Introduction
Ammonia (NH3) has long been recognized as an important form of reactive
nitrogen (Nr) in the atmospheric environment, playing a key role in
biogeochemical cycles from atmospheric chemical processes to deposition and
in subsequent environmental impacts (e.g., air pollution, reduced biodiversity,
acidification and eutrophication) (Fowler et al., 2009; Sutton et al.,
2008). NH3 reacts with nitric and sulfuric acids in air, forming
secondary inorganic aerosols (e.g., NH4NO3,
(NH4)2SO4) with long atmospheric lifetimes that can transport
these species far from sources and contribute 40 %–57 % of the fine particle
matter in megacities (Fowler et al., 2009; Huang et al., 2014; Yang et
al., 2011). Therefore, NH3 has received increasing attention in air
pollution research (Wang et al., 2015). In addition to
agriculture, which is considered the largest global NH3 source,
emissions from biomass burning, industry, vehicles and other sources
(Galloway et al., 2003; Sutton et al., 2008; Erisman et al., 2008; Sun et
al., 2016, 2017) can also be significant.
In China, annual NH3 emissions were approximately 2 and 3 times higher
than European and US emissions, respectively, over the period from 1990 to 2005
(Reis et al., 2009; Kang et al., 2016; Zhao and Wang, 1994; Klimont,
2001; EMEP, 2018; USEPA, 2018), and were estimated to be 14.6 Tg N yr-1
in 2010 (Liu et al., 2013) and 15.6 Tg N yr-1 in 2015 (Zhang et al., 2017).
Such high emissions, in addition to the important role NH3 plays in
degrading air quality, makes NH3 a key target to curb serious air
pollution in Chinese urban areas (Fu et al., 2017; Chang et al., 2016; Ye
et al., 2011; Wang et al., 2011). Some studies have indicated that reducing
NH3 concentrations could be an effective method for alleviating
secondary inorganic PM2.5 pollution in China (Gu et al., 2014;
Wang et al., 2015; Wu et al., 2016; Xu et al., 2017). However, NH3 has
received less attention from the government than SO2 and
NOx, which have been controlled since 2005 and were effectively
reduced during the 12th Five-Year Plan period (2011–2015) (Fu et
al., 2017). Currently there are strong arguments regarding the role of regional
transport in contributing to haze pollution in China (Guo et al., 2014;
Li et al., 2015), especially for severe haze episodes occurring during
stagnant meteorological conditions with a shallow boundary layer (Sun et
al., 2014; Zheng et al., 2015; Quan et al., 2013). The vertical characterization
of air pollutant concentration profiles may be helpful for elucidating
factors contributing to the formation and transport of regional haze events
(Quan et al., 2013; Tang et al., 2015; Wiegner et al., 2006). Many
studies have been conducted to improve our understanding of temporal and
spatial concentration dynamics of atmospheric NH3 and how they relate
to underlying factors (e.g., emission intensity and meteorological conditions) and air quality (Yamamoto et al., 1988, 1995; Bari
et al., 2003; Vogt et al., 2005; Lee et al., 1999). However, such studies in
China have generally focused on the spatial distribution of NH3 near
the ground (Ianniello et al., 2010; Wu et al., 2009; Meng et al., 2011;
Xu et al., 2015), whereas the vertical characterization of NH3 concentrations
has been very limited.
NH3 mixing ratios may vary significantly as a function of height,
as NH3 is a trace gas with both point and non-point sources, and it also has a tendency
to deposit rapidly to surfaces. In urban locations, like Beijing,
where NH3 is a key contributor to fine particle formation, local
sources (e.g., traffic) emit at the surface and are then mixed through the
boundary layer, while NH3 transported from agricultural sources outside
the city is presumably already mixed through the boundary layer. The
influence of these behaviors may be reflected in the vertical NH3
concentration gradients measured within the city. For example, dominant
local surface traffic emissions might give rise to a profile that peaks near
the surface, while NH3 transported into the urban area may be uniformly
mixed in the vertical or even decline near the surface due to loss by dry
deposition. Of course these patterns are expected to be further affected by
sinks, including surface deposition as well as by the fine particle formation of
ammonium salts. NH3 vertical distribution measurements are also useful
for advancing satellite retrievals, which offer a great potential for
understanding the global distribution of gaseous NH3 (Shephard and
Cady-Pereira, 2015; Sun et al., 2015; Van Damme et al., 2015).
To our knowledge there are few studies reporting long-term observations of
the vertical distributions of NH3 in the lowest few hundred meters of the
atmosphere, including measurements at the BAO tower in the USA (Li et
al., 2017; Tevlin et al., 2017) and the CESAR site in the Netherlands (Dammers
et al., 2017). Li et al. (2017) analyzed vertical NH3 concentration
profiles at the BAO tower in Colorado, USA, reporting the minimum
concentration at the top of the tower, which slowly increased towards a peak
concentration at ∼10 m before a large reduction in
concentration was found at 1 m. The site was influenced by the transport of high NH3
concentrations from large animal feeding operations to the northeast.
Using higher time resolution measurements at the BAO tower, Tevlin
et al. (2017) pointed out that the surface can act as an occasional NH3
sink as well as a source. The CESAR study in the Netherlands showed that
vertical profile differences were mainly due to local and regional transport
influences (Dammers et al., 2017). Because the BAO and CESAR tower
sites are both located in suburban areas with low aerosol mass loadings,
observed vertical profiles of aerosol and gas species (Öztürk et
al., 2013; VandenBoer et al., 2013; Riedel et al., 2013) could be
substantially different from those in megacities in China. Zhou et
al. (2017) measured vertical concentration profiles of NH3 and 7
other air pollutants at 10 heights (8, 15, 47, 80, 120, 160, 200, 240, 280 and
320 m) in urban Beijing, finding that NH3 concentrations peaked at 160 m.
However, only one vertical profile was measured and may not adequately
represent typical conditions. Until now, long-term monitoring of vertical
NH3 concentration profiles has not been carried out in China.
Here, we report a 1-year field campaign on the Beijing 325 m
meteorological tower to investigate vertical NH3 concentration profiles
and consider how temporal variations may relate to urban emission sources,
meteorological factors and air transport from more distant sources. Study
findings are relevant for our understanding of precursor NH3
distributions and the role of NH3 in the formation of severe aerosol
pollution in China; furthermore, they will provide benchmarks to assist in meeting air
quality goals and policy needs in future.
(a) Modeled NH3 emissions distribution (0.1∘,
∼10 km) over the North China Plain in 2015 including the location of the monitoring
site shown as a black dot. NH3 emission estimates are from the
inventory of Zhang et al. (2018) at a 0.1∘ horizontal resolution.
(b) Map of Beijing showing the location of the monitoring tower.
(c) The 325 m meteorological tower and ALPHA passive samplers.
Materials and methods
Site description
The sampling site is located at the State Key Laboratory of Atmospheric
Boundary Layer Physics and Atmospheric Chemistry (LAPC), Institute of
Atmospheric Physics (IAP), Chinese Academy of Sciences (CAS) in urban
Beijing (39∘58′ N, 116∘22′ E; Fig. 1).
The site is approximately 0.8 km north of the Third Ring Road, 1.3 km south
of the Fourth Ring Road and 0.2 km west of the Beijing–Tibet expressway,
which are three transport arteries encircling Beijing, each with average
traffic volumes of over 200 000 vehicles day per day in 2016 (Beijing
Transport Institute, 2017); therefore, this site represents a typical urban site that is mainly surrounded
by residential areas.
NH3 measurement
From 16 March 2016 to 16 March 2017, weekly atmospheric NH3 samples
were collected at 16 heights on the 325 m meteorological tower using ALPHA
passive samplers (adapted low-cost high absorption, Centre for Ecology and
Hydrology, Edinburgh, UK) except for a few samples with slightly different
durations due to tower maintenance schedules. The samplers operate on the
principle of diffusion using an acid-coated filter to capture the NH3.
A PTFE (Teflon) membrane is placed directly at the mouth of the sampler,
forming a quiescent boundary layer in front of the sample membrane. A
stable, turbulent-free diffusion path length is achieved behind the
membrane, whilst allowing gaseous NH3 to diffuse through for capture
and minimizing the sampling of NH4+ aerosol (Tang et al., 2014).
NH3 was sampled at 2, 8, 15, 32, 47, 63, 80, 102, 120, 140, 160, 180,
200, 240, 280 and 320 m a.g.l. (above ground level). At each height, three ALPHA
samplers were deployed under a PVC shelter to protect the samplers from rain
and direct sunlight (shown in Fig. 1). NH3 samples were
extracted with 10 mL high-purity water (18.2 MΩ-cm) and analyzed
using a continuous-flow analyzer (Seal AA3, Germany). Three field (travel) blanks were prepared for each batch of samples, which were analyzed together with
the abovementioned samples, and used to blank correct sample results and determine the method
detection limit (MDL) values. MDL was calculated using the following equation:
MDL ≥t×Sb×N1+N2N1×N2,
where the t value is given at the 95 % confidence level for the appropriate of
degrees of freedom, Sb is the blank standard deviation, N1 and
N2 are the number of sample measurements (single measurement,
N1=1) and the number of analyzed blanks, respectively.
From the field blanks, the MDL was calculated to be 0.31 µg m-3 for a 1-week ALPHA passive
NH3 sample. All lab measurements were conducted in the Key Laboratory
of Plant-Soil Interactions, Chinese Ministry of Education, China
Agricultural University. More details regarding the passive samplers and the related
laboratory preparation and analysis can be found in Xu et al. (2015).
Meteorological data
Meteorological parameters, including wind speed (WS), wind direction (WD),
relative humidity (RH) and temperature (T), were obtained at all sampling
heights except 2 m; the temperature was also not available at 8 m. WS and WD
were measured using four-cup anemometers (model O1OC, Met One Instruments),
and RH and T were measured using a T/RH sensor (model HC2-S3, ROTRONIC).
Time series of the vertical distribution of weekly atmospheric NH3
concentrations (±1σ) in Beijing urban (16 March 2016–16 March 2017).
Data analysis
Repeated-measures analysis of variance (ANOVA) was used to test changes in
the NH3 concentration along vertical profiles. When the ANOVA results were
significant, the Tukey's honest significant difference (HSD) test was used
to determine the significance of the difference between means with a
significance level of P<0.05. The coefficient of determination was
used to test the linear correlations with a significance level of P<0.05.
All of the statistical analyses were conducted using SPSS version 23.0
(IBM Corp., Armonk, NY, USA).
Potential source contribution function analysis (PSCF) (Ashbaugh et
al., 1985) of atmospheric NH3 was performed using MeteoInfo (TrajStat
package) (Wang, 2014), where 72 h back trajectories arriving at the
monitoring site (IAP tower) at each height were calculated every 3 h for the
entire study period. The average NH3 concentration for each cluster was
computed using the cluster statistics function. NH3 pathways could then
be associated with the high concentration clusters. The number of trajectory
segment endpoints falling in a grid cell (i, j) is nij. The number of
trajectory endpoints associated with the data with NH3 concentrations higher than an arbitrarily set criterion for each
height during the four seasons (75th percentile for NH3 was set
here) is mij (Table S1 in the Supplement). The PSCF value for the ijth cell is then calculated as mij/nij. A weighting function Wij was
applied to reduce the uncertainties of small values of nij
(Polissar et al., 1999). Weighted PSCF values (WPSCF) were calculated
by multiplying a particular Wij (≤1.00) if the total number of the
endpoints for one grid cell was lower than 3 times the average of the
endpoints per each cell. Higher WPSCF values indicate higher potential
contributions of NH3 to the receptor site (IAP tower).
Wij=1.0080<nij0.7020<nij≤800.4210<nij≤200.05nij≤10
Results
Vertical profiles of NH3 concentrations
The time series of weekly averages of NH3 concentrations from
16 March 2016 to 16 March 2017 are shown in Fig. 2. The weekly NH3
concentration across all heights averaged 13.3±4.8 µg m-3
during the year-long study period. Individual weekly concentrations ranged
from 4.4 µg m-3 at 2 m to 25.3 µg m-3 at 32 m.
Nearly all (99.6 %) of the weekly NH3 concentrations along the
profile exceeded 5 µg m-3. Summer concentrations were
generally the highest. Maximum NH3 concentrations mostly occurred
between 32 and 63 m, decreasing both towards the surface and the top of
the tower. Minimum concentrations mostly occurred at 2 and 320 m (Fig. S1 in the Supplement).
Significant differences of annual average NH3 concentrations
across the vertical profile were only found between the “maximum
concentration” height and the top two heights, i.e., 280 and 320 m (Fig. 3i).
Even at 320 m, the annual average NH3 concentration was still
relatively high at 11.3 µg m-3 (Fig. 3i). During the whole
observation period, the daily average boundary layer height was generally
above 320 m, indicating that a good portion of the sampling occurred within a
well-mixed boundary layer (Fig. S2).
Seasonal vertical concentration profiles exhibited fairly similar shapes to
the annual average profile, although there were some important differences in absolute
concentration values and the magnitude of vertical gradients within the
profiles (Fig. 3). The average NH3 concentration across the profile
from high to low was observed in summer (18.2 µg m-3), spring
(13.4 µg m-3), autumn (12.1 µg m-3) and winter
(8.3 µg m-3). Proportional declines of the NH3 concentration from
the peak to higher and lower elevations differed between seasons: the
greatest proportional decline was seen in autumn (28.1 % decrease from 63 to 320 m), followed by winter
(23.8 %), summer (20.5 %) and spring (15.8 %) (Fig. S3).
Comparison of seasonal vertical NH3 concentrations with the
mean (dots), median, 10th, 25th, 75th and 90th percentiles of the NH3
concentrations of each height for the IAP tower (Beijing, this study; a, c, e, g, i) and BAO tower (USA, Li et al., 2017; b, d, f, h, j).
The lowercase letters next to the boxes denote the statistical difference
in the NH3 concentration between all heights, where a one-way ANOVA was
used, at the p<0.05 level.
Meteorological variability
Vertical NH3 concentration profiles varied substantially during the
sampling period, along with vertical changes in meteorological parameters.
Bivariate polar plots (Fig. 4) show that high NH3 concentrations below
47 m were mostly observed during periods with low wind speeds (<4 m s-1).
As heights and associated wind speeds increased, the relationship
between NH3 concentrations and wind speed weakened. For example, at
280 m, the highest concentration was observed when the wind speed was also high
(up to an average of ∼15 m s-1).
Wind direction also plays an important role in air pollution transport. Transport
from the northwest was typically associated with low NH3 concentrations
at all heights, consistent with the absence of large emissions sources in
the mountains northwest of Beijing. It is noteworthy that high NH3
concentrations at near-surface heights (8 and 15 m) always coincide with
winds from the south, including the southeast and southwest directions. High
NH3 concentrations appear to be associated with winds from the
northeast from 32 m to 80 m. Above 80 m, winds from the south contribute
more to high NH3 concentrations. Major regions of agricultural NH3
emissions are located south and east of Beijing.
The frequency distributions of wind directions and NH3
concentration for all height during the observation period.
Radial data are WS (m s-1) as a function of WD (∘). The colors
denote the NH3 concentrations (µg m-3).
Probability density of NH3 concentrations (µg m-3)
at different ranges of temperature1 (∘C) and relative humidity2 (%)
for 14 heights. 1 Temperature includes four subsets: <4, 4–12, 12–20 and
>20∘C; 2 relative humidity includes four subsets: <25 %, 25–50 %,
50–75 % and >75 %.
To further investigate observed variability, we show the probability density
function of NH3 concentrations in relation to the relative humidity (RH)
and temperature (T) (Fig. 5). Clear positive relationships between T and
NH3 concentrations were found at all heights from low RH to high RH.
When T was low (T<12 ∘C), the NH3 concentration mostly fell
below 10 µg m-3 under all RH conditions. The occurrence of
high NH3 concentrations increased with T>12 ∘C, which
is not surprising given that agricultural NH3 emissions increase
with T; furthermore, higher T and lower RH also shift the equilibrium of the
NH3(gas) + HNO3(gas) ↔ NH4NO3(particulate) system toward the
gas phase. Statistically, a strong positive relationship was found between
NH3 and T at all heights from the surface to the top of the tower
(R2∼0.6; Fig. S4); both the slope and the correlation
coefficients were similar across all heights. Although, a positive
correlation between NH3 and RH and a negative correlation between
NH3 and WS were found, the correlation coefficients were quite low.
Weighted potential source contribution analysis (WPSCF) of atmospheric
NH3 in Beijing from 16 March 2016 to 16 March 2017.
Potential source analysis
Analysis of the relationship between local wind direction and NH3
concentrations does not fully clarify the potential source regions
contributing to observed NH3 at the sampling site (Fig. S6). Some
seasonal variations were observed, i.e., the frequency of high NH3
concentrations were greater under southerly winds than northwesterly winds in
the spring, the increased frequency of high NH3 concentrations were
associated with southerly and easterly winds in the summer and autumn, and
NH3 concentrations still exceeded 5 µg m-3 during winter
with relatively frequent winds from the northwest.
To examine the relationship between air transport and NH3
concentrations more rigorously, weighted PSCF (WPSCF) during the four
seasons were calculated for several measurement heights (2, 63, 180 and 320 m)
(Fig. 6). In summer, from the surface to the tower top, a strong influence
from source areas to the south of Beijing was seen, coinciding with regions
(e.g., Tianjin, Henan, Hebei and Shandong provinces) characterized by
elevated anthropogenic emissions of NH3 (Fig. 1), largely from
agricultural activities (Zhang et al., 2009; Gu et al., 2012). During
summer, regions to the north and west of the monitoring site had low WPSCF
values, whereas high WPSCF values to the south and southeast were common during
spring. High WPSCF values were mainly located northwest and southeast of
Beijing in autumn, while their WPSCF values were typically lower in winter
than during other seasons.
It is important to remember that aerosol–gas partitioning can also strongly
influence measured NH3 concentrations. To investigate seasonal phase
changes between NH3 and NH4+, we define the NH3 gas
fraction (FNH3 = the gaseous NH3 concentration divided by the
sum of the gaseous NH3 and fine particulate NH4+
concentrations), where the concentrations are expressed in molar units.
Monthly average partitioning for these reduced inorganic nitrogen forms from
a nearby urban monitoring site, 10 km from the IAP tower, is plotted in
Fig. S8. The NH3 gas fraction (FNH3) was found to be the highest
in summer (0.83 in August) and the lowest in winter (0.36 in February). As
expected, gas phase NH3 is favored in the warmer months, while particle
phase NH4+ is favored in the cooler months, with a gradual
transition. Weekly NH4+ concentrations at the tower were estimated
using weekly NH3 concentrations divided by monthly FNH3,
and WPSCF analysis of the sum of NH3 + NH4+ was then performed (see
results in Fig. S9). Results of this total WPSCF (NH3 + NH4+)
analysis yielded similar patterns to the NH3 WPSCF analysis
for all heights and seasons, indicating the importance of the identified
source regions for both the gaseous and particulate atmospheric forms of emitted NH3.
Overview of measured vertical NH3 concentrations
(µg m-3) in previous studies and in this study.
Heights (m)/
The Netherlands
BAO tower, USA
IAP tower, Beijing
NH3
Rural area
Meteorological tower
(µg m-3)
0–5
6.8 (1 m)
8.3
4.7
–
12.5
6.5 (4 m)
5–10
–
–
5.0
7.9
13.4
10–20
9.6
–
–
15.8
13.8
20–40
–
6.2
4.61
–
14.2
40–60
–
–
4.19
12.8
14.1
60–80
–
–
–
12.5 (80 m)
14.3 (63 m)
14.2 (80 m)
80–100
–
3.6
3.6
–
13.9
100–150
–
–
3.09
12.4 (120 m)
14.0 (120 m)
13.8 (140 m)
150–200
4.5
2.1
2.72
14.0 (160 m)
13.5 (160 m)
6.7 (200 m)
13.3 (180 m)
12.7 (200 m)
200–250
–
–
2.39
9.1
12.1
250–300
–
–
2.25
7.3
11.8
300–350
–
–
–
7.6
11.3
Period
2014
13 Dec 2011–9 Jan 2013
10–25 Feb 2009
16 Mar 2016–16 Mar 2017
References
Dammers et al. (2017)
Erisman et al. (1988)
Li et al. (2017)
Zhou et al. (2017)
This study
Discussion
Vertical NH3 concentration profiles
The North China Plain is a well-known “hotspot” for NH3 emissions due
to the rapid development of industrialization, urbanization and intensive
agriculture (Kang et al., 2016; Y. Zhang et al., 2010). In our study, high
atmospheric NH3 concentrations (13.3±4.8 µg m-3)
were found up to 320 m a.g.l. in urban Beijing
(16 March 2016–16 March 2017), and were much higher than the average annual NH3
concentration (3.3±1.4 µg m-3) observed across a vertical
profile at the 300 m rural BAO tower, USA (Li et al., 2017). Some studies of
NH3 vertical distribution found that the NH3 concentration
decreased significantly with height. For example, Tevlin et al. (2017)
reported an overall increase in summertime NH3 mixing ratios
toward the surface of 6.7 ppb or 5.1 µg m-3 (89 %) during the
day and 3.9 ppb or 3.0 µg m-3 (141 %) at night. In the BAO
tower study (Li et al., 2017), which also measured concentrations using
passive (Radiello) samplers deployed for 1- to 2-week sample periods, the
concentration profiles showed a similar overall vertical distribution: the
minimum NH3 concentration was observed at the top of the tower, it slowly increased towards a
peak concentration at ∼10 m, and a sharp reduction was then seen near the
surface. By contrast, our results showed much smaller decreases in NH3
concentrations in the upper air in urban Beijing (Table 1), with only a
1.18 µg m-3 (9.5 %) average decrease from the surface to the
top of the tower (Fig. 3i). The flatter shape of the Beijing vertical profile may reflect a
combination of strong local (e.g., vehicle) and regional (e.g., industrial and
agricultural emissions) sources (Figs. 2 and 6) in our study, the fact
that deep mixing layers regularly enveloped the full height of the tower
within the surface boundary layer so that all sources influencing the tower
measurements were vertically well mixed (Fig. S2), and/or the averaging of
more distinct profiles over the week-long sample periods. In contrast to the
“rural” boundary layer above the fields surrounding the BAO tower, the
mixing in the Beijing urban area could be greatly enhanced by larger surface
roughness (e.g., the average urban building height is ∼50 m) and surface
heating (Baklanov and Kuchin, 2004). Higher time resolution vertical profile
measurements are needed in the future to untangle the influence of these
potential factors.
Distinct seasonal variations in NH3 concentrations were found (Fig. 2),
which were statistically most strongly associated with temperature rather than relative
humidity or wind speed (Fig. S4). High temperatures enhance NH3
emissions from soil, applied fertilizers, animal waste, vertical mixing and
increase volatilization of NH3 from NH4NO3 particulate matter
(Bari et al., 2003; Ianniello et al., 2010; Li et al., 2014; Lin et al.,
2006; Meng et al., 2011; Plessow et al., 2005; Walker et al., 2004;
Zbieranowski and Aherne, 2012). While high (low) mixed-layer heights in
spring and summer (autumn and winter) could dilute (concentrate) NH3 in
the surface boundary layer (Fig. S3), average NH3 concentrations across
the profile were actually high in summer/spring and low in winter/autumn,
consistent with the strong temperature-driven seasonal variation of the NH3
concentration and the greater NH4NO3 particle formation during
cold periods in autumn and winter. Conducting simultaneous measurements of
fine particle composition at different heights in future studies would be
valuable for more closely evaluating the influence of changes in
phase-partitioning.
Li et al. (2017) found a vertical difference of approximately 75 %
from the concentration peak near the surface to the top of the BAO tower in
winter (Fig. 3j), and attributed this strong vertical gradient to the
occurrence of low level temperature inversions which trapped emissions
closer to the surface during this period. During our study in Beijing, the vertical
gradient was only 28 % in winter (maximum concentration found at 32 m),
consistent with a deeper average boundary layer. However, inversions did
limit the vertical mixing of NH3 during some periods in Beijing.
Examination of the thermal inversion layer probability at 06:00 and 15:00 LT
(Fig. S7b and c) revealed that T inversions (0.22±0.26 ∘C)
frequently occurred between 102 and 160 m. Consequently, persistent higher
NH3 concentrations begin at a lower altitude (Fig. S7a) as also
observed by Tevlin et al. (2017). Because the time resolution of our
Beijing study was one sample per week, we could not catch the changes
between the daytime and nighttime NH3 vertical mixing. Compared to
NH3 monitoring in real time (Tevlin et al., 2017), weekly sampling
smooths diurnal vertical distributions and makes it harder to identify the
influence of local surface sources or sinks.
Surfaces can act either as sources or sinks of NH3, depending on
the surface NH3 content, ambient NH3 concentrations, and local
meteorology and surface type (Tevlin et al., 2017; L. Zhang et al.,
2010 ). The maximum NH3 concentration occurrence at 2 m in
Beijing and the concentration decrease with increased height may reflect an
important surface source of NH3, although our limited time resolution
makes such conclusions tentative. The influence of the evaporation of
dew/precipitation may also be important. Some studies found that dew is both
a significant nighttime reservoir/sink and strong morning source of
NH3 (Wentworth et al., 2016; Teng et al., 2017).
Potential source analysis
Areas south of Beijing with high WPSCF values appear to be important
NH3 source regions (Fig. 6), suggesting regional transport from high
agricultural NH3 emission areas (e.g., Hebei, Henan, Shandong provinces) contributed significantly to atmospheric NH3 in the Beijing urban
region. Consistently higher NH3 concentrations were observed during
periods with winds from the southeast, south and southwest at all heights, especially in
summer (Fig. S6). Although NH3 has a limited atmospheric lifetime with
respect to dry deposition, concentrations in these agricultural NH3
source regions can be extremely high (Shen et al., 2011) while
significant NH3 can be tied up in longer-lived ammonium nitrate
particles that partially dissociate to release NH3 back to the gas
phase in response to NH3 loss by dry deposition (Ianniello et
al., 2011; Kang et al., 2016; Xu et al., 2017). The WPSCF (Fig. 6) and
NH3 emissions distribution (Fig. 1a) both suggest the importance
not only of regional transport from nearby areas, but also the potential for
local emissions to play an important role in sustaining the high NH3
level in Beijing, e.g., vehicular traffic (Chang et al., 2016; Pan et al.,
2018a). As discussed above, stagnant meteorological conditions with low WS
and T inversions allow local emissions, such as those from urban traffic, to
accumulate. Additionally, the topography of the mountains to the west and
north of Beijing effectively traps polluted air over Beijing during
southerly airflow, an effect reported in many Beijing particulate matter
studies (Xia et al., 2016; Wu et al., 2009; Zhao et al., 2009).
Generally, NH3 source regions identified in the WPSCF analysis (Fig. 6)
suggest that regional transport from the south exerts an important
influence on Beijing NH3 concentrations throughout the year. The area
south of Beijing (e.g., Hebei, Henan and Shandong provinces) is a hotspot of
NH3 emission (Zhang et al., 2018), and half of NH3 emissions
have been estimated to deposit as NH3 at urban sites in the North China
Plain (Pan et al., 2018b). In addition, seasonal patterns of NH3
potential sources (Fig. 6) matched well with the seasonal surface NH3
concentrations in China (Zhang et al., 2018). In detail, NH3
concentrations were typically highest in summer, and south winds produced
higher NH3 concentrations than other wind directions (Fig. S6). Spring
and summer had a similar wind direction distribution (Fig. S6) and wind
speeds (Fig. S5), but corresponding NH3 concentrations were lower in
spring. This may reflect decreased emissions in regions to the south during
cooler spring temperatures and the increased partitioning of NH3 into fine
particles during this cooler season. As shown above aerosol–gas partitioning
strongly influences NH3 concentrations; high FNH3 during warm
periods, especially summer, favored greater NH3 gas concentrations due
to the thermodynamic tendency for NH4NO3 to dissociate to NH3
and HNO3 at high temperatures. Although FNH3 was low in winter,
indicating that NH4+ is the dominant NHx form in this cold season,
winter NH3 concentrations across all heights still averaged
8.3±2.6 µg m-3, with a similar wind direction distribution as other
seasons, except at high altitudes (i.e., 240 and 320 m; Fig. S6).