has been enhanced in most urban areas of China because strict emission directives have been promulgated to abate the emission and environmental impacts of SO₂ (Xie et al., 2020; Song et al., 2019). In this work, we observed that NH₄NO₃ volatilization was a ubiquitous phenomenon for particles during transport, resulting in a shift in part-wise nitrate from the fine mode to the coarse fraction; this shift has also been reported in the offshore areas of the UK (Yeatman et al., 2001). Thus, we think that considering only fine-fraction nitrate may result in the conversion rate of NO_x to nitrate being partly underestimated at some times, especially in the summer. Moreover, the deposition velocity of coarse particles is usually faster than that of fine particles; therefore, the above process would appreciably elevate the deposition of N into the environment. Indeed, abundant NO₂, O₃, and NH₃ co-occurrence is common in the East Asian atmosphere, and under these conditions, secondary inorganic aerosols can be effectively produced, leading to a PM_2.5 loading explosion in the urban atmosphere of China (Wu et al., 2020c; Wang et al., 2016). Given this, harmonious reductions in NO₂, O₃, and NH₃ will be urgent in further mitigation strategies to improve air quality and alleviate other potential effects.