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Contents
1
An Overview of Atmospheric Reactive Nitrogen in China
from a Global Perspective . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .
Part I
2
1
Reactive Nitrogen Emission and Deposition in China
Anthropogenic Emissions of SO2, NOx, and NH3 in China . . . . . . .
Monitoring Atmospheric Nitrogen Deposition in China . . . . . . . . . .
13
41
3
Modelling Atmospheric Nitrogen Deposition in China . . . . . . . . . . .
67
4
5
Reactive Nitrogen Budgets in China . . . . . . . . . . . . . . . . . . . . . . . .
Part II
87
Contribution of Atmospheric Reactive Nitrogen to China’s
Air Pollution
6
Contribution of Atmospheric Reactive Nitrogen to Haze
Pollution in China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 113
7
Contribution of Atmospheric Reactive Nitrogen to Ozone
Pollution in China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 135
xi
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xii
Contents
8
Contribution of Atmospheric Reactive Nitrogen to Acid
Deposition in China . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 155
Part III
Impacts of Nitrogen Deposition on China’s Ecosystems
9
Impacts of Nitrogen Deposition on Forest Ecosystems in China . . . 185
10
Impacts of Nitrogen Deposition on China’s Grassland
Ecosystems . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 215
11
Impacts of Nitrogen Deposition on China’s Desert Ecosystems . . . . 245
12
Impacts of Nitrogen Deposition on China’s Lake Ecosystems:
Taking Lake Dianchi as an Example . . . . . . . . . . . . . . . . . . . . . . . . 263
Part IV
Reactive Nitrogen Regulation
13
Nitrogen Regulation in China’s Agricultural Systems . . . . . . . . . . . 297
14
National Regulation of SO2 and NOx Emissions in China . . . . . . . . 311
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Chapter 1
An Overview of Atmospheric Reactive
Nitrogen in China from a Global
Perspective
Abstract Atmospheric reactive nitrogen (N), as an important component of global
N cycle, has been significantly altered by anthropogenic emissions and consequently
induced worldwide impacts on air pollution and ecosystem services. Due to rapid
agricultural, industrial, and urban development, China has been experiencing a rapid
increase in reactive N emissions and deposition since the late 1970s. Based on a
literature review, this book summarizes recent research on (1) atmospheric reactive
N as a global environmental issue (Chap. 1), (2) the emission, deposition, and budget
of atmospheric reactive N (Chaps. 2, 3, 4 and 5), (3) the contribution of reactive N to
air pollution (e.g., haze, surface O3, and acid deposition) (Chaps. 6, 7 and 8), (4) the
impacts of N deposition on sensitive ecosystems (e.g., forests, grasslands, deserts,
and lakes) (Chaps. 9, 10, 11 and 12), and (5) the regulatory strategies for mitigation
of atmospheric reactive N pollution from agricultural and nonagricultural sectors in
China (Chaps. 13 and 14).
1.1
Atmospheric Reactive Nitrogen as a Global
Environmental Issue
Nitrogen (N) gas (N2) accounts for a volume fraction of approximately 78% of the
atmosphere, but it is unusable by most living organisms. Prior to the industrial
revolution, biological N fixation was the dominant pathway of new reactive N inputs
to the biosphere, while anthropogenic creation of new reactive N was negligible
1
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X. Liu and E. Du
(Vitousek et al. 2013). Human population was ultimately constrained by food
production during that long period, and the anthropogenic alternation of global N
cycle was minor compared with natural N flows (Fowler et al. 2013). The case has
changed since the early 1900s due to the industrial conversion of N2 to ammonia
(NH3), known as the Haber-Bosch process, which has substantially increased global
food production and sustained the growth of population. In turn, the growing
population further drives an increase in consumption of energy and other natural
resources, resulting in various environmental issues across the globe. The burning of
fossil fuels for energy production has increased the emissions of nitrogen oxides
(NOx), which is formed as a by-product during combustion. As a result, anthropogenic creation of reactive N has been dramatically accelerated (Gruber and Galloway
2008; Galloway et al. 2008). Specifically, global anthropogenic emissions of reactive N have risen from approximately 13 Tg N year 1 in 1860 (Galloway et al. 2004)
to approximately 100–115 Tg N year 1 in 2000 (Duce et al. 2008; De Vries et al.
2017), causing a fundamental change in the N cycle and a cascade of negative
impacts on earth systems (Galloway et al. 2003, 2008; Fowler et al. 2013).
As reactive N moves along its biogeochemical pathway, it causes a sequence of
effects, known as the N cascade (Galloway et al. 2003). The increase in reactive N
emissions to the atmosphere can cause air pollution via a combination of physical
and chemical processes. For instance, NH3 and NOx are both involved in the
formation of haze, and nitrate and ammonium are major compounds of atmospheric
particulate matter (Zhang et al. 2012; Wu et al. 2016). Moreover, anthropogenic
emissions of NOx play an important role in the formation of tropospheric ozone (O3)
during photochemical pollution episodes (Crutzen 1988). In view of the fact that
SO2 emissions have been successfully curbed in many countries, the contribution of
reactive N precursors (i.e., NH3 and NOx) to acid deposition has become increasingly important at a global scale (Galloway 2001; Dentener et al. 2006; Vet et al.
2014). These effects of reactive N on air pollution can expand to a larger extent via a
short- or long-distance transportation. Consequently, the occurrence of haze, O3
pollution, and acid deposition can result in various negative effects, such as damage
to human health, an alteration to climate systems, and a reduction of ecosystem
function and services (Schulze 1989; Kampa and Castanas 2008; Ramanathan and
Feng 2009; Du et al. 2017).
When deposited to the biosphere, reactive N can exert both beneficial and
deleterious effects on land and aquatic ecosystems, depending on the level of N
deposition and background N availability. As net primary productivity is widely
limited by N availability in natural ecosystems (Vitousek and Howarth 1991; Elser
et al. 2007; Bai et al. 2010; Song et al. 2012), N deposition can thus stimulate plant
growth and increase carbon (C) sequestration in these N-limited ecosystems
(De Vries et al. 2009, 2014; Du and De Vries 2018). When exceeding a certain
critical load, N deposition can, however, exert negative impacts on ecosystem health
and function. For instance, N deposition can cause a loss of plant biodiversity and
ranks the third driver of biodiversity loss after land use change and climate change at
the global scale (Bobbink et al. 2010; Sala et al. 2000). Other negative effects of N
deposition, such as nutrient imbalances, soil acidification, and increasing availability
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1 An Overview of Atmospheric Reactive Nitrogen in China from a Global Perspective
3
of toxic heavy metals, can partially counteract the fertilizing effect of N deposition
on plant growth (Bowman et al. 2008; Du et al. 2016; Tian et al. 2018; De Vries et al.
2014). In some extreme cases, excess N deposition can cause ecosystem N saturation, resulting in a reduction in biodiversity and primary productivity and
transforming the ecosystem to a net N source (Aber et al. 1998; Yue et al. 2019).
Currently, West Europe, the United States, China, and India are four hotpots of
reactive N emission and deposition (Dentener et al. 2006; Vet et al. 2014). However,
the temporal trends of reactive N emission and deposition differ in these regions. As
a result of substantial emission reduction of NOx and NH3, atmospheric deposition of
both nitrate and ammonium has leveled off in the Europe since the early 1990s
(Tørseth et al. 2012). In the United States, a reduction of NOx emissions has
substantially decreased nitrate deposition since the middle 1990s, while ammonium
deposition has grown continuously due to an absence of ammonia emission regulation (Li et al. 2016; Du 2016). In China, emissions of both NH3 and NOx kept
increasing continuously during the period 1980–2010 and drove an enhancement of
N deposition (Liu et al. 2013, 2016b). The Chinese government has started to curb
NOx emissions since 2010, and satellite observations indicate a reduction of average
NO2 column densities by 32% from 2011 to 2015 (Liu et al. 2016a). With the
increase in N use efficiency in agriculture together with stricter emission controls on
sulfur dioxide (SO2) and NOx, both wet and dry N deposition stabilized and even
showed decreasing trends in China (Liu et al. 2016b; Yu et al. 2019). In contrast, as
driven by growing consumption of N fertilizers and fossil fuels, emissions of NOx
and NH3 in India both have increased rapidly (Abrol et al. 2017). In view of an
absence of national mitigation strategies, the increase of emission and deposition of
reactive N will likely continue in India in next decades.
The environmental concerns have motivated integrated assessments of the change
in N cycle in the hotspot regions. The European Nitrogen Assessment first provided
an integrated and comprehensive evaluation of the sources, effects, and regulation
policy of reactive N at the European scale (Sutton et al. 2011). The Science Advisory
Board of the US Environmental Protection Agency also completed a national report,
to assess the current inputs, flows, and consequences of reactive N and to provide
specific management strategies to reduce the negative environmental impacts
(Doering et al. 2011). Recently, a group of Indian scientists have published a
national assessment on the sources of reactive N, the consequent environmental
and climate effects, and management options and policies (Abrol et al. 2017). These
assessments have substantially improved our understanding of anthropogenic alteration of N cycle and provide solid basis for N management in Europe, the United
States, and India. However, such an assessment is absent in China, even though
China currently is the largest emitter of reactive N and experiencing the highest level
of N deposition across the globe.
Since the early 2000s, monitoring, experimental, and modeling efforts have been
emerging explosively in China to assess the emission, atmospheric deposition, and
environmental effects of reactive N (Liu et al. 2011, 2017; Du et al. 2013; Tian et al.
2018). Based on a literature review, this book summarizes recent research on
(1) atmospheric reactive N in China from a global perspective (Chap. 1), (2) the
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X. Liu and E. Du
Fig. 1.1 Scheme of the emission, deposition, effects, and regulation of reactive nitrogen in the
environment
emission, deposition, and budget of atmospheric reactive N in China (Chaps. 2, 3, 4
and 5), (3) the contribution of atmospheric reactive N to air pollution (e.g., haze,
surface O3, and acid deposition) (Chaps. 6, 7 and 8), (4) the impacts of N deposition
on sensitive ecosystems (e.g., forests, grasslands, deserts, and lakes) (Chaps. 9, 10,
11 and 12), and (5) the regulatory strategies for the mitigation of atmospheric
reactive N pollution in China (Chaps. 13 and 14) (see Fig. 1.1).
In Chap. 1 (this chapter), Liu and Du briefly summarized all chapters’ key points
in this book in a global prospective, including each chapter’s main findings/conclusions and their eco-environmental and/or policy implications.
1.2
The Emission, Deposition, and Budget of Reactive
Nitrogen in China
China has been undergoing rapid socioeconomic development since the late 1970s,
and industrial production and agricultural utilization of reactive N have been simultaneously increased (Gu et al. 2012, 2015). In the meanwhile, growing energy
production from fossil fuels has also increased unintended N emissions (mainly
NOx) to the atmosphere. Along with the improvement of living standards, the
changing dietary structure (e.g., increasing consumption of meat and milk) and
lifestyle (e.g., increasing number of automobile per household) further stimulate
the creation, utilization, and emission of reactive N at national scale. The rapid
increase in anthropogenic reactive N inputs to the atmosphere, hydrosphere,
pedosphere, and biosphere has caused various environmental issues (Liu et al.
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1 An Overview of Atmospheric Reactive Nitrogen in China from a Global Perspective
5
2011; Zhang et al. 2012; Wu et al. 2016; Tian et al. 2018). As a result, the Chinese
government has started to curb atmospheric NOx emissions since 2010, while NH3
has not been regulated yet. In this context, an evaluation of reactive N emission,
deposition, and budget in China is essential for impact assessment and national N
regulation.
Atmospheric NH3 and NOx emissions have increasingly attracted the public
attention in China. In Chap. 2, Zhang Q et al. assessed spatial-temporal variations
in NH3, NOx, and SO2 emissions in China and related uncertainties. The results
indicate that the total reactive N emission increased continuously over the past three
decades, after which the NOx emission started to decline from 2011. However, the
NH3 emission is expected to remain consistently high levels. Atmospheric N deposition is one of the main pathways of external N inputs to terrestrial and marine
ecosystems. In Chap. 3, Liu et al. reviewed monitoring methods, monitoring networks, and monitoring results of N wet and dry deposition in China. In Chap. 4,
Zhang L et al. reviewed numerical modeling approaches for simulating atmospheric
N deposition and their recent applications on N deposition to China. Both monitoring and modeling results indicated an increase of N deposition during the 1980s and
2000s. The annual total N deposition to China was estimated to be 7.9–20.1 Tg N
year 1, and reduced N (NHx) accounted for 60–80% of the total N deposition. In
Chap. 5, Gu et al. evaluated the budgets for reactive N in China from 1980 to 2015.
The results indicate a tripling of anthropogenic reactive N creation, which was
associated with an even more rapid increase in N fluxes to the atmosphere and
hydrosphere. These increasing N flows may cause consequent threats to human
health, the sustainability of crop production and sensitive ecosystems, and the
environment.
1.3
The Contribution of Reactive Nitrogen to Air Pollution
in China
Reactive N plays a significant role in atmospheric chemistry and is closely associated
with air pollution, such as haze, troposphere O3, and acid deposition (Liu et al.
2017). In Chap. 6, Pan et al. reviewed the mechanisms of aerosol formation related to
reactive N and the contribution of reactive N to haze pollution in China. Specifically,
nitrate and ammonium are major compounds of atmospheric particulate matter,
contributing approximately one-third of fine particulate matters (e.g., PM2.5).
Although the concentration of amines in the atmosphere is probably two or three
orders of magnitude lower than that of ammonia, amines can significantly assist the
growth of both neutral and ionic clusters. In Chap. 7, Feng et al. reviewed the
mechanism of O3 formation and summarized the spatial-temporal patterns of
ground-level O3 in China. Generally, high O3 concentrations frequently occur in
China’s major metropolitan agglomerations such as the Jing-Jin-Ji region, the
Yangtze River Delta, and the Pearl River Delta. Moreover, the ambient O3
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X. Liu and E. Du
concentration at almost all monitoring sites has exceeded the threshold of sensitive
plant growth. The results also suggest current O3 is threatening human health and
ecosystem services. In Chap. 8, Yu and Duan overviewed the contribution of
reactive N to acid deposition, the acidification of soil and surface water,
acidification-buffering processes, and future prospects of acid deposition control in
China. The results indicate N deposition contributed increasingly to acid deposition
in China in recent decades. Surface waters across China are generally not sensitive to
acid deposition, while acidifying effect of N deposition on soil is likely more
important than S deposition due to N transformations.
1.4
The Impact of Nitrogen Deposition on Sensitive
Ecosystems in China
The enhancement of N deposition in China has aroused increasing concerns about its
effect on ecosystem health and function (Du et al. 2015). Forest covers more than
one-fifth of the national land area in China and provides fundamental ecosystem
services. In Chap. 9, Du et al. summarized current understanding of the N deposition
impacts on soil chemistry and N transformation, soil microorganisms and enzymes,
plant physiology and biodiversity, and ecosystem carbon balance in China’s forests.
Experimental results and modeling estimates generally indicate a fertilization effect
of N deposition on forest growth and consequent C sequestration. However, highlevel N deposition has been increasingly evidenced to cause N leaching loss, soil
acidification, nutrient imbalance, increased N2O emissions, and decreased soil CH4
uptake, which likely offsets the positive effect on ecosystem C storage over time.
Meanwhile, N deposition likely changed both species composition and richness of
plant and soil microbial communities in China’s forests.
Grasslands account for 40% of national land area in China and have an essential
role in regional economic development and ecological security. In Chap. 10, Lü et al.
reviewed the impacts of N deposition on China’s grasslands by focusing the changes
of above- and belowground biodiversity and biogeochemical (carbon and nutrient)
cycling. The results indicate that N deposition can substantially increase soil N
availability, alter fluxes of greenhouse gases, and threat plant biodiversity in the
grasslands of China. The impacts of N deposition on other ecosystems, such as
deserts and lakes, were also assessed. In Chap. 11, Zhou X et al. overviewed the
response of China’s desert ecosystems to increasing N deposition. The results
indicate that desert ecosystems are sensitive to increasing N deposition, and the
effect of N deposition is strongly interacted with precipitation. Elevated N deposition
has significantly influenced aquatic ecosystems, especially with regard to their N
budgets and phytoplankton growth potentials. In Chap. 12, Zhou F et al. reviewed
the effect of N deposition on eutrophic lakes by taking Lake Dianchi as an example.
They estimated that annual N deposition accounted for 15.7–16.6% of total N loads
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1 An Overview of Atmospheric Reactive Nitrogen in China from a Global Perspective
7
and the proportion was as high as 27–48% in May and June when toxic blooms were
initiated and proliferated.
1.5
Regulation Strategies of Reactive Nitrogen Emission
As increasing atmospheric emission and deposition of reactive N in China have
substantially contributed to air pollution and significantly threatened ecosystem
health, national N regulation is of great importance to mitigate negative effects of
reactive N in the environment. Agricultural systems and energy production systems
are major sources of reactive N emission to the atmosphere. In Chap. 13, Yan et al.
reviewed the N inputs and losses of reactive N in China’s agricultural systems as
well as the strategies to increase N use efficiency. They show that the total application of N fertilizer has increased by 150% in China’s croplands from 1980 to 2010,
while N use efficiency has decreased continuously. They also reviewed N management strategies to improve N use efficiencies, such as integrated soil-crop system
management, knowledge-based N management and livestock manure partially substitute synthetic fertilizer, as well as national N regulation projects to reduce N losses
from agricultural systems, such as soil testing and fertilizer recommendation program and “zero growth of the fertilizer and pesticide consumption by 2020” plan. In
Chap. 14, Zhao and Xia evaluated the effect of recent national policy strategies of
energy conservation and emission reduction on the emissions of SO2 and NOx. SO2
emission has reduced substantially since 2006 due to the improved use of flue gas
desulfurization in the power sector and implementation of new emission standards in
key industrial sources; NOx emission has started to decrease from 2011 due to the
penetration of selective catalytic/non-catalytic reduction systems in the power sector.
Transportation is playing an increasingly important role in regional air pollution,
with the emissions from stationary sources gradually controlled.
1.6
Outlook
China has experienced rapid economic growth via industrialization and urbanization
over the past four decades (1978–2018), while this growth has also consumed
increasing energy and raw materials and induced various environmental issues.
The increase of reactive N emission and deposition is one example that has attracted
public concerns due to its contribution to air pollution and negative effects on
ecosystem services. In general, atmospheric reactive N emission is closely associated with the occurrence of secondary aerosol (e.g., PM2.5) pollution, increasing
tropospheric O3 concentrations, and acid deposition. High-level N deposition in
China can significantly alter structure and function of various ecosystems in China,
especially those (semi-)natural ecosystems in eastern and southern regions. In view
of these environmental issues, the Chinese government has recently implemented
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X. Liu and E. Du
increasing measures to improve N use efficiency in agricultural systems and reduce
reactive N emission from energy production and transportation systems (Liu et al.
2016c). Moreover, a national reformation of economy systems is also expected in the
near future. These changes promise a shift from increase to decrease in reactive N
emission and deposition in China. Future research efforts on the trends and
eco-environmental and human health effects of atmospheric reactive N will have
critical implications for future national reactive N regulations in China and other
rapidly developing countries.
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