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atmosphere
Article
Characteristics and Formation Mechanisms of Fine
Particulate Nitrate in Typical Urban Areas in China
Xinlei Ge 1, *, Yanan He 1, Yele Sun 2, Jianzhong Xu 3, Junfeng Wang 1, Yafei Shen 1and
Mindong Chen 1
1Jiangsu Key Laboratory of Atmospheric Environment Monitoring and Pollution Control (AEMPC),
Collaborative Innovation Center of Atmospheric Environment and Equipment Technology (CIC-AEET),
School of Environmental Science and Engineering, Nanjing University of Information Science & Technology,
Nanjing 210044, China; 13057589806@163.com (Y.H.); Njwangjunfeng@gmail.com (J.W.);
shen1225@nuist.edu.cn (Y.S.); chenmd@nuist.edu.cn (M.C.)
2State Key Laboratory of Atmospheric Boundary Layer Physics and Atmospheric Chemistry,
Institute of Atmospheric Physics, Chinese Academy of Sciences, Beijing 100029, China;
sunyele@mail.iap.ac.cn
3State Key Laboratory of Cryospheric Sciences, Northwest Institute of Eco-Environment and Resources,
Chinese Academy of Sciences, Lanzhou 730000, China; jzxu@lzb.ac.cn
*Correspondence: caxinra@163.com; Tel.: +86-25-5873-1394
Academic Editor: Guey-Rong Sheu
Received: 17 January 2017; Accepted: 20 March 2017; Published: 22 March 2017
Abstract:
Nitrate is a very important aerosol component, thus elucidation of its characteristics
and formation mechanisms is essential and important for effective reduction of aerosol pollution.
In this work, highly time-resolved submicron aerosol (PM
1
) data measured by Aerodyne aerosol
mass spectrometers (AMS) in Nanjing, Beijing and Lanzhou during both summer and winter
were integrated to investigate the nitrate behaviors in urban China air. Results showed that
nitrate occupied 1/8–1/4 of PM
1
mass, typically higher than those observed in rural/remote
regions. Relative mass fractions of nitrate also varied significantly at different pollution levels.
Nitrate mass fractions generally increased with the increase of PM1loadings during summer, while
the contributions during winter increased first and then decreased with the increase of pollution
levels. We further propose that there are at least three mechanisms that likely govern the urban nitrate
behaviors: Type I—thermodynamics driven, Type II—photochemistry driven, and Type III—planetary
boundary layer (PBL) dynamics driven. Analyses of the ammonium-sulfate-nitrate data revealed
that ammonium nitrate was able to form before sulfuric acid was fully neutralized in some urban
areas. Our findings provide useful insights into the characterization and reduction of fine particulate
nitrate pollution.
Keywords:
fine aerosols; nitrate; aerosol mass spectrometer; diurnal variation; formation mechanism
1. Introduction
Fine particulate matter (PM
2.5
) pollution is a very serious and pressing environmental issue
in the densely populated areas of China [
1
–
3
]. PM
2.5
also has profound effects on the visibility, air
quality, human health, and the Earth’s climate, etc. Nitrate is one of the major components of PM
2.5
,
which can account for 4.5%–25% of the PM
2.5
mass, depending upon locations and environments [
4
–
7
].
Previous studies also showed that nitrate contributions to PM
2.5
mass could increase with the increase
of PM
2.5
mass loadings [
8
]. Indeed, nitrate was found to be an very important contributor to the
high PM
2.5
concentrations during heavy haze events in China [
9
]. In this regard, elucidation of the
atmospheric behaviors of nitrate is critical to reduce the PM2.5 pollution.
Atmosphere 2017,8, 62; doi:10.3390/atmos8030062 www.mdpi.com/journal/atmosphere
Atmosphere 2017,8, 62 2 of 12
Nitrate may associate with different cations in the particle phase. Under marine environments,
it may be partially present as NaNO
3
due to the reaction: HNO
3
+ NaCl = NaNO
3
+ HCl, but this
reaction mainly affects the coarse particles [
10
] rather than fine particles (PM
2.5
). Typically, fine aerosol
nitrate exists in the form of ammonium nitrate, which is generated by the neutralization reaction:
HNO
3
+ NH
3
= NH
4
NO
3
. The precursor nitric acid (HNO
3
) is overwhelmingly produced by secondary
oxidation processes, and only a very small portion is emitted directly, for example, from volcano
eruptions [
11
]. The secondary formation of HNO
3
includes two major pathways. During daytime,
HNO
3
is mainly produced from OH-initiated oxidation of NO
2
[
12
]. During nighttime, NO
2
can
be oxidized by O
3
to form NO
3
radical, which then reacts with NO
2
to form N
2
O
5
[
13
]; N
2
O
5
then
heterogeneously reacts with water vapor on the particle surface and produces HNO
3
(this process is
favored by low temperature and high relative humidity) [
14
,
15
]. Even when HNO
3
is available, the
neutralization of HNO
3
with NH
3
can only occur efficiently under ammonium-rich conditions, as
NH
3
prefers to react with sulfuric acid (H
2
SO
4
) first [
16
]. Some earlier studies [
17
,
18
] demonstrate that
a molar ratio of NH
4+
/SO
42−
larger than 1.5 can be treated as an ammonium-rich condition, meaning
that NH
4
NO
3
can be produced effectively after the formation of (NH
4
)
3
H(SO
4
)
2
. Nevertheless, this
boundary value (1.5) may change as both homogenous and heterogeneous formations of NH
4
NO
3
are
possible in the real atmosphere [
19
], and the dominant pathway may vary at different locations, or
during different seasons in the same location [20].
Traditionally, atmospheric particles are collected onto filters, and then nitrate concentration is
determined by using ion chromatography (IC). Such nitrate data has low time resolution, as the filters
are collected every few hours or days. From the year 2000, Aerodyne aerosol mass spectrometers
(AMS) [
21
,
22
], which can determine the aerosol compositions quickly and in real-time, were used
widely around the world. Due to the relatively simple fragmentation pattern in the AMS ionization
scheme, ammonium nitrate is recommended as a standard mass calibration material for all versions
of AMS, and the AMS is able to quantify nitrate relatively well. Given that the AMS also provides
highly-time resolved data, it is particularly useful for investigating the temporal variations and diurnal
cycles of nitrate. In this study, we integrated and analyzed the AMS data collected from a few typical
cities in China (Nanjing, Beijing, and Lanzhou) during both summer and winter. The purpose of this
work is to provide an overview of the characteristics of ambient aerosol nitrate in typical urban areas of
China, and propose the dominant mechanisms that govern such nitrate characteristics under different
atmospheric environments. Our findings can help researchers interpret the nitrate behaviors in future
studies, and provide useful insights into the effective reduction of particulate nitrate pollution.
2. Experiments
In this work, we selected Nanjing, Beijing, and Lanzhou—three cities representing Yangtze River
Delta (YRD), Beijing-Tianjin-Hebei (Jing-Jin-Ji), and northwestern part of China, respectively—for our
sampling locations. The sampling locations, sampling periods, and the versions of Aerodyne AMS used
for these field campaigns are shown in Figure 1. For Nanjing, a soot-particle AMS (SP-AMS) [
23
,
24
]
was deployed in urban Nanjing (Jiangxin Island) from 11 August to 18 September 2014, and in the
campus of Nanjing University of Information Science and Technology (NUIST) from 20 February to
23 March 2015 (details in [
25
]). For Beijing, an Aerosol Chemical Speciation Monitor [
26
] (ACSM,
a simplified version of Aerodyne AMS) was deployed during summer (26 June–28 August 2011) and
winter (21 November 2011–20 January 2012), both at the Institute of Atmospheric Physics (IAP) of
Chinese Academy of Sciences (CAS). For Lanzhou, the measurements were performed by using a high
resolution time-of-flight AMS (HR-ToF-AMS) [
27
] in the Cold and Arid Regions Environmental and
Engineering Research Institute (CAREERI) during summer (11 July–7 August 2012) and at Lanzhou
University during winter (10 January–4 February 2014) (the two sites are only about 500 meters apart).
Note that the results of the Beijing summer data have been published in Sun et al. [
28
], and winter data
was presented in Sun et al. [
29
,
30
], while results of Lanzhou summer and winter were presented in
Xu et al. [31]
and Xu et al. [
32
], respectively. The nitrate data from Beijing and Lanzhou were re-visited
Atmosphere 2017,8, 62 3 of 12
here, together with the new Nanjing data, to unravel the aerosol nitrate behaviors and governing
factors in different representative urban areas of China.
Atmosphere 2017, 8, 62 3 of 12
Figure 1. Sampling periods, locations, and aerosol mass spectrometers (AMS) versions of the datasets
used in this work.
Due to the limitation of transmission efficiency of the AMS lens, AMS is able to measure particles
mainly in the submicron meter range, and species that can vaporize at ~600 °C, thus the
AMS-measured particles are referred to as non-refractory PM
1
(NR-PM
1
) [21]. During the
aforementioned field measurements, although different types of AMS were deployed, all of them
(SP-AMS, ACSM, and HR-ToF-AMS) used ammonium nitrate to conduct the mass calibration, since
it is easy to evaporate and has a very simple fragmentation pattern (mainly m/z 30—NO
+
, and
m/z 46—NO
2+
from nitrate). The AMS is thus powerful in quantifying nitrate unless there are large
amounts of organic or metal nitrates, since they can affect NO
+
and NO
2+
signals significantly [33].
Quantification of other aerosol components, including sulfate, chloride, ammonium, and total
organics, were all based on their relative ionization efficiencies to nitrate. Only SP-AMS can
simultaneously determine the refractory black carbon (rBC) mass, since it is equipped with an
additional laser vaporizer, and such rBC data were included here for Nanjing datasets. For Lanzhou,
a multi-angle absorption photometer (MAAP) and a single-particle intracavity laser-induced
incandescence photometer (SP2) were used to measure rBC during summer and winter, respectively
[31,32]. No rBC mass concentrations were available for the Beijing datasets. The total organics can be
further segregated into a few factors/sources (primary sources including traffic, cooking, biomass
burning, coal combustion, etc., and secondary sources) via positive matrix factorization (PMF) [34,35].
Other supporting data included the meteorological parameters (temperature, relative humidity—RH,
wind speed and direction, pressure, precipitation, and solar radiation) and concentrations of the
gaseous pollutants (CO, SO
2
, O
3
, NO
2
).
3. Results and Discussion
3.1. Average Mass Contributions, Diurnal Patterns, and Driving Factors
Figure 2 shows the campaign-averaged mass contributions of nitrate to the total PM
1
(left panel)
and the corresponding diurnal cycles of nitrate concentrations (right panel) during summer and
winter, in Nanjing, Beijing, and Lanzhou, respectively. It is clear that for these three cities, the PM
1
pollutions were all heavier during winter than during summer. This result is consistent with the air
pollution status in most cities of China [3]. Nitrate concentrations were also elevated in winter—
11.1 μg/m
3
(winter) versus 3.4 μg/m
3
(summer) in Nanjing and 7.2 μg/m
3
(winter) versus 3.9 μg/m
3
(summer) in Lanzhou, but in Beijing, nitrate mass loading was slightly higher during summer than
it during winter (12.4 versus 10.9 μg/m
3
, respectively). The mass fractions of nitrate also increased
during winter compared with those during summer in Nanjing and Lanzhou, while the fractions
decreased during winter from those during summer in Beijing, mainly due to the increased
Figure 1.
Sampling periods, locations, and aerosol mass spectrometers (AMS) versions of the datasets
used in this work.
Due to the limitation of transmission efficiency of the AMS lens, AMS is able to measure particles
mainly in the submicron meter range, and species that can vaporize at ~600
◦
C, thus the AMS-measured
particles are referred to as non-refractory PM
1
(NR-PM
1
) [
21
]. During the aforementioned field
measurements, although different types of AMS were deployed, all of them (SP-AMS, ACSM, and
HR-ToF-AMS) used ammonium nitrate to conduct the mass calibration, since it is easy to evaporate
and has a very simple fragmentation pattern (mainly m/z 30—NO
+
, and m/z 46—NO
2+
from nitrate).
The AMS is thus powerful in quantifying nitrate unless there are large amounts of organic or metal
nitrates, since they can affect NO
+
and NO
2+
signals significantly [
33
]. Quantification of other aerosol
components, including sulfate, chloride, ammonium, and total organics, were all based on their relative
ionization efficiencies to nitrate. Only SP-AMS can simultaneously determine the refractory black
carbon (rBC) mass, since it is equipped with an additional laser vaporizer, and such rBC data were
included here for Nanjing datasets. For Lanzhou, a multi-angle absorption photometer (MAAP) and a
single-particle intracavity laser-induced incandescence photometer (SP2) were used to measure rBC
during summer and winter, respectively [
31
,
32
]. No rBC mass concentrations were available for the
Beijing datasets. The total organics can be further segregated into a few factors/sources (primary
sources including traffic, cooking, biomass burning, coal combustion, etc., and secondary sources)
via positive matrix factorization (PMF) [
34
,
35
]. Other supporting data included the meteorological
parameters (temperature, relative humidity—RH, wind speed and direction, pressure, precipitation,
and solar radiation) and concentrations of the gaseous pollutants (CO, SO2, O3, NO2).
3. Results and Discussion
3.1. Average Mass Contributions, Diurnal Patterns, and Driving Factors
Figure 2shows the campaign-averaged mass contributions of nitrate to the total PM
1
(left panel)
and the corresponding diurnal cycles of nitrate concentrations (right panel) during summer and winter,
in Nanjing, Beijing, and Lanzhou, respectively. It is clear that for these three cities, the PM
1
pollutions
were all heavier during winter than during summer. This result is consistent with the air pollution
status in most cities of China [
3
]. Nitrate concentrations were also elevated in winter—11.1
µ
g/m
3
(winter) versus 3.4
µ
g/m
3
(summer) in Nanjing and 7.2
µ
g/m
3
(winter) versus 3.9
µ
g/m
3
(summer) in
Lanzhou, but in Beijing, nitrate mass loading was slightly higher during summer than it during winter
Atmosphere 2017,8, 62 4 of 12
(12.4 versus 10.9
µ
g/m
3
, respectively). The mass fractions of nitrate also increased during winter
compared with those during summer in Nanjing and Lanzhou, while the fractions decreased during
winter from those during summer in Beijing, mainly due to the increased contributions from organic
constituents. Overall, we find that in urban areas of China, aerosol nitrate can occupy ~1/8 to ~1/4 of
the total PM
1
mass based on the measurements shown here. This relative contribution of nitrate is
generally within the range observed in other urban/suburban areas, but is typically higher than those
measured in rural or remote areas [
36
]. This finding clearly shows that urban traffic emissions are the
key factor to nitrate formation and thus represent an important contributor to the aerosol pollution
in China.
Atmosphere 2017, 8, 62 4 of 12
contributions from organic constituents. Overall, we find that in urban areas of China, aerosol nitrate
can occupy ~1/8 to ~1/4 of the total PM1 mass based on the measurements shown here. This relative
contribution of nitrate is generally within the range observed in other urban/suburban areas, but is
typically higher than those measured in rural or remote areas [36]. This finding clearly shows that
urban traffic emissions are the key factor to nitrate formation and thus represent an important
contributor to the aerosol pollution in China.
Figure 2. Average fine aerosol particulate matter (PM1) mass loadings, chemical compositions and
diurnal variations of the mass concentrations of nitrate: (a,b) Nanjing summer, (c,d) Nanjing winter,
(e,f) Beijing summer, (g,h) Beijing winter, (i,j) Lanzhou summer, and (k,l) Lanzhou winter (Beijing
datasets did not include refractory black carbon (rBC); the whiskers above and below the boxes are the
90th and 10th percentiles, the upper and lower boundaries of the boxes are the 75th and 25th percentiles,
and the lines in the boxes indicate the median values and the dots indicate the mean values; note Figure
2e–l are reproduced and modified from previous studies [28,29,31,32]).
31.4%
17.3%
29.2%
16.6%
0.9%
4.6%
Total PM
1
: 19.9 µg/m
3
(a)
Org.
NO
3
-
SO
4
2-
Cl
-
rBC
NH
4
+
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
(b) Nanjing Summer
26.0%
24.1%
Total PM
1
: 46.5 µg/m
3
(c)
23.3%
16.9%
4.0%
5.8%
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
(d) Nanjing Winter
40.1%
24.8%
Total PM
1
: 49.9 µg/m
3
(e)
18.1%
16.1%
1.0%
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
(f) Beijing Summer
51.6%
16.3%
Total PM
1
: 66.8 µg/m
3
(g)
14.0%
12.9%
5.2%
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
(h) Beijing Winter
47.0%
10.0%
16.0%
11.0%
4.0%
12.0%
Total PM
1
: 24.5 µg/m
3
(i)
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
(j) Lanzhou Summer
16.5%
12.5%
10.3%
3.0%
6.4%
Total PM
1
: 57.3 µg/m
3
(k)
51.3%
25
20
15
10
5
0
Mass Conc. (µg m
-3
)
12345678910 1112131415 161718192021 222324
Hour of Day
(l) Lanzhou Winter
Figure 2.
Average fine aerosol particulate matter (PM
1
) mass loadings, chemical compositions and
diurnal variations of the mass concentrations of nitrate: (
a
,
b
) Nanjing summer, (
c
,
d
) Nanjing winter,
(
e
,
f
) Beijing summer, (
g
,
h
) Beijing winter, (
i
,
j
) Lanzhou summer, and (
k
,
l
) Lanzhou winter (Beijing
datasets did not include refractory black carbon (rBC); the whiskers above and below the boxes are
the 90th and 10th percentiles, the upper and lower boundaries of the boxes are the 75th and 25th
percentiles, and the lines in the boxes indicate the median values and the dots indicate the mean values;
note Figure 2e–l are reproduced and modified from previous studies [28,29,31,32]).
Atmosphere 2017,8, 62 5 of 12
The AMS measurement has a very fine time resolution. For the field measurements included
in this work, the sampling resolutions were 5 to 30 min, allowing us to derive the average diurnal
cycles of nitrate which cannot be obtained from filter-based studies. As shown in Figure 2, the diurnal
patterns of nitrate presented different behaviors among different cities and seasons. In Nanjing, nitrate
varied relatively less during winter than it did during summer, overall, both seasons presented a
similar trend—peaking in early morning and reaching minimum in the afternoon. In fact, this pattern
was also observed by prior studies conducted during spring [
24
], autumn [
37
], and winter in urban
Nanjing [
38
], suggesting that this behavior is similar throughout the year in Nanjing. For Beijing,
summer nitrate concentrations responded similarly to the trends observed in the summer in Nanjing,
while during winter, Beijing’s nitrate concentrations displayed an opposite trend—peaking in the
late afternoon/early evening (~7–8 p.m.) and dropping to minimum values in the early morning.
This wintertime diurnal pattern was confirmed by another study conducted later in the same site [
39
].
Sun et al. [40]
also conducted one-year continuous aerosol measurement, and found that the nitrate
diurnal patterns differed greatly among different seasons, but that the afternoon increase only appeared
during winter, indicating a specific wintertime nitrate formation pathway in Beijing. In case of Lanzhou,
during summer, nitrate concentrations displayed a sharp peak in the morning (~10–11 a.m.), then
gradually decreased to relatively low levels at other times; during winter, the diurnal cycle was
somewhat similar to that observed in Beijing during winter, but changed more dramatically and
peaked earlier (about 3 p.m.). This afternoon nitrate peak in Lanzhou was also observed by another
AMS campaign conducted during 27 October to 3 December 2014 [41].
In order to elucidate the dominant factors that drive the different nitrate diurnal cycles presented
above, we further calculated the diurnal variations of temperature, RH, solar radiation, and appended
them along with that of nitrate for the six datasets in Figure 3. We also computed the theoretical
dissociation constants of NH
4
NO
3
(K
p
) by assuming the system was at thermodynamic equilibrium.
For Nanjing summer, nitrate varied closely with K
p
(r= 0.92) and RH (r= 0.92), but oppositely with
temperature (r=
−
0.92). Nanjing wintertime nitrate behaved similarly as it did during summer, with
correlation coefficients of r= 0.89 with K
p
,r= 0.93 with RH, and r=
−
0.90 with temperature. As is well
known, low temperature and high RH favor the thermodynamic gas/particle partitioning of NH
4
NO
3
,
thus these results verify that homogenous reaction between ammonia and nitric acid was the dominant
mechanism controlling nitrate formation during both summer and winter in Nanjing. Note that the
nitrate variations did not respond to the changes of solar radiation, indicating that photochemical
activities did not directly influence the nitrate formation in Nanjing. We propose this nitrate formation
mechanism as “Type I—thermodynamics driven”.
In the case of Beijing, nitrate in summer generally also matched well with the thermodynamic
parameters (r= 0.84 with K
p
,r= 0.88 with RH, and r=
−
0.89 with temperature). On the contrary,
wintertime nitrate correlated oppositely with these parameters (r=
−
0.67 with K
p
,r=
−
0.34 with
RH, and r= 0.58 with temperature), but positively responded to the increase of solar radiation during
daytime, reflecting the dominant contribution from photochemical production over thermodynamic
gas/particle partitioning. This mechanism is also proposed by Sun et al. [
29
], and here we define it as
“Type II—photochemistry driven”.
In the case of Lanzhou, the nitrate diurnal pattern during summer had no correlations with K
p
(
r=−0.02
), temperature (r=
−
0.09), and RH (r= 0.15), indicating an insignificant role of gas/particle
partitioning (Type I); although no solar radiation data was available, the sharp decrease of nitrate
during afternoon demonstrated that it was not associated with photochemical production too (Type II).
Xu et al. [
31
] interpreted that the dynamics of planetary boundary layer (PBL) and down-mixing of
nitrate produced from nocturnal chemistry in the morning, likely caused such nitrate behavior during
summer in Lanzhou. This behavior is also likely related with the specific topography of Lanzhou,
as a prior study reported that pollution of secondary species could be enhanced due to mountain
trapping [
42
]. This mechanism is defined as “Type III—PBL dynamics driven”. Nitrate in Lanzhou
during winter had negative correlations with the thermodynamic parameters (r=
−
0.80 with K
p
,
Atmosphere 2017,8, 62 6 of 12
r=−0.93
with RH, and r= 0.91 with temperature); it also had an afternoon peak, indicating it was
mainly driven by photochemical production (Type II)—a mechanism proposed by Xu et al. [
32
] as well.
Here we propose three types of mechanisms that can govern urban aerosol nitrate formation.
However, it should be noted that there are likely other mechanisms. For example, Yang et al. [
43
]
showed that transportation of haze from other regions could lead to the high nitrate concentrations
observed in Beijing.
Atmosphere 2017, 8, 62 6 of 12
Here we propose three types of mechanisms that can govern urban aerosol nitrate formation.
However, it should be noted that there are likely other mechanisms. For example, Yang et al. [43]
showed that transportation of haze from other regions could lead to the high nitrate concentrations
observed in Beijing.
Figure 3. Diurnal patterns of temperature (top panel, left y axis), relative humidity (RH) (top panel, right
y axis), nitrate concentration (bottom panel, left y axis), solar radiation and equilibrium constant (Kp) of
ammonium nitrate (bottom panel, right y axis) (K=K
298expa
−1+b1+ln
−
,
for reaction NH4NO3(p) ↔ NH3(g) + HNO3(g). Kp(298) is the equilibrium constant at 298 K (3.36 × 1016
atm−2), a = 75.11 and b = −13.5 [16]): (a) Nanjing summer, (b) Nanjing winter,
(c) Beijing summer, (d) Beijing winter, (e) Lanzhou summer, and (f) Lanzhou winter. (Type I—
thermodynamics driven; Type II—photochemistry driven; Type III—PBL dynamics driven. Please
refer to the main text for more details).
3.2. Influences of Other Aerosol Components
Due to influences of other aerosol components including primary species directly emitted from
various sources, and the secondary components produced by different pathways, the mass
contributions of nitrate to PM1 may vary greatly at different pollution levels. In Figure 4, we plotted
the mass percentages of nitrate against the total PM1 concentrations (divided into a number of bins).
-2
0
2
Temp. (
o
C)
24211815129630
60
50
40
30
RH (%)
300
200
100
0
Solar Radia. (W/m
2
)
16
14
12
10
8
6
NO
3
Mass Conc. (µg/m
3
)
8
6
4
2
0
K
p
(x 10
19
atm
-2
)
(d) Beijing Winter
30
28
26
24
Temp. (
o
C)
24211815129630
80
70
60
50
RH (%)
500
400
300
200
100
0
Solar Radia. (W/m
2
)
18
16
14
12
10
8
6
NO
3
Mass Conc. (µg/m
3
)
5
4
3
2
1
0
K
p
(x 10
16
atm
-2
)
(c) Beijing Summer
-6
-3
0
3
Temp. (
o
C)
24211815129630
Hour of day
50
40
30
RH (%)
18
15
12
9
6
3
NO
3
Mass Conc. (µg/m
3
)
40
30
20
10
0
K
p
(x 10
19
atm
-2
)
(f) Lanzhou Winter
28
26
24
22
20
Temp. (
o
C)
24211815129630
Hour of day
80
70
60
50
40
RH (%)
8
6
4
2
0
NO
3
Mass Conc. (µg/m
3
)
6
4
2
0
K
p
(x 10
16
atm
-2
)
(e) Lanzhou Summer
26
24
22
Temp. (
o
C)
24211815129630
100
90
80
70
RH (%)
400
300
200
100
0
Solar Radia. (W/m
2
)
6
5
4
3
2
1
NO
3
Mass Conc. (µg/m
3
)
7
6
5
4
3
2
K
p
(x 10
16
atm
-2
)
(a) Nanjing Summer
9
8
7
6
5
4
Temp. (
o
C)
24211815129630
75
70
65
60
55
50
RH (%)
300
200
100
0
Solar Radia. (W/m
2
)
16
14
12
10
8
6
NO
3
Mass Conc. (µg/m
3
)
10
8
6
4
2
K
p
(x 10
18
atm
-2
)
(b) Nanjing Winter
Type I Type I
Type I Type II
Type III Type II
Figure 3.
Diurnal patterns of temperature (top panel, left yaxis), relative humidity (RH) (top panel, right
yaxis), nitrate concentration (bottom panel, left yaxis), solar radiation and equilibrium constant (
Kp
) of
ammonium nitrate (bottom panel, right yaxis) (
Kp=Kp(298)expna298
T−1+bh1+ln298
T−298
Tio
,
for reaction NH
4
NO
3
(p)
↔
NH
3
(g) + HNO
3
(g). K
p
(298) is the equilibrium constant at 298 K
(
3.36 ×1016 atm−2)
, a = 75.11 and b =
−
13.5 [
16
]): (
a
) Nanjing summer, (
b
) Nanjing winter, (
c
) Beijing
summer, (
d
) Beijing winter, (
e
) Lanzhou summer, and (
f
) Lanzhou winter. (Type I—thermodynamics
driven; Type II—photochemistry driven; Type III—PBL dynamics driven. Please refer to the main text
for more details).
3.2. Influences of Other Aerosol Components
Due to influences of other aerosol components including primary species directly emitted
from various sources, and the secondary components produced by different pathways, the mass
contributions of nitrate to PM
1
may vary greatly at different pollution levels. In Figure 4, we plotted
the mass percentages of nitrate against the total PM1concentrations (divided into a number of bins).
Atmosphere 2017,8, 62 7 of 12
Atmosphere 2017, 8, 62 7 of 12
Figure 4. Variations of the mass percentages of nitrate as a function of the total non-refractory PM1
(NR-PM1)concentrations (left y axis, note the NR-PM1 did not include rBC for all datasets in this
figure), and the percentage of data points in each bin (right y axis): (a) Nanjing summer, (b) Nanjing
winter, (c) Beijing summer, (d) Beijing winter, (e) Lanzhou summer, and (f) Lanzhou winter (the box
plot symbols are the same as those described in Figure 1; note Figure 4c–f are reproduced and
modified from previous studies [28,29,31,32]).
For Nanjing summer, besides a small pool of data (1.3% of total) with PM1 concentrations
exceeding 50 μg/m3, the mass fractions of nitrate increased with the increase of PM1 mass loadings.
Generally, this finding is consistent with the measurement results of Zhang et al. [37] during summer
in Nanjing. The data of Lanzhou during summer showed a more obvious increasing trend, as shown
in Figure 4e. For Beijing, for the majority of data (98.7% of total), nitrate concentrations also displayed
an increasing trend. Nitrate concentrations decreased slightly (from 30% to 27%) only for some
extremely polluted periods (1.3% of total). These results suggest that although nitrate concentrations
were overall low, as it is easy to evaporate due to high temperatures in summer, the contribution of
nitrate to PM1 is generally higher during polluted periods than during periods of clean(er) air. This
finding highlights the importance of NOx emission control (such as from traffic or industry) to reduce
nitrate concentrations during heavy haze events in summer.
On the other hand, wintertime nitrate behaviors were different. First, on average, the
fluctuations were smaller during winter than during summer—Nanjing: summer (13%–28%) versus
winter (18%–28%), Beijing: summer (11%–30%) versus winter (11%–18%), Lanzhou: summer
(6%–27%) versus winter (12%–21%). Secondly, the relative contributions of nitrate presented a
general trend which increased first then decreased with the increase of PM1 concentrations. Different
from summer where the nitrate mass fractions decreased only for a very small fraction of data, the
decrease in nitrate fractions during winter occurred for a much larger portion of data (~11% in
40
35
30
25
20
15
10
5
0
Mass Perc. (%)
(a) Nanjing Summer
Total data points: 2889
30
25
20
15
10
5
0
% of points per bin
(b) Nanjing Winter
Total data points: 3199
40
35
30
25
20
15
10
5
0
Mass Perc. (%)
(c) Beijing Summer
Total data points: 5098
30
25
20
15
10
5
0
% of points per bin
(d) Beijing Winter
Total data points: 4980
40
35
30
25
20
15
10
5
0
Mass Perc. (%)
200150100500
NR-PM
1
mass conc. (µg/m
3
)
(e) Lanzhou Summer
Total data points: 3484
200150100500
NR-PM
1
mass conc. (µg/m
3
)
30
25
20
15
10
5
0
% of points per bin
(f) Lanzhou Winter
Total data points: 3528
Figure 4.
Variations of the mass percentages of nitrate as a function of the total non-refractory PM
1
(NR-PM
1
)concentrations (left yaxis, note the NR-PM
1
did not include rBC for all datasets in this figure),
and the percentage of data points in each bin (right yaxis): (
a
) Nanjing summer, (
b
) Nanjing winter,
(
c
) Beijing summer, (
d
) Beijing winter, (
e
) Lanzhou summer, and (
f
) Lanzhou winter (the box plot
symbols are the same as those described in Figure 1; note Figure 4c–f are reproduced and modified
from previous studies [28,29,31,32]).
For Nanjing summer, besides a small pool of data (1.3% of total) with PM
1
concentrations
exceeding 50
µ
g/m
3
, the mass fractions of nitrate increased with the increase of PM
1
mass loadings.
Generally, this finding is consistent with the measurement results of Zhang et al. [
37
] during summer
in Nanjing. The data of Lanzhou during summer showed a more obvious increasing trend, as shown in
Figure 4e. For Beijing, for the majority of data (98.7% of total), nitrate concentrations also displayed an
increasing trend. Nitrate concentrations decreased slightly (from 30% to 27%) only for some extremely
polluted periods (1.3% of total). These results suggest that although nitrate concentrations were overall
low, as it is easy to evaporate due to high temperatures in summer, the contribution of nitrate to
PM
1
is generally higher during polluted periods than during periods of clean(er) air. This finding
highlights the importance of NO
x
emission control (such as from traffic or industry) to reduce nitrate
concentrations during heavy haze events in summer.
On the other hand, wintertime nitrate behaviors were different. First, on average, the fluctuations
were smaller during winter than during summer—Nanjing: summer (13%–28%) versus winter
(18%–28%), Beijing: summer (11%–30%) versus winter (11%–18%), Lanzhou: summer (6%–27%)
versus winter (12%–21%). Secondly, the relative contributions of nitrate presented a general trend
which increased first then decreased with the increase of PM
1
concentrations. Different from summer
where the nitrate mass fractions decreased only for a very small fraction of data, the decrease in
nitrate fractions during winter occurred for a much larger portion of data (~11% in Nanjing, ~22% in
Atmosphere 2017,8, 62 8 of 12
Beijing, and ~4% in Lanzhou). This wintertime “increase-decrease” nitrate behavior was confirmed
by other studies, such as a study by Yao et al. during winter in Lanzhou [
41
]. For Nanjing in winter,
further analyses demonstrated that the elevation of primary organic aerosol (OA) contributions was an
important reason causing the decrease in nitrate fractions at high PM1loadings. Sun et al. [29] found
that for winter in Beijing, the decrease in nitrate fractions were mainly due to increases in contributions
from sulfate and OAs (especially primary OAs, such as coal combustion OAs), and the aqueous-phase
oxidation pathway likely contributed significantly to sulfate [
30
]. In the case of winter in Lanzhou,
Xu et al. [
32
] showed that the increased contribution from primary OAs (including biomass burning
OAs, coal combustion OAs, traffic- and cooking-related OAs) led to the decrease in nitrate fractions
at high PM
1
concentrations. These results point out the need and importance of emission control of
the primary sources in urban areas during hazy days. The results also suggest that emission control
of secondary aerosol gaseous precursors in some areas is likely also important, as the production of
secondary species (including both inorganic and organic species [
44
–
46
]) might be enhanced during
hazy days. This is also proposed by Huang et al. [
9
] based on the PM
2.5
results for a few Chinese
megacities during January 2013.
3.3. Competition with Ammonium Sulfate/Bisulfate
For submicron aerosols, nitrate is overwhelmingly present in the form of NH
4
NO
3
rather than
other metal nitrates. NH
4
NO
3
is formed from the neutralization of HNO
3
with NH
3
. However, under
real atmospheric conditions, HNO
3
has to compete with H
2
SO
4
, as NH
3
tends to first react with
H
2
SO
4
thermodynamically. Previous studies show that NH
4
NO
3
can start to form when the molar
ratio of NH
4+
/SO
42−
is larger than 1.5 (ammonium-rich). In fact, we found that the molar ratios of
NH
4+
/SO
42
for the AMS data investigated here were in most cases larger than 2. Note that due to
possible influences from reaction of NH
3
with HCl (although it was likely minor, as chloride fractions
were much lower than those of nitrate and sulfate), the NH
4+
molar concentrations calculated here
excluded the portions that were neutralized by HCl. The results indicate the abundance of ammonia,
which is favorable for nitrate formation. This is different from the filter-based PM
2.5
results reported
by Ye et al. [
20
] in Shanghai, from where a large portion of data was in the ammonium-poor regions
(NH4+/SO42−< 1.5), especially during summer and autumn.
In order to further investigate the relationship between NH
4
NO
3
, NH
4
HSO
4
, and (NH
4
)
2
SO
4
,
we define an excess molar NH
4+
relative to (NH
4
)
2
SO
4
:
Excess NH4+= (NH4+/SO42−−2) ×SO42−
.
In Figure 5, we show the scatter plots of the excess molar NH
4+
concentrations versus molar
concentrations of nitrate for all datasets. In this figure, a slope larger than 1 means that NH
4
NO
3
forms when sulfuric acid is fully neutralized by ammonia (namely, ammonium sulfate). A slope with
a positive value but less than 1 means that partial NH
4+
is associated with bisulfate in addition to
nitrate. As (NH
4
)
2
SO
4
plus any fraction of NH
4
HSO
4
can result in a compound with the NH
4+
/SO
42
stoichiometric ratio larger than 1.5—that of (NH
4
)
3
H(SO
4
)
2
, in this case, it therefore means that
NH
4
NO
3
can start to form after the formation of (NH
4
)
3
H(SO
4
)
2
. The correlations shown in Figure 5
are generally strong (r
2
of 0.84–0.96), but it should be noted that the fitted slopes only represent the
statistical average values. For each individual dataset, there are data points with molar ratios larger or
less than 1 (especially for Beijing winter), and the deviations of fitted lines from 1:1 lines are sometimes
subtle; for instance the slopes of Beijing Winter (1.02) and Lanzhou summer (0.98) datasets are very
close to 1. Nevertheless, from a statistical standpoint, on average, during summer in Nanjing, and
during both summer and winter in Beijing, nitrate seemed significantly likely to be generated following
the formation of (NH
4
)
2
SO
4
, while it was able to form after the formation of (NH
4
)
3
H(SO
4
)
2
during
winter in Nanjing, and during both summer and winter in Lanzhou.
Atmosphere 2017,8, 62 9 of 12
Atmosphere 2017, 8, 62 9 of 12
Figure 5. Scatter plots of the molar concentrations of excess NH
4+
relative to (NH
4
)
2
SO
4
versus NO
3−
(the NH
4+
molar concentrations were the measured NH
4+
molar concentrations minus the amounts
used to neutralize HCl): (a) Nanjing summer, (b) Nanjing winter, (c) Beijing summer, (d) Beijing
winter, (e) Lanzhou summer, and (f) Lanzhou winter (data pointes were colored by time).
4. Conclusions
This work investigated the characteristics and formation mechanisms of fine aerosol nitrate in
typical urban areas of China using the Aerodyne AMS data collected in Nanjing, Beijing, and
Lanzhou, during both summer and winter. Results showed that nitrate could occupy up to a quarter
of urban PM
1
mass, with concentrations that were typically higher than those found in rural/remote
regions. During summer, the relative mass contributions of nitrate generally increased with the
increase of PM
1
mass loadings, highlighting the importance of NO
x
emission control; during winter,
nitrate contributions increased first then decreased with the increase of PM
1
concentrations, with the
decrease likely caused by the increases of both primary and secondary species, suggesting that
control of both primary particles and secondary aerosol precursors is essential for effectively
reducing the heavy haze pollution. Furthermore, by investigating the relationships between the
Figure 5.
Scatter plots of the molar concentrations of excess NH
4+
relative to (NH
4
)
2
SO
4
versus NO
3−
(the NH
4+
molar concentrations were the measured NH
4+
molar concentrations minus the amounts
used to neutralize HCl): (
a
) Nanjing summer, (
b
) Nanjing winter, (
c
) Beijing summer, (
d
) Beijing winter,
(e) Lanzhou summer, and (f) Lanzhou winter (data pointes were colored by time).
4. Conclusions
This work investigated the characteristics and formation mechanisms of fine aerosol nitrate in
typical urban areas of China using the Aerodyne AMS data collected in Nanjing, Beijing, and Lanzhou,
during both summer and winter. Results showed that nitrate could occupy up to a quarter of urban
PM
1
mass, with concentrations that were typically higher than those found in rural/remote regions.
During summer, the relative mass contributions of nitrate generally increased with the increase of
PM
1
mass loadings, highlighting the importance of NO
x
emission control; during winter, nitrate
contributions increased first then decreased with the increase of PM
1
concentrations, with the decrease
likely caused by the increases of both primary and secondary species, suggesting that control of both
Atmosphere 2017,8, 62 10 of 12
primary particles and secondary aerosol precursors is essential for effectively reducing the heavy haze
pollution. Furthermore, by investigating the relationships between the diurnal patterns of nitrate and
those of meteorological parameters, etc., we proposed three types of mechanisms that can govern
the nitrate variations in urban China: “Type I—thermodynamics driven”, “Type II—photochemistry
driven”, and “Type III—PBL dynamics driven”. Competition between the formation of ammonium
nitrate and ammonium bisulfate/sulfate were discussed, and we found that NH
4
NO
3
could start to
form before sulfuric acid was fully neutralized and after (NH4)3H(SO4)2was formed.
Overall, this work provides useful insights into nitrate formation, and thus is valuable for
reducing aerosol nitrate pollution. The three types of mechanisms we proposed are valuable to aid in
the interpretation of nitrate behaviors documented in other studies. Nevertheless, this study indicates
that nitrate formation can be complex, and does not exclude other mechanisms. Different nitrate
pollution events occurring in the same location might be dominated by different mechanisms as
well [
43
]. Future work is thus still necessary to carefully elucidate the atmospheric behaviors of nitrate
under different atmospheric environments.
Acknowledgments:
This work was financially supported by the Natural Science Foundation of China
(21407079, 21577065, and 91544220), Jiangsu Natural Science Foundation (BK20150042), Jiangsu Provincial
Specially-Appointed Professors Foundation, Jiangsu Innovation and Entrepreneurship Program, and the Startup
Foundation for Introducing Talent of NUIST (2014r064).
Author Contributions:
Xinlei Ge conceived the idea; Xinlei Ge and Yanan He wrote the manuscript; Xinlei Ge,
Yanan He, and Junfeng Wang performed Nanjing field measurements and data analyses; Yele Sun provided Beijing
data, and Jianzhong Xu provided Lanzhou data; Yafei Shen and Mindong Chen provided valuable comments and
suggestions for the development of the manuscript.
Conflicts of Interest: The authors declare no conflict of interest.
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