Acid deposition in Asia: Emissions, deposition, and ecosystem effects
, Qian Yu
, Qiang Zhang
, Zifa Wang
, Yuepeng Pan
, Jie Tang
, Jan Mulder
State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment, Tsinghua University, Beijing, 100084, China
Center for Earth System Science, Tsinghua University, Beijing, 100084, China
State Key Laboratory of Atmospheric Boundary Layer Physics and Atmospheric Chemistry, Institute of Atmospheric Physics, Chinese Academy of Sciences,
Beijing, 100029, China
Norwegian Institute for Water Research, 0349, Oslo, Norway
Center for Atmosphere Watch and Services, Chinese Academy of Meteorological Sciences, Beijing, 100081, China
Department of Environmental Sciences, Norwegian University of Life Sciences, 1432, Ås, Norway
Here review more recent studies on acid deposition in Asia, especially in Eastern Asia.
Surface waters are generally not sensitive to acid deposition in comparison with soils.
Soil acidiﬁcation is not very serious because of base cation deposition, N denitriﬁcation, and SO
Received 8 March 2016
Received in revised form
5 July 2016
Accepted 7 July 2016
Available online 9 July 2016
We review and synthesize the current state of knowledge regarding acid deposition and its environ-
mental effects across Asia. The extent and magnitude of acid deposition in Asia became apparent only
about one decade after this issue was well described in Europe and North America. In addition to the
temperate zone, much of eastern and southern Asia is situated in the tropics and subtropics, climate
zones hitherto little studied with respect to the effects of high loads of acid deposition. Surface waters
across Asia are generally not sensitive to the effects of acid deposition, whereas soils in some regions are
sensitive to acidiﬁcation due to low mineral weathering. However, soil acidiﬁcation was largely
neutralized by such processes as base cation deposition, nitrate (NO
) denitriﬁcation, and sulfate (SO
adsorption. Accompanying the decrease in S deposition in recent years, N deposition is of increasing
concern in Asia. The acidifying effect of N deposition may be more important than S deposition in well
drained tropical/subtropical soils due to high SO
adsorption. The risk of regional soil acidiﬁcation is a
major threat in Eastern Asia, indicated by critical load exceedance in large areas.
©2016 Elsevier Ltd. All rights reserved.
Acid deposition became an issue of major concern in Asia in the
early 1980s, nearly one decade after widespread acid deposition
was recognized in Europe and North America (e.g. Bhatti, 1992).
Before the establishment of national monitoring networks for acid
deposition in Asia, isolated surveys of acidity level and chemical
composition of rainwater in some Asian countries (such as China,
Japan, and India) indicated the occurrence of acid rain (Bhatti,
1992). Speciﬁc regions with decreasing pH trends in precipitation
included southern China (south of the Yangtze River) (Zhao and
Sun, 1986; Galloway et al., 1987), southern (especially along the
east coast) and northeastern India (Varma, 1989), and some areas in
Japan (Hara, 1997) and Korea (Chung et al., 1996).
Nationwide surveys of acid rain began in the 1980s. In China,
these efforts were sponsored by the National Environmental Pro-
tection Agency (NEPA) of China from 1982 (Wang and Wang, 1996;
Fujita et al., 2000). The Japanese Acid Deposition Survey (JADS) has
been conducted since September 1983 by the Japan Environment
Agency (Seto et al., 2004; Okuda et al., 2005). These nation-wide
monitoring networks provide the longest record of wet
E-mail address: email@example.com (L. Duan).
All the co-authors contributed equally to this work.
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Atmospheric Environment 146 (2016) 55e69
deposition in Asia. To create a common understanding of the state
of acidic deposition problems in East Asia, the Acid Deposition
Monitoring Network in East Asia (EANET) began regular deposition
monitoring activities in January 2001. Currently, this network
consists of 54 monitoring sites in thirteen countries, including
Cambodia, China, Indonesia, Japan, Lao P.D.R, Malaysia, Mongolia,
Myanmar, Philippines, Republic of Korea, Russia, Thailand, and
Vietnam (EANET, 2011). Monitoring of wet and dry deposition,
together with ecological impacts, has been conducted as part of the
activities of this network.
In addition to monitoring, modeling has been carried out to
analyze the spatial variations and source-receptor relations of acid
deposition. The early modeling studies were carried out not only
for the whole of Asia (Kotamarthi and Carmichael, 1990; Arndt and
Carmichael, 1995; Arndt et al., 1998), but also more speciﬁcally for
Eastern Asia (Huang et al., 1995; Kim and Seog, 2003; Park et al.,
2005). Recently, the ensemble-mean depositions of sulfur (S) and
nitrogen (N) over Eastern Asia was presented based on eight
regional chemical models used in a model inter-comparison study
for Asia (MICS-Asia; Wang et al., 2008).
Asia is now the global hotspot of S and N deposition (Vet et al.,
2014). Since the early 2000s, the global maximum of both S and N
deposition is found in East Asia, including regions like eastern
China and South Korea. Other areas of high deposition in Asia
include sections of Pakistan, India, Bangladesh, Myanmar, Thailand,
Laos, North Korea, and Japan (Vet et al., 2014). Both monitoring and
modeling indicate that S and N deposition increased in China from
the 1980se2000s, most likely due to increased sulfur dioxide (SO
and nitrogen oxides (NOx) emissions. Recently, some researchers
have suggested that S deposition in China started to decrease as
early as 2006 (Zhao et al., 2009, 2013).
Anthropogenic inputs of S and N into terrestrial ecosystems
impact soil and surface water, causing acidiﬁcation and eutrophi-
cation (Bouwman et al., 2002). Because long-term data on surface
water chemistry are limited in the Asian region, the re-
measurement of previously surveyed rivers and lakes or the
assessment of public data on water quality are among the few
options to assess the current situation regarding water acidiﬁcation
and N leaching (Duan et al., 2011). So far, very few areas have shown
water acidiﬁcation. This is even true for areas with acidic soils and
high rates of acid deposition (Komai et al., 2001; Chen et al., 2012).
In comparison, soil acidiﬁcation, as indicated by a signiﬁcant
decrease in soil pH and increase in aluminum (Al) mobilization, and
increased N leaching (Aber et al., 2003), has been commonly re-
ported in East Asia (Larssen et al., 2011; Asano and Uchida, 2005;
Fang et al., 2011). Most of the available Asian data on the impacts
of acid deposition originate from Japan and China. Across these
nations, the biogeochemical cycles of major solutes like S, N and
calcium (Ca) are shown to be different from those in Europe and
North America, probably due to the warm and humid climate,
different soil and vegetation types, and different deposition char-
acteristics in China (such as high Ca deposition) (Chen and Mulder,
2007; Larssen et al., 2011).
The critical load of acid deposition is deﬁned as ‘‘a quantitative
estimate of an exposure to one or more pollutants below which
signiﬁcant harmful effects on speciﬁed sensitive elements of the
environment do not occur according to present knowledge’’
(Nilsson and Grennfelt, 1988). The critical load concept was
developed in the 1980s to support effective acid rain policy. It has
been widely used in international negotiations to reduce of SO
NOx emissions in Europe (Hettelingh et al., 1995a) with the aim of
decreasing S and N deposition in excess of critical loads through
cost-optimal emission abatement. Critical loads of acid deposition
have been determined and mapped in several studies for a number
of regions such as southeastern Asia (Hettelingh et al., 1995b) and
northern Asia (Bashkin et al., 1995), and for some countries
including China (Duan et al., 2000a; Zhao et al., 2009), Japan
(Shindo et al., 1995) and South Korea (Park and Lee, 2001). These
studies, mainly focus on acidiﬁcation, and may support the use of
critical loads as a basis for transboundary pollution impact analysis
and co-emission reduction negotiation in Asia in the future.
In this paper, we review and synthesize the current state of
knowledge regarding acid deposition and its environmental effects
across Asia, in particular East Asia. The trends of emissions of
acidifying precursors such as SO
, NOx, and ammonia (NH
deposition of S and N in recent years are summarized, and the ef-
fects of acid deposition on soil and surface water are discussed. We
hope the review may be useful for future studies and policy-
2. Emission and deposition
Driven by a dramatic economic development, Asian anthropo-
genic emissions of SO
, NOx, and NH
show increasing proportions
in the global budgets since 1970s (e.g., Fig. 1). For all of the three
acidifying precursors, more than 35% of the global emissions were
contributed by Asia in 2005, mainly contributed by China and India
(EC-JRC/PBL, 2011; Smith et al., 2011).
Emission trends of SO
, NOx, and NH
among Asian regions (Fig. 2). China’sSO
continuously before 2006, and declined after 2006 due to wide
application of ﬂue-gas desulfurization (FGD) in power plant units
since 2005. Decreasing trends were found for SO
Japan and South Korea due to the much earlier implementation of
stringent emission control measures. The rapid increase in SO
emission from Southeast Asia has slowed since 1997 (Fig. 2).
A rapid increase in NOx emissions (counted as NO
observed for China between 1990 and 2011 (Fig. 2), due to the
increasing energy consumption in industry and transportation
sectors, and lack of adequate control measures. NOx emissions from
China began to decrease in 2012 because of the wide applications of
selective catalytic reduction (SCR) in coal-ﬁred power plants (Wang
et al., 2014). Decreasing emission trends for Japan and South Korea
in recent years were driven by the stringent emission standards
implemented for vehicles (Wang et al., 2014).
In contrast, NH
emission trends were relatively ﬂat among
Asian regions, with the exception of China and Southeast Asia
(Fig. 2). NH
emissions from China increased during 1980e1996,
and then decreased (Kang et al., 2016). The decrease in China’sNH
Fig. 1. Global SO
emissions by region (North America ¼USA þCanada; East
Asia ¼Japan þChina þSouth Korea) (Smith et al., 2011).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e6956
emissions was attributed to a decline in ammonium bicarbonate
applications as fertilizer. A sharp increase in NH
Southeast Asia was mainly driven by fertilizer applications
(Kurokawa et al., 2013).
Fig. 3 presents the spatial distribution of Asian SO
, NOx, and
emissions in 2010 (from MIX inventory; Li et al., 2015). Mainly
produced during fuel consumption, SO
and NOx emissions were
concentrated in populated regions with high energy consumption
like East China and India. NH
emissions were distributed widely
among regions with intensive agriculture activities, such as China,
India and Southeast Asia.
Observational studies on spatial distributions of S and N depo-
sition for the whole of Asia are very limited. The annual bulk S
deposition at ﬁve forest sites in southern and southwestern China
ranged from 24 to 160 kg ha
, which is in the same range as,
Fig. 2. Emissions of acidifying precursors in several countries/regions in Asia (Data source: MEIC for China, www.meicmodel.org; REAS for other regions, Kurokawa et al., 2013).
Fig. 3. Emission of acidifying precursors in Asia in 2010 (Unit: Gg/gird; Data source: MIX Asian emission inventory; Li et al., 2015). The map resolution is 0.25(latitude) by 0.25
L. Duan et al. / Atmospheric Environment 146 (2016) 55e69 57
or higher than, that seen in most of central Europe in 1980, when
acid deposition was at its peak (Larssen et al., 2006). Very high total
deposition of S and N also occurred in north China (Pan et al., 2012,
2013). For example, higher S deposition was observed at industrial
and urban sites (50e100 kg ha
), reﬂecting a higher contri-
bution of dry deposition (mostly of gaseous SO
for about 70%) than of wet deposition (Pan et al., 2013). In addition,
the total N deposition in northern China (30e100 kg ha
also signiﬁcantly larger than that in other Asian countries such as
Japan because of high rates of wet deposition and gaseous NH
deposition (Pan et al., 2012). The annual S deposition for six remote
EANET sites in Japan during 2000e2004 ranged from 1.9 to
7.5 k g h a
. For the same six sites, the N deposition ranged
from 1.5 to 6.4 kg ha
(about half as NO
), which was lower
than values for urban sites in Japan, with average values of
7.8 k g h a
(Seto et al., 2007).
Most of monitoring data showed a decreasing trend of wet
deposition of SO
and thus an increasing trend of precipitation pH
in Japan since early 1990s (Seto et al., 2004), in parallel with a
decreasing trend in the SO
emission from Japan. For example, the
acid deposition survey by the Japan Environment Agency showed
an increasing trend of pH and a decreasing trend for both S and N
deposition from the period 1986e1988 to the period 1989e1993
(Hara, 1997). The pH data collected at the 29 stations in the ﬁrst
period showed a range of annual mean pH from 4.4 to 5.5 (Kitamura
et al., 1991). The range changed to 4.5 to 5.8 during the second
period (Hara, 1997).
Although the monitoring of S and N deposition was very limited
in China, both in space and time, long-term records of precipitation
pH are available. The overall precipitation pH showed three stages
(Fig. 4). In the period 1992e1999, the precipitation pH showed an
increasing trend. During 2000e2006, however, a decreasing trend
of the precipitation pH was observed in North China, Central China,
East China, and South China. The precipitation over North China,
Central China, and South China became more acidiﬁed during
1992e2006, with more pronounced trends in North China and the
north of Central China. A slight increase in the precipitation pH was
found in Southwest China, an area characterized by the most severe
acid rain for about two decades since the early 1980s. Consequently,
the center of the most severe acid rain area, south of the Yangtze
River, moved eastwards (Tang et al., 2010). Accompanying the rapid
reduction of SO
emission after 2006 (Fig. 2), precipitation pH
began to increase (Fig. 4).
Based on the EANET dataset of precipitation chemistry, yearly
wet deposition of SO
was averaged for all sites in each
country (Fig. 5). The annual SO
deposition was signiﬁcantly larger
in China than in South Korea and Japan, whereas the NO
tion was similar in these countries. However, the pH of rainfall in
China was higher due to the high buffering of precipitation acidity
by emissions of basic particulate matter (PM), including soil dust
(Larssen and Carmichael, 2000), anthropogenic dust (Zhu et al.,
2004; Lei et al., 2011), and NH
(mainly from agricultural activ-
ities; Kang et al., 2016). For example, the Ca
rainwater in northern China is much higher than in southern China
or in the United States and other industrial nations (Wang and
Wang, 1996; Wang et al., 2012; Cao et al., 2013). It is estimated
that if the soil-derived bases (particularly CaCO
) were eliminated
from the atmosphere, the precipitation pH in these northern sites
would average approximately 3.5 (Galloway et al., 1987).
Long-range transport of acidic pollutants has been suggested to
occur from the Asian continent to Japan and Korea. This may be one
reason why the trend of S and N deposition in Korea and Japan
(Fig. 5) did not coincide well with the SO
and NOx emission (Fig. 2).
For example, high atmospheric non-sea salt SO
observed at the Japan Sea coast during southwesterly wind, and the
concentration of anthropogenic sulfate and nitrate aerosol at the
central North Paciﬁc site almost doubled from 1981 to the mid-
1990s, parallel to increased SO
emission from China (Prospero
et al., 2003; Hideaki et al., 2008). Anthropogenic SO
emission from China contributed to the increase in SO
concentration in the wet-only samples from 2000 to 2007 in
coastal areas of Korea (Park et al., 2015). Higher S and N deposition
occurred on the western coasts of Korea than on the eastern coasts
(Kang et al., 2004). During 2000e2007, when SO
China increased by 53%, there was a longitudinal gradient in urban
concentration in Japan, with values decreasing the further the
site was from the Asian continent. This result demonstrates that, in
spite of the relatively short tropospheric lifetime of SO
transport of increasing SO
from the Asian continent can partially
counteract the local reduction of SO
emission downwind, and even
override it in some southwestern areas of Japan (Lu et al., 2010b).
emission from China began to decrease after 2006,
ambient air SO
concentration and SO
concentrations in precip-
itation also started to decrease, while rain pH increased in the
whole of East Asia (Lu et al., 2010b).
Unlike S, atmospheric N deposition rates in eastern Asia have
dramatically increased during recent decades due to the emissions of
NOx and NH
from combustion processes and agricultural activities,
respectively, not only from China (Liu et al., 2011), but also from Japan
(Kannari et al., 2001). For example, average bulk N deposition of all
available monitoring sites throughout China increased from
13.2 kg ha
in 1980s to 21.1 kg ha
in 2000s (Liu et al.,
2013). It has also been reported that forests in Central Japan
revealed a high level of N deposition and NO
in stream water (signal
forNsaturation)(Baba et al., 1995; Babaand Okazaki,1998; Ohrui and
Mitchell, 1997, 1998; Matano et al., 2001; Ham et al., 2010). Wet
deposition of N variedbetween 5.7 and 16.7 kg ha
contribution of NH
) from rural sites to urban areas during
1999e2002 (Paramee et al., 2005). A literature review on N deposi-
tion from 69 forest ecosystems at 50 sites throughout China indicated
that the wetdeposition of N rangedfrom 2.6 to 48.2 kg ha
an average of 16.6 kg ha
(Fanget al., 2011). Ammoniumwas the
dominant form of N at most sites, accounting for, on average, 63% of
total inorganic N deposition (Fang et al., 2011).
A decreasing trend of wet deposition has been observed in
Southeast Asia for SO
but not for NO
(EANET, 2006, 2011), which
coincides with the trend of SO
and NOx emissions (Fig. 2). The wet
deposition of SO
also decreased by about 5% between
2000e2002 and 20 05e2007 in northeast India, one of the areas
with highest S deposition in India (11.0 kg ha
(Vet et al., 2014). However, the wet deposition of N increased by
more than 30% in this area, which measured among the highest wet
N deposition rates in the world with values of about 20 kg ha
(Vet et al., 2014).
A recent study using the Nested Air Quality Prediction Modeling
System (NAQPMS), that coupled the cloud-process and aqueous
chemistry module from the Community Multi-scale Air Quality
(CMAQ) modeling system, indicated that a very high wet deposition
of both S and N (in the range of 16e25 kg ha
21e32 kg ha
, respectively) occurred in northern, south-
western, and eastern China in 2007 (Ge et al., 2014). The total
deposition of S and N was even higher than 70 kg ha
50 kg ha
respectively in these areas (Fig. 6). Strong
neutralization of precipitation by soil aerosols over northeastern
Asia was also estimated, with the increase in annual mean pH by
0.8e2.5 in northern China and Korea, while less than 0.1 in
southern China and Japan (Wang et al., 2002). In comparison with
natural sources such as desert, which mainly contributes to high
deposition in northern and northwestern China, the anthro-
pogenic sources of Ca
-content particulate matter such as cement
L. Duan et al. / Atmospheric Environment 146 (2016) 55e6958
production and coal combustion result in high Ca
eastern and southern China (Fig. 7). The highest Ca
modeled is even comparable with the S deposition on an equivalent
basis, indicating signiﬁcant neutralization of precipitation acidity
by base cations.
Here only surface water and soil acidiﬁcation/eutrophication
were focused on, although acid deposition has other impacts on
agriculture, human health, and infrastructure.
3.1. Water acidiﬁcation
Surface water acidiﬁcation, causing severe ecological damage in
Scandinavia and to some extent also elsewhere in northern Europe
and North America, used to be the main concern of damage caused
by acid deposition. Although extremely high deposition rates for
occur in East Asia, especially in China and Japan, only very few
streams in forested catchments in central Japan and on islands have
suffered from acidiﬁcation (Nakano et al., 2001; Kurita and Ueda,
2006; Yamada et al., 2007; Matsubara et al., 2009; Nakahara
et al., 2010). A large buffering capacity of the soil and high
Fig. 4. Spatial and temporal distribution of precipitation pH in China: (a) Maps of average annual precipitation pH; (b) Trend of nationally average precipitation pH and acid rain
frequency (F) during 1992e2014 (Data from Acid Rain Monitoring Network; CMA, 2014).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e69 59
alkalinity of the inland waters inhibit acidiﬁcation of the inland
water ecosystem in Japan under current levels of acidic precipita-
tion (Suzuki, 2003). The survey of rain and head water chemistry
between 1991 and 1997 revealed that the watershed ecosystem did
not show any direct evidence of water acidiﬁcation in spite of de-
cades of elevated H
deposition (estimated at 0.43 keq ha
more since 1960s; Ikeda and Hmada, 2001). Many monitoring
studies conducted during the 1990s showed that stream waters in
Fig. 5. Monitoring sites of EANET (Acid Deposition Monitoring Network in East Asia) and trend of average wet deposition in China, Japan, and South Korea (Data source: http://
Fig. 6. S and N deposition in Asia modeled by the Nested Air Quality Prediction Modeling System (NAQPMS) (Ge et al., 2014).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e6960
Japan (e.g., Baba and Okazaki, 1998; Komai et al., 2001; Ohte et al.,
2001) and China (Duan et al., 2000b; Hao et al., 2001a; Ye et al.,
2002) were well buffered (with high ANC) and there were no
signs of a long-term trend of pH decline.
Although some aquatic ecosystems with low alkalinity, consid-
ered to be sensitive to acidiﬁcation, are present in China, especially
in the south and northeast, until recently detailed surveys of those
headwater streams were scarce. A recent survey of headwater
streams showed enhanced concentration of SO
than 0.5 meq/L and 0.1 meq/L, respectively) in stream waters in
southwestern China under very high atmospheric deposition of S
and N (Fig. 8). Much lower concentration of SO
in stream waters in Japan and South Korea (Fig. 8).
As can be seen from Fig. 8, there may be a big difference be-
tween China and Japan in the acidifying potential of N, which is
mainly due to the difference in N input (relatively small in Japan as
in the US, and much higher in China as in Northwest Europe; Figs. 5
and 6). In Japan even with N deposition below 10 kg ha
(1.0 g m
), a generally used threshold above which nitrate
leaching to streams is predicted (Dise et al., 1998), there was evi-
dence of elevated NO
concentration of stream water in a larch
forest (Nakahara et al., 2003). The enhanced NO
accelerate surface water acidiﬁcation in central Japan. One example
was the Lake Ijira catchment, which is one of the forested catch-
ments with wet N depositions (19 kg ha
) among the highest
levels observed in Japan (Nakahara et al., 2010). In contrast,
leaching to stream waters are not likely to occur
in China, unless N deposition exceeds 25 kg ha
(2.5 g m
, or even more in some region) (Fig. 8).
Typical for much of southern China, where soil and forest types
(Haplic Acrisol and subtropical evergreen coniferous forests,
respectively) are similar as in SW2, and N deposition (both NH
-N) is commonly high, is that forested catchments act as
large N sinks, shown by relatively low NO
concentration (about a
few mg/L) in stream water (Larssen et al., 2011). Most of the NO
lost from solution upon its transport through the groundwater
discharge zone (GDZ) before reaching the stream, due to denitri-
ﬁcation (Zhu et al., 2013a). This is conﬁrmed by studies of natural
abundance, where enrichment of
O in the GDZ were
attributed to the denitriﬁcation process (Yu et al., 2016). Denitriﬁ-
cation is an acid neutralizing process, and contributes to the higher
soil pH in the GDZ than on the well-drained hill slopes. Denitriﬁ-
cation results in not only the production of N
gas, but also emission
of nitrous oxide (N
O), a potent greenhouse gas. In wet years, N
emission is accounted for close to 10% of the total N sink at the
Tieshanping site in southwestern China (Zhu et al., 2013b). Thus,
the groundwater discharge zones of N-saturated forests in southern
China are hot-spots of denitriﬁcation, thus protecting stream water
from excessive acidiﬁcation due to NO
leaching. In addition,
relative low leaching of SO
(in comparison with S deposition)
occurred in many stream waters in Japan and southwestern China
even with high S deposition (Fig. 8), indicating considerable S sink
in the forested catchments, mainly due to SO
adsorption by soils
(Vogt et al., 2007; Huang et al., 2015).
3.2. Soil acidiﬁcation
Most of streams studied in Japan and China are not acidiﬁed,
although hillslopes receive acidic precipitation similar to or higher
than acid-sensitive areas of eastern North America and northern
Europe, and soil water in the root zone is often found to be acidiﬁed
(Miyanaga and Ikeda, 1994; Sato and Takahashi, 1996; Toda et al.,
2000; Asano and Uchida, 2005; Iizumi et al., 2005; Ebise and
Nagafuchi, 2002; Larssen et al., 2011).
Acidiﬁcation of soils has been widely observed in China. In the
1980s, soil pH decreased by 0.1e0.5 units on Lushan Mountain in
southern China (Pan et al., 1993). On Hengshan Mountain in
southern China, surface soil pH decreased by 0.5e1.1 for different
soil types from 1983 to 2001 (Wu et al., 2005). A decreasing trend in
surface soil pH during 1980e2009 was also found on Taishan
Mountain in northern China (Zhang and Li, 2010). Regionally, sig-
niﬁcant soil acidiﬁcation across major forest ecosystems and
grasslands was found in China during 1980se2000s (Yang et al.,
2012, 2015). Maximum decrease in soil pH was found in ever-
green forests (on average from 5.4 in the 1980s to 4.8 in the 2000s)
in southern China (Yang et al., 2012). Soil pH in the surface layer
declined signiﬁcantly over the last two decades across grassland in
northern China, with an overall decrease of 0.63 units (Yang et al.,
2012). The decrease in soil pH in China may be partly attributed
to acid deposition. However, soil pH was also decreased by 0.5 units
due to overuse of N fertilizers rather than acid deposition (although
the mechanism of soil acidiﬁcation is the same) (Guo et al., 2010).
In Japan and Korea, dominant soils are developed from granite
bedrock, and they are characterized by a low acid buffering capacity
(Yagasaki et al., 2001). Soil acidiﬁcation was also observed in Japan.
For example, the averagepH of the surface mineral soils in a forested
catchment in central Japan decreased from 4.5 in 1990to 3.9 in 20 03,
with the rate of decrease of 0.07 pH units per year noticeably higher
than many previously reported values in Europe and North America
(Nakahara et al., 2010). Surface soils of Andosols also showed de-
creases in soil pH, the dissolution of aluminum, and the formation of
precipitates, such as aluminum hydroxysulfate and basic iron sulfate
(Onodera et al., 2002; Takahashi and Higashi, 2013).
In comparison to soil, soil water in the root zone is more
Fig. 7. Ca
deposition estimated with origin in the desert areas (a; Larssen and Carmichael, 2000) and anthropogenic emission in 2005 (b; Zhao et al., 2011) in China.
L. Duan et al. / Atmospheric Environment 146 (2016) 55e69 61
dynamic and responds more quickly to acid deposition. In order to
supplement the existing monitoring data and gather new infor-
mation on Chinese systems, a set of integrated monitoring sites was
established through a Chinese-Norwegian cooperative project, the
Integrated Monitoring Program on Acidiﬁcation of Chinese
Terrestrial Systems (IMPACTS; Larssen et al., 2006). Precipitation
composition, as well as soil, water, and vegetation effects were
being intensively studied at ﬁve forested sites, which represented
acid-sensitive forested ecosystems in southern and southwestern
China. All sites were exposed to ambient acid deposition. Repre-
sentative for areas with high S and N deposition, the Tieshanping
site in southwestern China showed signiﬁcant soil acidiﬁcation
through signiﬁcantly higher leaching of strong acid anions (SO
) in soil solution (SS) than base cations (especially Ca
(Fig. 9;Larssen et al., 2011). Elevated H
ﬂuxes and associated high
inorganic monomeric aluminum (Al
)ﬂuxes were observed in soil
solution (Fig. 9). It should be noted that the Ca
input ﬂuxes (in
throughfall) are high compared to Ca deposition in most other
countries (due to elevated atmospheric transport of Ca
this input of Ca
, the ﬂux of H
in soil solution would be
Although forest dieback in Eastern Asia associated with acid
deposition has not been widespread, several phenomena such as
abnormal defoliation have been reported, not only in Japan (Izuta,
1998; Nakahara et al., 2010) but also in China (Larssen et al.,
2006; Wang et al., 2007) and Korea (Lee et al., 2005). For
example, severe defoliation is observed at two of the ﬁve IMPACTS
sites, Tieshanping and Luchongguan (near Guiyang in Guizhou
province), with 40e50% of defoliation and maximum 6% tree death
of Masson pine (Pinus massoniana) at the Tieshanping site (Larssen
et al., 2006; Wang et al., 2007). This was attributed to air pollution
and soil acidiﬁcation, although other stress factors such as insect
attacks and summer drought may have been important as well
(Wang et al., 2007). A literature review of the effect of acidiﬁcation
on forests in Japan indicated that the most important indicator for
soil acidiﬁcation is the Ca/Al molar ratio of soil water in the root
zone (Hirano et al., 2007). Japanese coniferous tree species such as
Japanese cedar and red pine are relatively sensitive to a reduction in
(Ca þMg þK)/Al molar ratio in soil solution (Izuta, 1998). These
results were similar as other earlier studies in Europe and North
America (e.g., Sverdrup and Warfvinge, 1993). Recently, soil acidi-
ﬁcation, with nutrient imbalance and low (Ca þMg þK)/Al molar
ratios (<10), was found to hamper the sound growth of both Jap-
anese cedar and Japanese cypress, and is one of the most likely
causes of the decline of temple and shrine forests in Kyoto (Ito et al.,
2011). The (Ca þMg þK)/Al molar ratio of the soil water at Tie-
shanping, southwest China was lower than 2.0 (Huang et al., 2014),
the critical limit widely accepted for Masson pines (Gao et al., 1992),
which are widely distributed in the subtropical areas of southern
and southwestern China. This species seems very sensitive to soil
SO42- concentration (meq·L-1)
S deposition (g·m-2·yr-1)
N deposition (g·m-2·yr-1)
S deposition (g·m
N deposition (g·m
Fig. 8. SO
concentrations in stream waters in Eastern Asia. Data from three regions of China: northeast (NE), west of southwest (SW1), and east of southwest (SW2)
(Chen et al., 2012; Xu et al., 2013). Data from Japan (Baba et al., 2001; Ebise and Nagafuchi, 2002; Farah et al., 2006; Iizumi et al., 2005; Kawakami et al., 2001; Kobayashi et al., 2013;
Komai et al., 2001; Nakahara et al., 2003; Nakahara et al., 2010; Sato and Takahashi,1996). Data from Korea (Jeon and Nakano, 2001). Left ﬁgures show S and N deposition modeled
by CMAQ (Zhao et al., 2013).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e6962
acidiﬁcation in comparison with some species in Europe and North
America, where 1.0 is widely used as a critical limit of
(Ca þMg þK)/Al (Sverdrup and Warfvinge, 1993).
Generally, chemical weathering of minerals is the only long-
term sustainable source of alkalinity neutralizing acid input in
North America and Northern Europe. Because of the distribution of
soils (such as Haplic Podzol and Albic Luvisol developed from
granite) with lower weathering rates in northeast China (Duan
et al., 2002), where the natural conditions such as the soil and
vegetation types are quite similar to those in North America and
Northern Europe, the soil cannot provide a strong acidity buffering
capacity. Although a much higher soil weathering rate occurred for
Haplic Luvisol in the west part of southwestern China (indicated as
SW1 in Fig. 8), the soil weathering rates of Haplic Acrisol in the east
part of southwestern China (indicated as SW2 in Fig. 8) and large
region in southern China were similar to those in northeastern
China (Duan et al., 2002). More signiﬁcant acidity buffering
capacity is therefore produced by several other processes including
The overall catchment budget (i.e., the difference between
throughfall deposition ﬂux and stream water output) indicated
considerable retention of N and S in the ﬁve forested catchments in
southern and southwestern China (Fig. 9;Larssen et al., 2011). As
for N, several studies based on the mass balance approach indicate
that after its deposition, NH
nitriﬁes to NO
(Larssen et al., 2011;
Huang et al., 2015). Since relatively little NO
uptake is reported in
N saturated forests due to limitation of other nutrients (Huang
et al., 2014), signiﬁcant NO
leaching occurs, which is associated
with strong acidiﬁcation in the rooting zone of the soil (Chen et al.,
2004). The acidifying effects of N deposition were also found in ﬁeld
studies at Dinghushan in the tropics of China (Lu et al., 2014), where
the increased Al and decreased base cation concentrations were
attributed to the loss in biodiversity (Lu et al., 2011). A literature
review on N leaching from 69 forest ecosystems at 50 sites
Fig. 9. Fluxes of SO
, and inorganic monomeric aluminum (Al
) in wet deposition (marked WO), throughfall (TF), soil-water (SS) and streamwater (W) in
Tieshanping (TSP), a subtropical forested catchment in Chongqing, southwest China (Larssen et al., 2011). Assume that throughfall ﬂuxes represent a reasonable estimate of total
deposition. The horizontal lines in each diamond box show the median, the range of the boxes is SD and the square inside each box shows the average. The numbers indicate ﬂuxes
for the individual years (1 represents 2001, 2 represents 2002, etc.). The size (height) of the diamond illustrate the variation between plots and between years; the numbers show
the variation between years only.
L. Duan et al. / Atmospheric Environment 146 (2016) 55e69 63
throughout China indicated that overall 22% of the N input via
throughfall was leached from soil, which is lower than the 50e59%
observed for European forests (Fig. 10;Fang et al., 2011). Note that
there may be large differences in NO
ﬂuxes between soil water
(Fig. 10) and stream water (Fig. 8). In China, elevated N leaching by
soil water (e.g., NO
-N concentrations exceeding 1.0 mg L
in forest ecosystems when they receive N deposition of more than
(Fang et al., 2011), while N deposition needs to be
above 25 kg ha
for considerable N leaching into stream
waters (Fig. 8). The big gap between the two thresholds may
attribute to the active denitriﬁcation occurring in the groundwater
discharge zones, indicated by much lower NO
ﬂuxes in stream
water than in soil water (Fig. 9;Larssen et al., 2011; Zhu et al.,
2013a). The N input thresholds for elevated NO
leaching in soil
water seems lower than those found in Europe and North America
(at approximately 10 kg ha
;Dise et al., 1998), where the
denitriﬁcation rate in head water catchments may be relatively
The main cause of substantial net loss of SO
catchment is likely the adsorption of SO
in soil layers (Vogt et al.,
2007; Duan et al., 2013; Huang et al., 2015), which produces OH
and can neutralize soil acidiﬁcation. It coincides with very low
leaching of SO
by stream waters in southern China and Japan with
high S deposition (Fig. 8). The anion exchange capacity is high in
these soils with low pH and a high content of Al oxides (Vogt et al.,
2007). By contract, soil nitriﬁcation produces H
to acidify soil,
while denitriﬁcation mainly neutralizes surface water acidiﬁcation.
It seems that the acidifying effect of N deposition may be more
important than S deposition in the well-drained tropical/subtrop-
ical soils. For some catchments where SO
saturation may occur,
with the SO
ﬂux in soil water similar as that in throughfall (Fig. 9),
reduction to sulﬁdes, as implied by high groundwater
table and effective denitriﬁcation, may be another SO
sink in the
sub-soil or even more likely in the groundwater discharge zones.
In summary, elevated Ca
deposition and signiﬁcant sinks of N
(denitriﬁcation) and S (sorption, including reduction, adsorption
and precipitation) explain the issue of why there is little surface
water acidiﬁcation in China. Other processes like SO
are only temporarily important (until approaching SO
saturation). Denitriﬁcation and SO
reduction may be more per-
manent sinks of acidity but depend on soil N status and S status. In
sorption is most likely reversible, implying that SO
desorption may delay the increase in soil water pH after a decrease
in S deposition. Moreover, modeling results by MAGIC indicated
that the current regulation of SO
emission abatement could not
signiﬁcantly increase soil water pH values, the (Ca þMg þK)/Al
molar ratio, or soil base saturation to the level of 2000 before 2050,
and the emission reduction of particulate matter would offset the
beneﬁts of SO
reduction by greatly decreasing the deposition of
base cations, particularly Ca
(Duan et al., 2013). Continuous
droughts in southwestern China in the future might also delay
acidiﬁcation recovery (Duan et al., 2013).
Excess nitrogen deposition has not only led to acidiﬁcation, but
also resulted in ecosystem eutrophication in Eastern Asia, shown as
changes in N dynamics, plant growth, or biodiversity. Atmospheric
N deposition could stimulate enzyme activities and accelerate N
transformation and cycling processes (Kim and Kang, 2011). For
example, N addition increased rates of net N mineralization and
nitriﬁcation, regulating organic matter decomposition (Mo et al.,
2006, 2007, 2008a; Mochizuki et al., 2012). Examination of six
forests in southern China and Japan indicated that in addition to
leaching, denitriﬁcation losses of NO
signiﬁcantly increased with
increasing N deposition (Fang et al., 2015), which increased soil N
emissions (Zhang et al., 2008).
Although N deposition could improve soil N availability and
result in an increased photosynthetic capacity and stimulation of
plant growth in N-limited ecosystems (Fan et al., 2007; Xia et al.,
2009; Bai et al., 2010), excess N input led to restriction to plant
growth or even damage to plants due to change in soil N status
(Fang et al., 2009; Lu et al., 2009; Xu et al., 2009), nutrient imbal-
ance (Yang et al., 2009), or reduction in net photosynthesis (Mo
et al., 2008b; Guo et al., 2014).
Biodiversity could also be signiﬁcantly affected by N deposition,
with the level depending on soil N status, vegetation composition,
dose and duration of N addition, and N requirements by different
species (Bai et al., 2010). Excessive N deposition normally reduced
biodiversity, including forest understory species (Lu et al., 2008,
2010a), grasses and forbs (Bai et al., 2010), and soil fauna (Xu
et al., 2006).
4. Critical loads
In Southeast Asia, comprising China, Korea, Japan, The
Philippines, Indo-China, Indonesia and the Indian subcontinent,
critical loads were ﬁrst computed and mapped as part of the impact
module of the Asian version of the Regional Air pollution INfor-
mation and Simulation model (RAINS-Asia) (Hettelingh et al.,
1995b). RAINS-Asia is used to assess abatement strategies for S
emissions through the application of the critical loads concept
(Streets et al., 1999). According to that study, low critical loads
(subject to high risk of acidiﬁcation) are found in southeastern Asia,
parts of the Himalayan range and the Tibetan plateau, parts of the
boreal forest in northern China, and the rain forest strip in south-
western India, while the dry regions in most of India and north-
western China show relatively high critical loads (Hettelingh et al.,
1995b). To improve the spatial resolution, critical loads were also
studied in many Asian countries such as Japan (Shindo et al., 1995;
Shindo and Fumoto, 1998; Hayashi and Okazaki, 2001), Russia
(Bashkin et al., 1995; Semenov et al., 2001), South Korea (Park and
Lee, 2001; Park and Shim, 2002; Park and Bashkin, 2001), India
(Gautam et al., 2010; Satsangi et al., 1995, 1998), and China (Zhao
Fig. 10. Throughfall N input versus dissolved inorganic nitrogen (DIN) leaching by soil
water in China (Fang et al., 2011).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e6964
and Seip, 1991; Duan et al., 2000a).
In most of the above studies, similar methods were applied in
Asia as in Europe, with some minor modiﬁcations (Duan et al.,
2000a; Posch et al., 2015). One of the most important modiﬁca-
tions of the widely used steady state mass balance (SSMB) method
was the consideration of base cation (BC) deposition (Zhao et al.,
2007a, b). As mentioned above, BC deposition, with a consider-
able fraction of anthropogenic origin, becomes the most important
source of ANC instead of weathering rate in China. Thus BC depo-
sition is considered variable instead of constant in the critical load
equation, which leads to the extended S-N-BC critical load function
(Zhao et al., 2007a). Both the maximum critical load of S (CLmax(S))
and the maximum critical load of N (CLmax(N)) would decrease
with the reduction of BC deposition (Fig. 11). If alkaline dust
emissions are controlled in the future, more efforts will be required
to prevent soil acidiﬁcation and ecosystem damage. This illustrates
the potential of future acidiﬁcation induced by reduced BC depo-
sition, if S and N deposition are not reduced correspondingly.
Based on the extended SSMB method, a map of critical loads for
S and N for China was developed under current BC deposition (Zhao
et al., 2009). It shows that the S critical loads in the northern and
northwestern China were generally higher than 30 kg ha
due to high weathering rates and natural deposition of base cations,
while the values could be lower than 0.3 kg ha
eastern China, with low temperatures and thus low weathering
rates, and in southern China, where both low weathering rates (due
to low content of weatherable minerals) and high vegetation up-
take of base cations occurs. Such results have been applied in
policy-making in China, such as the designation of the Acid Rain
Control Zones (Hao et al., 2001b) and the total emission control
planning. Under the current high BC deposition, the area exceeding
the CLmax(S) covered about 15.6% of mainland China (Zhao et al.,
2011). Unanticipated side effects of the control of primary PM
and thus BCs, particularly from the anthropogenic sources, may
wholly counteract the beneﬁts to regional acidiﬁcation of reduced
emissions of acid precursors, including large-scale abatement of
achieved since 2006 (Fig. 12). This suggests that policy-makers
may have little choice but to pursue even more stringent SO
Associated with rapid economic development, acid deposition
has become a major issue in Asia, especially in East Asia. Generally,
surface waters in Asia are not as sensitive to acid deposition in
comparison with soil. This is even true in acid forest soils in tropical
and subtropical regions, which are characterized by low mineral
weathering. East Asia is different from North America and North-
west Europe as the acidiﬁcation potential of atmospheric deposi-
tion is less than expected due to high base cation deposition,
deposition, derived from soil dust and particulate
matter from cement production and fossil fuel combustion.
Therefore, more attention should be paid to the trend of base cation
emission (both natural and anthropogenic) and deposition in Asia.
In addition, NO
denitriﬁcation and SO
adsorption are processes
that play a more prominent role in acid neutralization in soils of
East Asia than in Europe and North America.
Nitrogen deposition, especially of NH
, is of increasing concern
in Asia due to nitriﬁcation and nitrate leaching in N-saturated
ecosystems causing acidiﬁcation of soils and water. Enhanced NO
leaching has been observed in China and Japan. Although further
studies are needed, the acidifying effect of N deposition may be
more important than S deposition in well drained tropical/sub-
tropical soils due to high SO
As the biggest contributor of S and N emissions, China’s emis-
sions have begun to decrease in recent years, following Japan, South
Korea and some other countries. This has led to a decrease in S and
N deposition, and beginning of recovery from soil acidiﬁcation in
these countries. However, the large stores of adsorbed SO
expected to be desorbed, a process which delays the recovery of the
Fig. 11. Extended S-N-BC critical load function (Zhao et al., 2007a). To avoid exceedance, depositions of S, N, and BC should be limited below the shaded surface. More effort to
reduce S and/or N deposition should be taken to avoid acidiﬁcation as BC deposition is reduced, regardless of whether the critical load is currently exceeded (route P1 to P1*) or not
(route P2 to P2*). Small diagram shows the traditional S-N critical load function (Posch et al., 1995).
L. Duan et al. / Atmospheric Environment 146 (2016) 55e69 65
soil from acidiﬁcation. Thus, how quickly soils respond to decreased
deposition is uncertain. Risk of regional soil acidiﬁcation still exists,
as can be seen from critical load exceedance in large areas of East
Asia. Further studies on the effect of acid deposition in Asia are
therefore needed, not only for improving our understanding, but
also for supporting future policy-making.
The authors are grateful for the ﬁnancial support of the National
Natural Science Foundation of China (21221004), the State Envi-
ronmental Protection Public Welfare Project of China (201209001)
and the Collaborative Innovation Centre of Regional Air Quality. We
would also like to thank Julian Aherne and Douglas Burns for giving
us important comments on the manuscript.
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