Temporal Trend and Spatial Distribution of Speciated Atmospheric

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Temporal trend and spatial distribution of speciated atmospheric mercury emissions in China during 1978-2014 Qingru Wu, Shuxiao Wang, Guoliang Li, Sai Liang, CheJen Lin, Yafei Wang, Siyi Cai, Kaiyun Liu, and Jiming Hao Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b04308 • Publication Date (Web): 15 Nov 2016 Downloaded from http://pubs.acs.org on November 22, 2016

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Environmental Science & Technology

Temporal trend and spatial distribution of speciated atmospheric mercury emissions in China during 1978-2014 Qingru Wu,†,‡ Shuxiao Wang,*,†,‡ Guoliang Li,† Sai Liang,§ Che-Jen Lin,|| Yafei Wang, ⊥ Siyi Cai,† Kaiyun Liu,† and Jiming Hao†,‡

†

State Key Joint Laboratory of Environmental Simulation and Pollution Control, School of Environment,

Tsinghua University, Beijing 100084, China ‡

State Environmental Protection Key Laboratory of Sources and Control of Air Pollution Complex, Beijing

100084, China §

School of Natural Resources and Environment, University of Michigan, Ann Arbor, Michigan

48109-1041, United States ||

Center for Advances in Wter and Air Quality, Lamar University, Beaumont, TX 77710, USA

⊥

School of Statistics, Beijing Normal University, Beijing 100875, China

*Corresponding author. Tel.: +86 1062771466; fax: +86 1062773597. E-mail address: [email protected]

1

ACS Paragon Plus Environment


TOC 500

400

100 Share of Hg speciation (%)

Atmospheric Hg emissions (t)

600

80 60

Hg0 HgII Hgp

40 20 0

Coal-fired power plants Coal-fired industrial boilers Artisanal and small-scale gold production Lead smelting Zinc smelting Cement production Battery production Others

1980 1986 1992 1998 2004 2010

300

200

100

0 1980

1984

1988

1992

1996

2000

2004

2008

2012

2


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Abstract

2

Mercury pollution control has become a global goal. The accurate estimate of long-term mercury

3

emissions in China is critical to evaluate the global mercury budget and the emission reduction potentials.

4

In this study, we used a technology-based approach to compile a consistent series of China’s atmospheric

5

mercury emissions at provincial level from 1978 to 2014. China totally emitted 13,294 t of anthropogenic

6

mercury to air during 1978-2014, in which gaseous elemental mercury, gaseous oxidized mercury, and

7

particulate-bound mercury accounted for 58.2%, 37.1%, and 4.7%, respectively. The mercury removed

8

during this period were 2,085 t in coal-fired power plants (counting 49% of mercury input), 7,259 t in Zn

9

smelting (79%), 771 t in coal-fired industrial boilers (25%), and 658 t in cement production plants (27%),

10

respectively. Annual mercury emissions increased from 147 t in 1978 to 530 t in 2014. Both sectoral and

11

spatial emissions of atmospheric mercury experienced significant changes. The largest mercury emission

12

source evolved from coal-fired industrial boilers before 1998, to zinc smelting during 1999-2004,

13

coal-fired power plants during 2005-2008, finally to cement production after 2009. Coal-fired industrial

14

boilers and cement production have become critical hotpots for China’s mercury pollution control.

3



15

1 INTRODUCTION

16

Mercury (Hg) is a toxic pollutant that exists in the atmosphere as three operationally defined forms:

17

gaseous elemental mercury (Hg0), gaseous oxidized mercury (HgII), and particulate-bound mercury

18

(Hgp).1 Hg0 is the most abundant form (over 90%) in the atmosphere with residence time of 0.5–2 years.2

19

Due to its persistence, Hg could spread globally before depositing to the earth’s surface and

20

bio-accumulating in the environment.3 Correspondingly, local anthropogenic Hg emissions have led to

21

global Hg pollution and human and ecosystem health impact.4 Consequently, reducing atmospheric Hg

22

emissions has been a compulsive goal of Minamata Convention on Mercury.5

23

Constructing the long-term Hg emission inventory is important to assess the environmental Hg

24

budget and evaluate the emission reduction potential. Several studies constructed global Hg emissions

25

during 1850-2008.6, 7 However, these studies were resolved at continental or national level, and hence

26

they cannot support China’s Hg control strategy adequately. Researchers estimated that atmospheric Hg

27

emissions in China increased from 13 t in 1949 to 695 t in 2012.8 This inventory did not include the

28

emissions from intentional Hg use sectors (referring to production activities using Hg as raw materials,

29

including chlor-alkali production, caustic soda production, battery production, fluorescent lamp

30

production, thermometer production, and sphygmomanometer production in this study) and artisanal and

31

small-scale gold production (ASGM), which were significant emission sources in China before 2000.9 In

32

addition, the ASGM was even regarded as China’s largest emission source in recent 2010 inventory of

33

Arctic Monitoring and Assessment Programme and United Nations Environment Programme

34

(AMAP/UNEP).4 Therefore, missing these sources would lead to the underestimate of China’s Hg

35

emissions. Extensive efforts have been made to quantify China’s Hg emissions in specific years (e.g.,

36

199910, 200511, and 20104) and short-term periods (e.g., 1995-20039, 2005-201212, and 2000-201013).

37

These studies were different from each other in emission-estimation methodologies and sectoral

38

categories, making it challenging to develop long-term Hg emission inventories based on their results

39

directly.

40 41

In general, emissions from a particular source were calculated as its specific activity level multiplied by its source-specific emission factor.7,

9, 10, 12-14

Emission factors for other sectors, except for coal 4


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combustion, were rarely related with air pollution control devices (APCDs) in the inter-annual

43

inventories.9, 12 Thus, it is difficult to quantify the removed Hg of APCDs, which is also important in

44

evaluating Hg reduction potentials. In addition, Hg speciation is a key factor to assess the local deposition

45

and long-range transportation of Hg after emissions.15, 16 The continuous implementation of APCDs in

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China’s emission sources during past decades could influence the sectoral emission speciation of Hg.12, 13

47

Failure to track the changes of Hg speciation will reduce the accuracy of environmental impact

48

assessment of Hg emissions and control.

49

Rapid economic growth in China after the dawn of the Chinese Economic Reform (1978) have

50

consumed huge amounts of fuels and raw materials, which inevitably led to large emissions of air

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pollutants including Hg. To provide new insights on the evolution of sectoral and spatial Hg emissions

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accompanied by the economic growth and to support production-side Hg control strategies in China, this

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study quantified a consistent series of atmospheric Hg emissions from anthropogenic sources in China at

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the provincial level during 1978-2014. These results can be used in the input-output model to analyze

55

consumption-side drivers and make demand-side measures.17-20 The long-term spatial Hg speciation

56

profiles can also be used by the atmospheric transport model to assess the environmental benefit from

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emission reduction and to improve the knowledge of global biochemical cycle.

58 59

2 METHOD AND DATA

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Our method comprised three steps: (1) Emission factor generation; (2) Emission estimation; (3)

61

Uncertainty analysis.

62

Step (1): Emission factor generation

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This study applied two estimation methods to develop time varying Hg emission factors for

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anthropogenic sources in China during 1978-2014. We chose the estimation method for each source based

65

on its emission contribution and data availability. The detailed anthropogenic source categories and

66

methods applied were shown in Table S1.

67

The dynamic technology-based emission factors calculated by Equation 1 were generated by

68

considering the annual combustion/production processes and APCDs applied. Equation 1 was applied to 5



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estimate emissions from dominant emission sources, including coal-fired power plants (CFPPs),

70

coal-fired industrial boilers (CFIBs), residential coal combustion, other coal combustion, primary zinc

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smelting (abbreviated as Zn smelting), primary lead smelting (abbreviated as Pb smelting), primary

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copper smelting (abbreviated as Cu smelting), cement production, chlor-alkali production, caustic soda

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production,

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sphygmomanometer production. One significant challenge to compile historical Hg emission inventories

75

by using Equation 1 was to obtain the year-by-year application information on different

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production/consumption technologies and APCDs for the diverse emission sources. To address this, we

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acquired such information through literature review, expert judgments, best estimations, site investigations,

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field experiments, and interviews with industrial associations.

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battery

production,

fluorescent

lamp

production,

thermometer

production,

and

For other sources, we used Equation 2 to estimate the variation of speciated Hg emission factors,

80

referring to previous studies on estimating long-term Hg emissions.7,

14

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year-by-year emission factors fit transformed normal distribution function due to the dynamics of

82

technology change. The values of emission factor ef at a certain year was estimated by selecting values of

83

the parameters efa, efb, and s to correspond to the known or inferred time development pathway of

84

relevant technologies.

ef m ,l ,t = ∑ θ i −1M i ,l (1 − fi ,l ,t × w)∑ α l , j ∑ R j ,l Qm ,l , k (1 − Pj ,l ,k ,tη j ,l ,k ) i

j

k

(−

ef m ,l ,t = Qm ,l [(ef al − ef bl )e

( t − t0 ) 2 2 sl2

)

+ ef bl ]

It was assumed that the

(Equation 1)

(Equation 2)

85

where ef is the emission factor, in the unit of g t-1 or g corpse-1. m is the index for Hg species. l is the

86

index of emission sector. t is the calculated year, yr. i is the index of province. j is the index of

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combustion/production process (Table S1). k is the type of APCD combinations (Table S1). θ is the

88

transition factor which means the raw material consumption for unitary product yield (Table S2), %. M is

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Hg concentration in fuel or raw materials, g t-1. The detailed data are provided in Table S3 of Wu et al.21,

90

Table S4 of Zhang et al.13, and Table S3 in this study. f is the pretreatment rate of fuels or raw materials

91

consumed (Table S2), %. w is the Hg removal efficiency of the pretreatment, %. α is the application rate 6


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of different consumption or production processes (Table S4), %. R is the Hg release rate (Table S2), %.

93

For intentional Hg use sectors, the Hg release rate in this study meant total Hg loss to air due to the lacking

94

of detailed Hg release and APCDs information. Thus, we did not consider the APCD situations in these

95

sectors. For other sources, the Hg release rate means that Hg released into flue gas from fuel/raw

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materials. Q is the Hg speciation profile of different APCD combinations (Table S5), %. P is the

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application rate of different APCD combinations (Table S6), %. η is the probabilistic distribution of Hg

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removal efficiency of a certain type of APCD combination (Table S7), %.Hg speciation and removal

99

efficiencies were generated mainly from field experiments. We took the generation processes of Hg

100

removal efficiencies as an example to explain how field experiments were used in this study (Supporting

101

information S2). efa represents the emission level pre-1990 (Table S8), g t-1. efb is the best emission factor

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achieved in China (Table S8), g t-1. t0 is the time when the technology transition begins (pre-1990), yr. S

103

is the shape parameter of the curve. The largest emission factor for one sector from the literature was set

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as efa while the most recent localized emission factor was used as efb. Take the ASGM for example. The

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emission factor for ASGM was in the range of 400-15000 g t-1 and the most recent localized emission

106

factor was 2000 g t-1.7, 9, 11, 22, 23 Thus, efa and efb for the ASGM were set as 15000 and 2000 g t-1,

107

respectively.

108 109

The calculated emission factors by province for each sector were shown in supporting information (Calculated emission factor database. xlsx).

110

Step (2): Emission estimation

111

Based on the emission factors, speciated Hg emissions were calculated as follows.

Em ,l ,t = 10 6 × ef m ,l ,t × Al ,t

(Equation 3)

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where E is Hg emissions, t. A is the activity level, t or corpse. Provincial fuel consumption data were

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collected from China Energy Statistical Yearbooks.24 The data were revised according to the recently

114

released China Energy Statistical Yearbooks (2015).25 Provincial ASGM production data before 2002

115

were directly collected from yearbooks,26 while the data after 2002 were estimated based on the

116

information from China Gold Association that approximately 1%−3% of total gold production were

117

produced from ASGM.13 The coal combustion in cement production and iron and steel production was 7



118

calculated following the methods of Zhao et al.27, 28 Industrial production data by province were from

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relevant statistical yearbooks.24, 29-35

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Step (3): Uncertainty analysis

121

Monte Carlo simulations were used to produce the probabilistic emissions by taking into account the

122

probability distribution of key parameters. In equation 1, the key parameters included the Hg

123

concentration in fuel/raw materials and Hg removal efficiencies of APCDs. Hg concentration in fuel and

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raw materials fits lognormal distribution curve.13,

125

efficiencies of APCDs were shown in Table S7 and the generation processes were shown in supporting

126

information S2. In equation 2, the standard deviation of all investigated emission factors were used to

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generate the normal distribution of efm. In equation 3, the uncertainty of activity levels depend on the data

128

collection methods (Supporting information S3). We then ran the simulations for 10,000 times and got the

129

results in the form of a statistical distribution. Key characteristics of the simulation curves included P10,

130

P50, and P90 values. The P10, P50, and P90 meant that the probabilities of actual results less than

131

corresponding values were 10%, 50%, and 90%, respectively. The P10 and P90 values of the distribution

132

curve were set as the lower and upper limit of the simulation results. The (P50-P10)/P50 and (P90-P50)/P50

133

values were the lower and upper limit of the uncertainty range with a confidence level of 80%.

21

The distribution characteristics of Hg removal

134 135 136

3 RESULTS AND DISCUSSIONS 3.1 Hg emission trends by sector

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Atmospheric Hg emissions during 1978-2014 were 13,294 t, with 42% emitted before 2000 and 58%

138

emitted after 2000 (Figure 1). Annual Hg emissions increased from 147 t in 1978 to 530 t in 2014. Three

139

peaks in 1997, 2007, and 2011 could be identified in the emission estimate. During 1978-1997, Hg

140

emissions increased at an average annual growth rate (AAGR) of 5.5%. In this period, some present-day

141

small emission sources contributed large amounts of Hg emissions. For example, the summation of Hg

142

emissions from ASGM and battery production accounted for 19% of national emissions in 1997. The Hg

143

emissions decreased in 1998 triggered by Asian financial crisis, which led to the reduction of fuel

144

consumption.36 During 2000-2007, Hg emissions increased again at an AAGR of 5.7%, mainly due to the 8


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rapid increase (generally more than 10%) of fuel consumption.24 However, Hg emissions decreased by an

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AAGR of 0.8% during 2008-2010, due to the slight decrease of activity level growth and stricter SO2

147

control in some significant emission sources (e.g., CFPPs).37-40 In 2011, SO2 control devices were

148

gradually close to its maximal Hg reduction potential.41 For example, the installation rate of SO2 control

149

devices in CFPPs was next to 90%.41 Thus, we experienced the third peak in 2011. Hg emissions

150

decreased to 530 t in 2014 as a result of slowing activity levels, enhanced NOx control, and the

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elimination of backward production capacities (e.g., eliminating CFIBs with capacity less than 10 t/h).42,

152

43

153

spanned seven to nine years. For example, annual mean Hg0 concentration in Guiyang, southwestern

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China, increased from 8.40 ng m-3 in 2002 to 10.2 ng m-3 in 2010 with a mean annual rate of 0.16 ng m-3

155

yr-1 was found.44 The increasing in atmospheric Hg0 in Guiyang is consistent with the increasing

156

anthropogenic Hg emissions (7.1%) in this region.

To evaluate the emission trends, a preliminary assessment was conducted using monitoring data that

157

Major emission sources included CFPPs, CFIBs, NFMS, cement production, and battery production

158

during the study period. Hg emissions from CFPPs increased by 3.4 t annually before 2006. In 2006, the

159

emissions reached a peak (114 t). Afterwards, the co-benefits of SO2 and NOx control reduced Hg

160

emissions to 82 t in 2014. During 2005-2008, CFPPs were the largest Hg emission source in China.

161

CFIBs were the largest emission source before 1998. Emissions from CFIBs increased from 32 t in 1978

162

to 74 t in 2004 due to the increase of coal consumption. After 2005, Hg emissions from CFIBs generally

163

kept increasing, but the synergic Hg removal effect of APCDs slowed down the growth rate of Hg

164

emissions. Emissions from the NFMS increased from 46 t in 1978 to the peak of 227 t in 2004 when the

165

number of nonferrous metal smelters without proper APCDs greatly increased. Afterwards, Hg emissions

166

from the NFMS gradually decreased due to the elimination of small-scale smelters and enhanced SO2

167

emission control. In 2014, emissions from the NFMS were 116 t, approximately 22% of national

168

emissions. Zn smelting was the largest Hg emission source among all NFMS. Emissions from Zn smelters

169

led China’s Hg emissions during 1999-2004. The emission peak of Zn smelting was 126 t in 2004,

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accounting for 23% of national emissions. Emissions from ASGM increased from 7 t in 1978 to 47 t in

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1998. Since then, the emissions dramatically decreased mainly due to the ban of ASGM activities.45, 46 9



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Emissions from cement production continuously increased from 7 t in 1978 to 145 t in 2014. The

173

emissions from this sector were 5% of national emissions in 1978 and reached 27% in 2014. Cement

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production has been the largest emitter since 2009. Atmospheric Hg emissions from intentional Hg use

175

peaked in 1997 when the production of mercuric oxide battery reached the peak. After that, Hg emissions

176

from this sector continued to decrease. Except for the above sources, emission shares of other sources

177

were relative small. Atmospheric Hg emissions from residential coal combustion, other coal combustion,

178

other combustion sources (e.g., municipal wastes incineration), and iron and steel production increased

179

steadily during the study period.

180

Hg emissions in this study were 5%-57% smaller than the estimates in the same year by Streets et

181

al.10, Wu et al.9, and AMAP/UNEP4; but 2%-15% larger than the estimate by Zhang et al.13 The emission

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trend in this study was coincidentally similar to results reported by Tian et al.14 for the estimate before

183

1995 due to their lower emission estimate for Zn, Pb, Cu smelting and intentional Hg use, as well as

184

higher estimates for primary Hg ore mining. The difference in the estimates after 1995 was mainly caused

185

by the dramatic increasing emissions from Zn, Pb and Cu smelting. The detailed comparisons of our study

186

and previous estimates were shown in Supporting Information S4 and Figure S1-S3.

187

3.2 Spatial distribution trend of Hg emissions

188

Spatial distribution of Hg emissions in 1978, 2000, 2010 and 2014 were shown in Figure 2, Table

189

S10, and Table S11. In 1978, south central China and northwest China contributed the largest (23%) and

190

smallest (8%) shares. Emission shares from other regions were quite similar, from 16% to 19%. Liaoning

191

province discharged the largest amounts of Hg emissions (17 t, equal to 12% of national emissions).

192

Emissions from Zn/Pb smelting, CFPPs, and CFIBs jointly contributed to the large emissions in Liaoning

193

provinces (Figure S4 (f)). Emissions from other provinces were less than 15 t. The emission gap between

194

provinces was not obvious in this year. In 2000, emission share of south central China and east China

195

increased to 20% and 31%, respectively. Significant decrease of emission share was observed in northeast

196

China, from 16% in 1978 to 7% in 2000. The overall provincial emissions showed an upward trend

197

compared to that in 1978. The number of provinces with emissions more than 15 t increased to eleven.

198

Five of them were located in south central China and east China. Its emissions reached 42 t, 10


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approximately 11% of national emissions. Other significant emission provinces included Hunan and

200

Yunnan, emissions from which were 34 and 31 t, respectively. In 2010, Hg emissions in each region

201

continued to increase. The emission share of east China increased to 24% in this year due to its large

202

amounts of cement production. Emissions from Shandong and Jiangsu in this region increased to 46 and

203

32 t, respectively. Significant emission provinces in other regions included Henan, Hebei, Yunnan, Hunan,

204

and Inner Mongolia. Emissions of these provinces were larger than 30 t. From 2000 to 2010, the gap of

205

provincial emissions kept increasing. Emissions from Henan broke through 70 t while emissions from

206

Tibet were less than 1t. In 2014, the emission shares of east China and south central China reached 26%

207

and 28%, respectively. Emissions from large emitters such as Henan and Yunnan significantly decreased

208

and emission gaps between provinces slightly shrank.

209

During the study period, Henan province emitted the largest amount of Hg at 1,353 t, accounting for

210

10% of total cumulative emissions (Table S9). Provincial emissions were strongly affected by the

211

province-specific industrial activities and the implementation of APCDs over the study period. Before

212

2000, the dominant sources in Henan included the ASGM, Pb smelting and cement production (Figure S4

213

(a)). In 2000, these three sectors jointly contributed to 64% of Henan’s emissions. During 2001-2005,

214

Henan’s emissions increased sharply because of Pb smelting. Its emissions took up 51% of Henan’s

215

emissions in 2005. After 2005, Hg emissions from Pb smelting decreased while Hg emissions from

216

cement production increased, leading to varying Hg emissions in Henan province. In 2010, the large

217

emissions from cement production and Pb smelting kept Henan the top one emission province. With the

218

decreasing emissions from Pb smelting, the dominant emission source was cement production in Henan

219

province in 2014. Following Henan province were Hunan, Shandong, Hebei, Gansu, Liaoning, Jiangsu,

220

Yunnan, and Guangdong provinces (Figure S1(b)-(i)). Cumulative Hg emissions from these eight

221

provinces total 6,975 t (48% of total cumulative emissions). In Hunan, Gansu, Liaoning, and Yunnan

222

provinces, Hg emission trends were driven primarily by Pb and Zn smelting, CFPPs, and CFIBs. Cement

223

production were the dominant sources in Shandong, Hebei, Jiangsu, and Guangdong provinces.

224

3.3 Emission speciation

225

During 1978-2014, cumulative Hg0, HgII, and Hgp emissions were 7,732 (58.2%), 4,938 (37.1%) and 11



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627 t (4.7%), respectively. Speciation profile of emitted Hg experienced great changes during this period

227

(Figure 3), from 65/28/7 (Hg0/HgII/Hgp) in 1978 to 51/46/3 in 2014. Hg speciation profile has significant

228

impacts on the Hg transport distance.15 The decreasing proportion of Hg0 emissions signals an enhanced

229

Hg deposition in China.47, 48 Such shift was due to the high HgII proportion in the exhaust gases from

230

some large sources such as cement production, Zn smelting, and Pb smelting. In 2014, the HgII emissions

231

from cement production contributed 45% of national HgII emissions. Thus, if measures are taken to

232

control Hg emissions from cement plants, it is quite possible that we will see large benefits for local

233

environment in the coming years.

234

The fractions of Hg0 and Hgp showed an overall decreasing trend in each province, while the HgII

235

emission share gradually increased with variations in different provinces (Figure S5-S7). Fractions of

236

Hg0 in the emissions decreased by 22%-49% in Hainan, Qinghai, Tibet, Tianjin, and Xinjiang provinces.

237

For provinces lacking large-scale industrial activities, Hg emissions were mainly from intentional Hg use,

238

ASGM, and cremation. These sectors were characterized with high fractions of Hg0 emissions

239

(80%-100%), although their emission quantities were much smaller than those from large emission

240

sectors with lower Hg0 fractions (e.g., coal combustion and cement production). For other provinces, the

241

reduction of Hg0 speciation fraction was less than 20%, mainly due to the widespread application of

242

APCDs that modified the emission speciation. The maximal reduction of Hgp speciation fraction was 7%

243

in Jiangsu province due to the co-benefits of high efficient particulate matter (PM) control devices in

244

CFPPs, CFIBs and cement production.

245

3.4 Uncertainty analysis

246

The propagated overall uncertainties for the Hg emission estimates were shown in Figure 4. The 80%

247

confidence interval of uncertainty gradually reduced from (-30%, 44%) in 1978 to (-19%, 22%) in 2014.

248

The variation of Hg concentrations in coal, metal concentrates and limestone was the major contributor to

249

the uncertainties in the entire study period, accounting for 44%-62% of overall uncertainties. Estimates in

250

activity levels contributed 19%-34% of overall uncertainties and the Hg removal efficiency of APCDs

251

contributed to the remaining uncertainty, both of which can be improved with improved statistics and

252

field measurements of Hg removal of APCDs in future. Significant emitters with large uncertainties were 12


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Zn smelting (-63%, +81%), Pb smelting (-61%, +89%) and CFIBs (-46%, +49%) in 2014. The ASGM

254

activity was once prevailing gold smelting technology. Uncertainty from emissions of ASGM contributed

255

15% of the uncertainties of national emissions in 1998.

256

3.5 Implications

257

Atmospheric Hg emissions kept an overall rising trend in China during 1978-2011, which was

258

contrary to the global emission trends.49 Such an inconsistency resulted in substantially increasing global

259

attention on China’s Hg emissions, especially when other regions (eg., Europe, North America) have

260

achieved significant Hg reduction after 1990.7, 49 Therefore, China is facing large burden in Hg reduction.

261

The identification of the Hg removal trend in each sector in this study will highlight the Hg reduction

262

potential in different sectors and the prior control sources. According to the requirement of the convention,

263

CFPPs, NFMS (nonferrous metals smelting, refer in particular to Cu, Pb, Zn and industrial gold smelting),

264

wastes incineration, CFIBs, and cement production are the five key sources to be controlled. For CFPPs,

265

their Hg emissions and removal trends were shown in Figure 5(a). Hg removal during the study period

266

reached 2,085 t, approximately 49% of Hg input to CFPPs. The Hg removal trend in CFPPs indicated

267

ancillary benefits to atmospheric Hg abatement caused by the control of other pollutants (PM, SO2, and

268

NOX). In 2014, the application rate of electrostatic precipitator (ESP) and fabric filter (FF), flue gas

269

desulfurization towers (FGD), and selective catalytic reduction (SCR) reached 100%, 92.1%, and 83.2%,

270

respectively. These devices targeted to other pollutants synergically removed 73% of total Hg input to

271

China’s CFPPs in 2014. The average Hg removal efficiency is quite possible to improve given the issued

272

“ultra-low emissions” measure in 2015, which will promote Hg removal by using additional devices such

273

as ESP-FF or wet ESP. To further reduce Hg emissions, the China’s CFPPs can use co-benefit

274

enhancement techniques or dedicated Hg control technologies.50 However, the applications of these

275

technologies will be limited by the investment/operation costs and technology maturity and require new

276

policies’ support.

277

As to NFMS, their Hg input was much larger than that of other convention-related sectors. Take Zn

278

smelting for example (Emission trends of the other three NFMS subsectors were similar to Zn smelters).

279

Hg input to Zn smelting during 1978-2014 reached 9,239 t and approximately 7,259 t (approximately 13



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79%) were removed by APCDs (Figure 5(b)). Annual Hg removal increased from 6 t in 1978 to 701 t in

281

2014. The average Hg removal efficiency of APCDs in Zn smelters reached 93% in 2014. Thus, Hg

282

reduction potentials in Zn smelters are quite small. Hg emissions from municipal wastes incineration were

283

only 3.5 t in 2014. Although the incineration amount of municipal wastes kept increasing in the past

284

decade, municipal wastes incineration cannot be the dominant emission source in China in the near future

285

given the potential reduction of Hg concentration in the wastes due to the gradually elimination of

286

Hg-added products.

287

For CFIBs, total Hg removal during the study period was 771 t, only 25% of Hg input to this sector.

288

Although Hg removal efficiency in this sector has improved since 2005 (Figure 5(c)), their average Hg

289

removal efficiency was still only 42% in 2014. The low use proportion of washed coal (17%), low

290

application rate of high-efficient dust collectors (12%) and WFGD (12%), and large amount of

291

small-scale CFIBs (90% of total CFIBs in use) in 2014 promised larger Hg reduction potential in CFIBs

292

than that in CFPPs.51, 52 Relative policies will also promote Hg reduction in CFIBs.43, 53

293

As to the cement production (Figure 5(d)), Hg removal during the study period was 658 t,

294

accounting for 27% of total Hg input to cement production. Hg emission and removal trends in cement

295

production are determined by both APCDs and production processes. For cement production using

296

vertical shaft kiln or rotary kiln, Hg removal depended on APCDs. Thus, Hg removal by APCDs

297

increased from 1.0 t in 1978 to 22 t in 1996 when the application proportion of ESP/FF rose from 19% to

298

61%. However, for plants using the dry-process precalciner technology, the removed dust by ESP/FF was

299

recycled as raw materials, which led to the invalidation of the synergic Hg removal efficiencies of ESP/FF.

300

Therefore, when vertical shaft kiln and rotary kiln technologies were substituted by dry-process

301

precalciner technology after 1996, the co-benefits of Hg removal by ESP/FF were gradually diminished.

302

In 2014, the average Hg removal efficiency in cement production was only 13%, which implies great Hg

303

reduction potentials in this source. To reduce Hg emission from cement plants, it’s urgent nowadays to

304

use the dust shuttling technology to limit the build-up of Hg levels in the kiln dust and to fully realize the

305

synergic removal efficiency of ESP/FF. However, such measures still lack policy support in China.

306

In generally, CFIBs and cement production has substituted CFPPs and NFMS as the most prior 14


Page 15 of 25


307

control sources in China nowadays. In addition, attention should also be paid to non-convention related

308

sources such as iron and steel production. Hg emissions from this sector reached 32 t in 2014 and the

309

emissions are quite possible to increase due to the increasing trend of iron and steel production.

310

The spatial distribution trends of Hg emissions indicated different Hg control priority in different

311

provinces. Provinces with large atmospheric Hg emissions currently and cumulative Hg emissions (e.g.,

312

Henan, Hunan, and Shandong) should be the key control regions. In these provinces, strict air pollution

313

control measures and remediation technology for the contaminated sites should be applied. For provinces

314

such as Liaoning and Gansu provinces, although their atmospheric Hg emissions have been effectively

315

controlled currently, the large cumulative Hg emissions in the history have left large amount of

316

contaminated sites, especially the contaminated NFMS sites (Figure S4(e) and Figure S4(f)). Therefore,

317

it is urgent for these provinces to develop appropriate strategies for identifying and assessing sites

318

contaminated, and perform actions to reduce the risks posed by such sites.

319

Estimating long-term Hg emission trend in China is a challenging task due to the ever-evolving

320

transitions of industrial practices and APCD implementations. Given the best attempts, limited

321

availability on data documenting the details of raw materials, industrial practices, and APCDs still

322

represents a major barrier. In addition, some potential emissions sources such as secondary nonferrous

323

metal smelting, disposal of wastes from coal combustion were not included in this study due to data

324

unavailability, which required further study. To demonstrate the accuracy of the application information

325

and the robustness of the equations, calculated emission factors were evaluated against direct

326

measurements, as shown in Figure S8. It will be better to compare the emission inventory with direct

327

monitoring of atmospheric Hg concentration during the whole study period. However, the available data

328

in China are not sufficient to conclude a long-tern concentration trend. Although a preliminary assessment

329

has been conducted in this study, further evaluation of emission inventory using both chemistry and

330

transport model and atmospheric Hg observation data are needed in the future.

331

ACKNOWLEDGEMENT

332

This work was funded by 973 Program (2013CB430001), Natural Science Foundation of China

333

(21607090), and China Postdoctoral Science Foundation (2016T90103). The authors sincerely thank 15



Page 16 of 25

334

Zhonggen Li, Naiqiang Yan, Dingyong Wang, and Yongsheng Zhang for their sharing of Hg concentration

335

data and field experiment results.

336 337

SUPPORTING INFORMATION AVAILABLE

338

S1, Mercury (Hg) emission sources and parameters for estimation method; S2, Generation processes

339

of the distribution characteristics of Hg removal efficiencies; S3, Generation processes of the distribution

340

characteristics of activity data; S4, Comparison with previous studies; S5, Sectoral Hg emission trends in

341

typical provinces; S6, Spatial distribution of Hg emissions; S7, Hg speciation profiles and gridded Hg

342

emission inventories; S8, Comparison of calculated and tested emission factors. This information is

343

available free of charge via the Internet at http://pubs.acs.org/.

344

Gridded inventories are also available upon request.

16


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C. Estimating mercury emission outflow from East Asia using CMAQ-Hg. Atmos. Chem. Phys. 2010, 10 (4), 1853-1864. (49) Zhang, Y. X.; Jacob, D. J.; Horowitz, H. M.; Chen, L.; Amos, H. M.; Krabbenhoft, D. P.; Slemr, F.; St Louis, V. L.; Sunderland, E. M. Observed decrease in atmospheric mercury explained by global decline in anthropogenic emissions. Proc. Natl. Acad. Sci. U.S.A. 2016, 113 (3), 526-531. (50) United Nations Environment Programme (UNEP): Report of the group of technical experts on the development of guidance required under article 8 of the Convention; UNEP: Geneva, Switzerland, 2016. (51) National Energy Statistical Agency of China (NESA): China Energy Statistical Yearbook; NESA: Beijing, China, 2011. (52) Lin, Z. T. Research on the characteristic and inventory of mercury emission from coal-fired industrial boilers in China. Bachelor thesis, Tsinghua University, Beijing, China, 2015. (53) Ministry of Industry and Information Technology (MIIT); Ministry of Finance of the People's Republic of China (MF): Action plan of clean and efficient use of coal in industry; MIIT, MF: Beijing, China, 2014.

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FIGURES 600 Coal-fired power plants Coal-fired industrial boiler Residential boiler Other coal combustion Stationary oil combustion Mobile oil combustion Biomass incineration Municiple solid wastes incineration Cremation Large-scale gold production Artisanal and small-scale gold mining Copper smelting Lead smelting Zinc smelting Aluminium production Primary mercury mining Cement production Irron and steel smelting process Chlor-alkali production Caustic soda production Battery production Fluorescent lamp Thermometer Sphygmomanometer


500

400

300

200

100

0 1980 1984 1988 1992 1996 2000 2004 2008 2012

Year

Figure 1. National Hg emission trend by sector in China during 1978-2014.

21



Figure 2. Spatial distribution of atmospheric Hg emissions in (a) 1978, (b) 2000, (c) 2010, and (d) 2014.

22


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Share of Hg speciation (%)

100

0

Hg II Hg Hgp

80 60 40 20 0 1980

1986

1992

1998

2004

2010

Figure 3. Trend of Hg speciation profile during 1978-2014.

23



800

Uncertainty range


700

Atmospheric Hg emissions 600 500 400 300 200 100 0 1980

1984

1988

1992

1996

2000

2004

2008

2012

Year

Figure 4. Uncertainty range of atmospheric Hg emissions

24


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Figure 5. Hg emissions and removal in (a) CFPPs, (b) Zn smelting, (c) CFIBs, and (d) cement production. (WET– wet scrubber; ESP – electrostatic precipitator; FF – fabric filter; DC – dust collector, including cyclone, WET, ESP, and FF; IDRD – integrated dust removal devices; CFB-FGD – circulating fluidized bed – flue gas desulfurization; WFGD – wet flue gas desulfurization; WESP – wet electrostatic precipitator; SCR – selective catalytic reduction.)

25


Temporal Trend and Spatial Distribution of Speciated Atmospheric

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