Mercury Removal by Magnetic Biochar Derived from Simultaneous

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Mercury removal by magnetic biochar derived from simultaneous activation and magnetization of sawdust Jianping Yang, Yongchun Zhao, Siming Ma, Binbin Zhu, Jun Ying Zhang, and Chuguang Zheng Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b03743 • Publication Date (Web): 10 Oct 2016 Downloaded from http://pubs.acs.org on October 10, 2016

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Environmental Science & Technology

1

Mercury removal by magnetic biochar derived from simultaneous

2

activation and magnetization of sawdust

3

Jianping Yang, Yongchun Zhao※, Siming Ma, Binbin Zhu, Junying Zhang※, Chuguang Zheng

4

State Key Laboratory of Coal Combustion, School of Energy and Power Engineering, Huazhong

5

University of Science and Technology, Wuhan, 430074, China

6

ABSTRACT. Novel magnetic biochars (MBC) were prepared by one step pyrolysis of FeCl3−laden

7

biomass and employed for Hg0 removal in simulated combustion flue gas. The sample

8

characterization indicated that highly dispersed Fe3O4 particles could be deposited on the MBC

9

surface. Both enhanced surface area and excellent magnetization property were obtained. With the

10

activation of FeCl3, more oxygen-rich functional groups were formed on the MBC, especially C=O

11

group. The MBC exhibited far greater Hg0 removal performance compared to the non-magnetic

12

biochar (NMBC) under N2+4%O2 atmosphere at a wide reaction temperature window (120−250 °C).

13

The optimal pyrolysis temperature for the preparation of MBC is 600 °C, and the best FeCl3/biomass

14

impregnation mass ratio is 1.5 g/g. At the optimal temperature (120 °C), the Fe1.5MBC600 was

15

superior in both Hg0 adsorption capacity and adsorption rate than a commercial brominated activated

16

carbon (Br−AC) used for mercury removal in power plants. The mechanism of Hg removal was

17

proposed, and there are two types of active adsorption/oxidation sites for Hg0: Fe3O4 and

18

oxygen-rich functional groups. The role of Fe3O4 in Hg0 removal was attributed to the Fe3+(t)

19

coordination and lattice oxygen. The C=O group could act as act as electron acceptors, facilitating

20

the electron transfer for Hg0 oxidation.

21

KEYWORDS. magnetic biochar (MBC), mercury removal, flue gas

0

1

ACS Paragon Plus Environment


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TOC Art

23 24

INTRODUCTION

25

Mercury pollution has attracted worldwide attention in recent years because of its high toxicity

26

in the human health and environment 1. Coal−fired power plants are considered as one of the largest

27

anthropogenic mercury emissions source 1. The US Environmental Protection Agency (EPA) issued

28

the Mercury and Air Toxics Standards (MATS) in December 2011 with an intention to limit the

29

mercury emissions from power plants 2. As the largest coal-consuming country in the world, it is

30

estimated that China emitted about 25%–40% of global mercury annually 3. In such a case, China

31

government also issued the latest Emission Standard of Air Pollutants for Thermal Power Plants

32

(GB13223-2011) which aims to reduce mercury emission from power plants.

33

Mercury exists in three forms in coal combustion flue gas: oxidized mercury (Hg2+), particle

34

bound mercury (HgP) and elemental mercury (Hg0). Hg2+ and Hgp can be removed by the wet flue

35

gas desulfurization (WFGD) system and dust control devices. However, Hg0 is difficult to be

36

removed by existing air pollution control devices (APCDs) because of its insolubility in water and

37

volatility. In recent years, extensive technologies have been developed for the reduction of mercury

38

emissions from power plants. Injection of sorbent upstream of the particulate control device is

39

considered as one of the most promising approach to reduce mercury emissions. Different mercury

40

sorbents have been developed for elemental mercury removal, such as activated carbon

41

metals

42

activated carbon is one of the most studied sorbent for capturing elemental mercury from flue gas.

8-10

, metal oxides

11-17

, calcium based sorbents18, 19, zeolite 20, fly ash

2


4-7

, noble

21-23

, etc. Among these,

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43

However, in activated carbon injection (ACI) processes, a large C/Hg ratio (10,000–100,000 wt./wt.)

44

is required to achieve high (>90%) Hg0 removal efficiency. The modification with additives such as

45

halides and sulphur could significantly improve the mercury adsorption capacity of activated carbons

46

24-26

47

activated carbons accounts for a large portion of the overall cost of ACI

48

develop economic and effective alternatives of activated carbon for mercury removal.

49

. However, the high cost limited its commercial usage in power plants, where the production of 27

. Thus, it is essential to

Biochar (BC), a carbon-enriched porous substance from the pyrolysis of biomass, has been 28-32

50

demonstrated for the possibility in mercury removal from combustion flue gas

51

low-cost carbon material, BC was considered as a prospective alternative to activated carbon (AC)

52

for the removal of Hg0 from flue gas because of its large specific surface and abundant porous

53

structures. However, the low mercury adsorption capacity limited the practicability of BC in mercury

54

removal. In such a case, the additional activation or modification process is generally required to

55

obtain high mercury removal efficiency. The general method for improving the Hg0 removal capacity

56

is: optimization of microstructure and surface functional groups

57

on the surface of BC, such as sulfur

58

generally complex and time consuming. It is attractive to active the BC and improve the mercury

59

removal capacity during the sample preparation process but without additional activation process.

36

, halides

33-35

. As a kind of

; loading of active components

28, 30-32

, etc. However, the modification process is

60

Although the BC sorbents could serve as a viable alternative to AC, the powdered BC injected

61

into flue gas would be captured by dust control devices together with fly ash. In such a case, like

62

other powdered sorbent, the separation of spent BC sorbents from fly ash would be difficult. This

63

will result in the release of mercury adsorbed in spent sorbent and caused secondary mercury

64

pollution during the utilization of fly ash. Thus, it is necessary to develop a facile method to separate

65

the mercury-laden sorbents from fly ash. The introduction of a magnetic medium (such as Fe3O4,

66

γ-Fe2O3) to sorbents is an efficient method to separate the sorbent efficiently by external magnetic

67

field. The chemical co-precipitation method is widely used for the introduction of Fe3O4/γ-Fe2O3 3



68

37-39

69

pyrolysis processes, resulting in the complex of preparation process of magnetic biochar (MBC).

70

Moreover, the co-precipitation reaction has a negative effect on the porosity of products 40. One step

71

pyrolysis of ferric chloride (FeCl3) laden biomass is a facile method for the magnetization of BC.

72

During the pyrolysis process, besides the magnetization property, the microstructure and surface

73

functional groups could be improved with the catalytic effect of FeCl3 41, 42. Surface functional

74

groups on the carbon-based sorbents have been proven to be able to enhance the adsorption capacity.

75

In such a case, the activation and magnetization property can be simultaneously obtained during

76

pyrolysis.

. However, the magnetization by chemical co-precipitation is generally after the complete of

77

In the present study, novel MBCs prepared by one step pyrolysis of FeCl3-laden sawdust were

78

employed for Hg0 removal in simulated combustion flue gas. The effects of pyrolysis temperature,

79

FeCl3/sawdust impregnation mass ratio, and reaction temperature on Hg0 removal performance were

80

investigated. The Hg0 removal performance under simulated flue gas (SFG) atmosphere was also

81

studied and compared with that of commercial ACs used for mercury removal in power plants.

82

Further, the mechanisms involved in Hg0 removal over MBC were identified.

83

84

Synthesis of the MBC. The MBCs were prepared by one step pyrolysis of FeCl3-laden sawdust, a

85

naturally abundant lignocellulose biomass. The proximate and ultimate analysis as well as the

86

chemical composition of the raw sawdust were shown in Table S1 and Table S2, respectively. Before

87

used, the sawdust was washed, dried, crushed and sieved to 200~300 mesh. The sawdust was then

88

immersed into FeCl3 aqueous solution for 2 h under continuous agitation. After that, the solid residue

89

was separated and dried at 105°C for 12 h. In such a case, the FeCl3-laden sawdust was obtained,

90

which was used as precursor for the MBC. The precursor was then pyrolyzed at the setting

EXPERIMENTAL SECTION

4


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temperature (500, 600, 700, 800 °C) for 1 h under N2 flow (0.5 L⋅min-1). After cooling to room

92

temperature in N2 flow, the MBC was obtained. The obtained products were denoted as FeRMBCT,

93

where R represents the impregnation mass ratio of FeCl3·6H2O to sawdust (R= 0.5, 1, 1.5, or 2 g/g),

94

and T represents the pyrolysis temperature (T=500, 600, 700, or 800 °C). For comparison, the BCs

95

pyrolyzed from the sawdust without the laden of FeCl3 (R=0 g/g) were also prepared, which were

96

denoted as BCT.

97

Characterization of the samples. The physical–chemical characteristics of the sample, including the

98

chemical composition, textural properties, morphology, magnetism, crystal structure and surface

99

chemistry, have been studied by various characterization technologies, which was described in

100

details in the SI.

101

Mercury removal experimental apparatus and procedures. The Hg0 removal performances of

102

MBCs were investigated by a mercury adsorption fixed–bed system, as shown in Figure S1. The Hg0

103

concentration in the flue gas was about 85 µg·m−3, which was monitored by an online mercury

104

analyzer (VM3000 Mercury Vapor Monitor). In each set of experiment, 50mg of sample mixed with

105

about 2 g quartz sand was used. The height of sorbent in the reactor was about 10 mm. The flow of

106

flue gas feed into the reactor was 1.2 L· min-1. To identify the mercury speciation in the flue gas, a

107

mercury speciation conversion system was equipped. The impingers containing 10% KCl solution

108

(side Ⅰ) or 0.5mol·L−1 SnCl2/HCl solution (side Ⅱ) was placed between the reactor and mercury

109

analyzer, respectively. When the stream was passed through side Ⅰ, the oxidized mercury (Hg2+)

110

0 was captured by KCl solution and the Hg0 concentration ( Hg out ) was measured by mercury analyzer.

111

On the other side, the Hg2+ was reduced to Hg0 by SnCl2 and the total concentration of mercury

112

0 T 0 2+ T = Hg out + Hg out ( Hg out ) was measured, where the difference between Hg out and Hg out is the

113 114

2+ Hg out

concentration. The total Hg0 removal efficiency (ߟT), the Hg0 adsorption efficiency (ߟads), and the Hg0 5



115

oxidation efficiency (ߟoxi) were defined as follows:

∑

ηT =

η ads

0 Hg in0 − ∑ 0 Hg out t

0

∑

∑ =

ηoxi =

Q=

t

∑

t

0

Hg in0

T Hg in0 − ∑ 0 Hg out

× 100%

(1)

× 100%

(2)

× 100%

(3)

t

0

t

t 0

∑

t 0

Hg in0

T 0 Hg out − ∑ 0 Hg out t

∑

t 0

Hg in0

1 t2 (Cin − Cout ) × v × dt m ∫t1

(4)

116

0 Where Hg in0 and Hg out represent the Hg0 concentration at the inlet and outlet of reactor, and

117

T Hg out represents the total mercury concentration at the reactor outlet. t represents the accumulated

118

time of each set of experiment, and t=120 min in this work. Since Hg0 and Hg2+ in the outlet flue gas

119

might be both adsorbed on the adsorbents, the ߟads covers the Hg0 and Hg2+ adsorbed on the

120

adsorbents and the ߟoxi only represents the part of Hg2+ present in flue gas. The accumulate mercury

121

adsorption capacity were calculated by equation (4), where Q represents the mercury adsorption

122

capacity (µg⋅g-1), v represents the gas flow rate (m3⋅h-1), m represents the mass of sorbents (g), and t

123

represents the adsorption time (h).

124

The pseudo-first order model was applied to analyze the Hg0 capture process, where the

125

equilibrium mercury adsorption capacity could be obtained. The pseudo-first order model could be

126

expressed as follows: dqt = k1 ( qe − qt ) dt

127 128

(5)

According to the initial conditions of t=0 qt=0 and t=t qt=qt, a modified form of equation (5) could be obtained: qt = qe (1 − e − k1t )

(6)

6


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129

Where qt and qe represents the amount of mercury adsorbed in the sorbent at time t and at

130

equilibrium time (µg⋅g-1), respectively. k1 represents the rate constant of pseudo-first order equation

131

(min-1). The value of qe and k1 could be obtained by fitting the mercury adsorption curve.

132

The experimental conditions are summarized in Table S3. In set I and set II, the Hg0 removal

133

performances of different MBCs were studied at 120 °C under N2+4% O2 atmosphere. In set Ⅲ, the

134

Hg0 removal performances of the optimal sample were investigated at a wide temperature range of

135

30~350 °C with an intention to obtain the optimal reaction temperature. In set IV, the Hg0 removal

136

performance of the optimal sample was studied under SFG atmosphere (4% O2 + 12% CO2 + 300

137

ppm NO + 1200 ppm SO2 + 10 ppm HCl + 8% H2O, balanced with N2). Moreover, a commerical

138

brominated AC (Br-AC) used for mercury removal in power plants was selected for comparison. The

139

ultimate analysis and the textural properties are described in SI (Table S4 and S5). To obtain the

140

equilibrium mercury adsorption capacity by the pseudo-first order kinetic model, the 1000–min

141

mercury removal experiments were performed in set IV.

142

143

Sample characterization. The textural properties of various samples are listed in Table S6.

144

Compared to the BC600, the BET surface area and pore volume of MBCs significantly increased. The

145

pyrolysis temperature is a key factor affecting the textural properties of MBCs. Generally, a high

146

pyrolysis temperature is favorable to the increase of BET surface area and pore volume. This is

147

because the high temperature could accelerate the release of small organic molecules and

148

unconverted compositions of sawdust, resulting in the development of pore structure

149

the excessive high pyrolysis temperature will decrease the BET surface area and pore development.

RESULTS AND DISCUSSIONS

41

. However,

150

The crystalline structures of BCs were studied by XRD, as shown in Figure S2. For the sample

151

of BC600, only one weak diffraction peak at 26.6° was observed, which could be attributed to the

152

amorphous carbon. However, the amorphous carbon was disappeared, while Fe3O4, FeCl2, FeO(OH)

153

and Fe3C were appeared on the MBCs. Based on the XRD results, the transformation of iron species 7



154

during pyrolysis process could be explained by reactions (1)-(6), which is consistent with the

155

interpretation of Liu et al. [42]. The FeCl3 preloaded on the sawdust would be initially hydrolyzed to

156

Fe(OH)3 and FeO(OH) during the drying process (reactions (1)-(2)). As the pyrolysis reaction

157

proceeds, some reducing components like H2, CO, and amorphous carbon were formed. In such a

158

case, the FeO(OH) would be reduced to Fe3O4 at high temperature (reactions 3-5). However, with

159

the further increase of pyrolysis temperature to 800 °C, a portion of Fe3O4 could be converted into

160

Fe3C due to the interaction with amorphous carbon. Moreover, FeCl2 also appeared after pyrolyzed at

161

500 °C, which could be attributed to the reduction of residual FeCl3 after drying stage (reaction (6)). FeCl3 + 3 H2O → Fe(OH)3 + 3HCl

(1)

Fe(OH)3→ FeO(OH) + H2O

(2)

6 FeO(OH) + H2 → 2Fe3O4 + 4 H2O

(3)

6 FeO(OH) + 4 CO → 2Fe3O4 + 4 CO2

(4)

6 FeO(OH) + 4 C → 2Fe3O4 + 4 CO

(5)

2FeCl3 + H2 → 2FeCl2 + 2HCl

(6)

162

The magnetic hysteresis curves of MBCs are shown in Figure S3. The samples showed

163

negligible coercivity and magnetization hysteresis, indicating the superparamagnetic characteristic

164

of the prepared materials. In such a case, the spent MBC sorbents after mercury removal could be

165

easily separated from fly ash by external magnetic field. The FeCl3/sawdust impregnation mass

166

ratio is the key factors affecting the magnetism of materials, since it determined the content of

167

Fe3O4 in the MBC matrix. Moreover, the pyrolysis temperature could also significantly affect the

168

magnetization property of materials. This is because the pyrolysis temperature could significantly

169

affect the iron species on the MBCs, which is consistent with the XRD results.

170

The morphology of the MBCs is shown in Figure 1. It demonstrated that Fe3O4 particles were

171

formed and associated on the MBC surface. At lower FeCl3/sawdust impregnation mass ratio

172

(0.5-1.5), a good dispersion of Fe3O4 particles on the MBC surface was achieved (Figure 1(a)-(c)). 8


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173

However, with the increase of FeCl3/sawdust impregnation mass ratio to 2.0, some aggregation of

174

Fe3O4 particles was appeared over Fe2.0MBC600 (Figure 1(d)). The pyrolysis temperature could also

175

significantly affect the morphology of BC. After pyrolyzed at 800 °C, the surface of BC800 is much

176

more smooth (Figure 1(e)), while the surface of Fe1.5MBC800 is etched seriously (Figure 1(f)). This is

177

because the FeCl3 could act as catalysts during the pyrolysis process, accelerating the dehydration of

178

carbohydrate polymers at high temperatures

179

sawdust might be changed, where the formation of heavy tars that may block pore structures was

180

inhibited. This will result in the development of porous structures in the BC matrix, and even cause

181

the etch of BC surface.

42

. In such a case, the decomposition pathway of

182

The functional groups on various BC surfaces were investigated by FTIR spectra. As shown in

183

Figure S4 (a), for the sample of BC600, only one small peak at about 1600 cm-1 was observed, which

184

could be assigned to aromatic C=C

185

appeared on the MBC: -OH groups (hydroxyl or carboxyl) (3400 cm-1), carbonyl C=O (1648 cm-1),

186

and aromatic C=C (1603 cm-1)

187

about 557 and 465 cm-1, suggesting the introduction of Fe3O4 particles into MBCs 39, 43. The intensity

188

of C=O group is obviously increased with the increase of FeCl3-laden value. Thus, FeCl3 could

189

accelerate the formation of organic functional groups, which is favorable to the improvement of Hg0

190

removal capacity. In such a case, both the simultaneous magnetization and activation were obtained

191

during the preparation of MBCs. Moreover, the pyrolysis temperature also plays an important role in

192

the formation of functional groups (Figure S4 (b)). With the increase of pyrolysis temperature from

193

500 to 800 °C, both the intensity of organic functional groups and Fe-O bonds were eliminated

194

gradually. This suggested that the high pyrolysis temperature could accelerate the decomposition of

195

sawdust and the functional groups were eliminated accordingly. XPS was also performed to

196

quantitively characterize the functional groups on BCs. As shown Table S7, the FeCl3 could promote

39, 41

. However, a large amount of functional groups were

39, 41

.The vibrations of Fe–O bonds in Fe3O4 were also observed at

9



44, 45

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197

the formation of C=O group(289.0–289.2 eV

), while the content of oxygen-rich functional

198

groups decreased with the increase of pyrolysis temperature, which is in line with the FTIR results.

199

Since chloride played an important role in the adsorption and oxidation of Hg in sorbents, the

200

content of chloride in the MBC was determined by XRF. The content of chloride in Fe1.5MBC500 and

201

Fe1.5MBC600 is about 7.3% and 3.9%, respectively, while the content of chloride in Fe1.5MBC700 and

202

Fe1.5MBC800 could be negligible.

(a)

(b)

(c)

(d)

10


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(e)

(f)

203

Figure 1 ESEM secondary electron images of various MBCs: (a) Fe0.5MBC600, (b) Fe1.0MBC600, (c)

204

Fe1.5MBC600, (d) Fe2.0MBC600, (e) BC800, (f) Fe1.5MBC800

205

Mercury removal performance.

206

Effect of pyrolysis temperature. The effects of pyrolysis temperature on the Hg0 removal

207

performance are showed in Figure 2. The NMBCs obtained at various temperature (BC500/600/700/800)

208

all presented poor Hg0 removal performance. However, the ߟT of MBCs (Fe1.5BC500/600/700/800) were

209

significantly increased, and the Fe1.5BC600 presented optimal Hg0 removal performance. It is evident

210

that the MBCs had a large BET surface area compared to the NMBC (Table S6). This will promote

211

the physisorption of Hg0 over MBCs. Furthermore, Table S7 showed that more C=O group was

212

appeared on MBCs compared to NMBCs. The C=O group could act as the active

213

chemisorption/oxidation sites for Hg0 46, 47, resulting in the improvement of Hg0 removal capacity of

214

MBCs. In such a case, the Hg0 removal reaction was attributed to the combined action of

215

physisorption and chemisorption/oxidation. However, the sample of Fe1.5MBC700 with larger BET

216

surface area presented poorer Hg0 removal performance than Fe1.5MBC600. This implied that the

217

chemisorption/oxidation of Hg0 played a more important role than the physisorption.

218

Although the content of C=O group in Fe1.5MBC500 is similar to that in Fe1.5MBC600, the Hg0

219

removal performance of Fe1.5MBC500 is obviously poor. This suggested that the iron species played a

220

significantly role in Hg0 removal as well, since FeO(OH) and FeCl2 are the dominant iron species in 11



221

Fe1.5MBC500 with little amount of Fe3O4, while Fe3O4 is the unique iron species in Fe1.5MBC600

222

(Figure S2). This suggested that Fe3O4 on the sorbent surface is responsible for Hg0 removal as well.

223

When the pyrolysis temperature exceeds 600°C, the Hg0 removal performance dropped obviously,

224

which could be attributed to the decrease of C=O groups on sorbents surface (Table S7). Particularly,

225

the ߟT of Fe1.5MBC800 decreased seriously to 45.3%. This is because most of the C=O groups are

226

disappeared and Fe3O4 is converted into Fe3C.

227

The role of chloride in Fe1.5MBC500 in Hg0 removal was also studied. As shown in Figure S5(a)

228

and S5(b), there is no obvious variation of the speciation of chloride in Fe1.5MBC500 before and after

229

Hg0 adsorption test. Moreover, the XRF analysis showed that the content of chloride was not

230

decreased obviously as well. Thus, the chloride existed in MBC did not play any role in Hg0

231

removal. Although the Hg0 removal capacity can be improved after modified by FeCl3 47, 48, the

232

improvement by FeCl2 is not obvious in this study. This is mainly because the chloride in MBC

233

existed in the form of FeCl2, the Cl–Fe coordination is significantly varied with that of FeCl3.

234 235

Figure 2 Effect of pyrolysis temperature on Hg0 removal efficiency (FeCl3/sawdust impregnation

236

mass ratio is 0 and 1.5)

237

Effect of FeCl3/sawdust impregnation mass ratio. The effect of FeCl3/sawdust impregnation mass

238

ratio on Hg0 removal efficiency is showed in Figure 3(a). The Hg0 removal capacities were

239

significantly improved with the increase of FeCl3/sawdust impregnation mass ratio. This could be 12


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240

due to the appearance of more functional groups on MBCs with the activation of FeCl3 (Figure

241

S4(a)), and the content of Fe3O4 was increased as well with the increase of FeCl3–laden value in

242

precursor. In such a case, more active chemisorption/oxidation sites for Hg0 were appeared, resulting

243

in the improvement of Hg0 removal performance. However, although the sample of Fe2.0MBC600

244

possessed more functional groups and Fe3O4 than Fe1.5MBC600, Fe1.0MBC600 and Fe0.5MBC600, it

245

exhibited lower Hg0 removal capacity. This indicated that the excessive FeCl3 preloading is not

246

beneficial for the Hg0 removal, which could be due to the following reasons: ( ) the aggregation of

247

Fe3O4 particles on the Fe2.0MBC600 surface, (

248

hypothesis could be confirmed by the microstructure of Fe2.0MBC600 (Figure 1(d)). Moreover, the

249

Hg0 removal performance of pure Fe3O4 and Fe3O4+biochar was studied for comparision. As shown

250

in Figure 3 (b), the Hg0 removal efficiencies of Fe3O4 and Fe3O4+biochar were about 38.5% and

251

59.8%, respectively, which are far lower than that of MBCs. This confirmed that the disperation of

252

Fe3O4 is important for the formation of mercury adsorption and oxidation sites. Furthermore, the

253

agglomeration of Fe3O4 made the BET surface area and pore volume dropped seriously (Table S6).

254

This will result in the decrease of physisorption of Hg0 over Fe2.0MBC600. During the Hg0 removal

255

reaction process, the gaseous Hg0 was firstly adsorbed on the sorbents via physisorption to form

256

Hg0(ad); then Hg0(ad) would be further converted through chemisorption or oxidation on the active

257

sites of sorbents. In such a case, the inhibition of physisorption caused by the decrease of BET

258

surface will further hinder the occurrence of chemisorption or oxidation.

) the variation of textural properties. The first

13



259 260

Figure 3 (a) Effect of FeCl3/sawdust impregnation mass ratio on Hg0 removal efficiency (pyrolysis

261

temperature is 600 °C), (b) Hg0 removal efficiency of Fe3O4 and Fe3O4+NMBC.

262

Effect of reaction temperature. The optimal sample of Fe1.5MBC600 was selected to study the effects

263

of reaction temperature on Hg0 removal performance. As shown in Figure 4, at the temperature of 30

264

°C, the sample presented poor Hg0 removal performance (ߟT = 44.6%). However, the Fe1.5MBC600

265

present excellent Hg0 removal performance (ߟT > 90%) at a wide reaction temperature (120-250 °C).

266

With the further increase of reaction temperature, the ߟT was decreased obviously. To interpret this

267

observation, the adsorption and oxidation behaviors of Hg0 at various temperatures were studied. The

268

results showed that the adsorption of Hg0 played a dominant role in the Hg0 removal reactions except

269

at 350°C. At the temperature range between 120 to 180°C, about 3.4–4.8% of ߟoxi and above 90%

270

ߟads was obtained. The further increase of temperature could yield higher ߟoxi; however, the ߟads was

271

sharply decreased. The improvement of oxidation capacity is mainly because the reactants could

272

attain more kinetic energy with the increase of reaction temperature, accelerating the reaction of Hg0

273

oxidation

274

from the sorbent surface, which will suppress the Hg0 adsorption.

49

. However, the higher reaction temperature might result in the desorption of mercury

14


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275 276

Figure 4 Effect of reaction temperature on Hg0 removal efficiency

277

Comparison of the Hg0adsorption capacities of Fe1.5MBC600 and commercial AC. HCl, SO2, NO, and

278

H2O are the main constituents in real combustion flue gas, which would significantly affect the Hg0

279

removal performance. Thus, the Hg0 removal performance of Fe1.5MBC600 was studied under SFG

280

atmosphere at the optimal reaction temperature of 120 °C. As shown in Figure 5 (a), the Hg0 removal

281

performance was slightly inhibited under SFG atmosphere compared to that under N2+4%O2

282

atmosphere. However, the ߟT was still maintained above 80% after 100 min mercury removal test,

283

which is superior than that of commercial Br-AC. After 1000 min mercury removal test, the Hg0

284

removal capacity of commercial Br–AC was close to breakthrough, while the Fe1.5MBC600 obtained

285

70% breakthrough. The accumulate Hg0 adsorption capacity in 1000 min was calculated according to

286

Equation (4) and showed in Figure 5 (b). It could be observed that the Hg0 adsorption capacities of

287

Fe1.5MBC600 and commercial Br–AC in 1000 min tests were about 953 µg⋅g-1 and 636 µg⋅g-1,

288

respectively. The equilibrium adsorption capacities of Fe1.5MBC600 and commercial Br–AC were

289

obtained by the pseudo–first order adsorption kinetic model. The kinetic data (qe and k1) and

290

correlation coefficient (R2) obtained from pseudo–first order kinetic model were shown in Table S8.

291

The results showed that the equilibrium adsorption capacities of Fe1.5MBC600 and commercial 15



292

Br–AC were about 1279.6 µg⋅g-1 and 690.7 µg⋅g-1, respectively. Thus, the prepared MBC was

293

superior in mercury adsorption capacity compared to the commercial Br–AC. Moreover, the Hg0

294

adsorption rate of Fe1.5MBC600, represented by the slope of the accumulate mercury adsorption curve,

295

was also superior than that of commercial Br–AC in the whole test process.

296 297

Figure 5 Hg0 removal performance of Fe1.5MBC600 and Br-AC under SFG atmosphere

298

Identification of involved reaction mechanism. To understand the involved reaction mechanism, the

299

adsorption and oxidation behaviors of Hg0 over Fe1.5MBC600 were studied in a long time test (1000

300

min). As shown in Figure 6 (a), during the first reaction stage (120 min), the outlet total mercury

301

(HgT=Hg0+Hg2+) concentration was much lower than the inlet HgT. This suggested that most of

302

mercury including both Hg0 and Hg2+ were adsorbed on the sorbent surface. As the reaction

303

proceeds, the mercury adsorption efficiency decreased gradually, while the mercury oxidation

304

efficiency increased (Figure 6 (b)). After 1000 min test, the total mercury concentration (HgT) at the

305

inlet and outlet of the reactor was balanced, suggesting that the adsorption of mercury reached

306

equilibrium. However, large amount (about 25%) of Hg2+ was detected at the reactor outlet,

307

indicating that the oxidation of Hg0 was still proceeded. The oxidation of Hg0 when the adsorption

308

reached equilibrium could be attributed to the catalytic oxidation. This suggested that different sites

309

for the oxidation and adsorption of Hg0 were existed on the Fe1.5MBC600 surface. During the first

310

reaction stage, Hg0 could be chemisorbed on the active adsorption sites of Fe1.5MBC600 surface. In 16


Page 16 of 26

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311

addition, the Hg0 could be oxidized over the catalytic oxidation sites and then migrate to the adjacent

312

non–catalytic adsorption sites. This is mainly because if the Hg0 is oxidized by a site and then

313

adsorbed on the same site, the Hg0 oxidation capacity would lose with the decrease of active sites. As

314

the reaction proceeds, all the non–catalytic adsorption sites were occupied, resulting in the escape of

315

resultant Hg2+ into stream during the migration process.

316 317

Figure 6 (a) Typical plot of measured outlet Hg0 and HgT concentration (b) mercury adsorption and

318

oxidation efficiency. On the basis of the characterization results and above analysis, it could be concluded that

319 320

different active adsorption/oxidation sites are responsible for the removal of Hg0 over Fe1.5MBC600.

321

(1) The role of Fe3O4

322

As shown in the XPS spectra of Fe 2p (Figure 7 (a)), three peaks at 710.2 ev, 711.2 ev and

323

713.0 ev were observed on the fresh Fe1.5MBC600, which could be assigned to Fe2+, Fe3+ in

324

octahedral coordination (Fe3+(o)) and Fe3+ in tetrahedral coordination (Fe3+(t)) of Fe3O4, respectively

325

45

326

the spent Fe1.5MBC600. This suggested that the Fe3+(t) coordination in Fe3O4 could act as an active

327

adsorption/oxidation sites for Hg0, resulting in the formation of iron amalgamation (reaction (8)).

328

The O 1s spectra in the fresh and spent Fe1.5MBC600 were shown in Figure 7 (b), which could be both

329

divided into three peaks at 530.1/530.5 ev (lattice oxygen in metal oxides, O2-), 531.6/531.9 ev

. After Hg0 removal experiment, obvious variation was observed on the Fe3+(t) coordination over

17



330

(chemisorbed oxygen, O*), and 533.1/533.7 ev (C–O). The presence of O2- in fresh Fe1.5MBC600

331

could be attributed to the presence of Fe3O4. The metal oxide could generate charge imbalance,

332

vacancies and chemical bonds on the sorbent surface, which could introduce chemisorbed oxygen

333

into sorbent 31. After Hg0 adsorption experiments, the content of O* decreased from 48.5% to 29.8%,

334

while the content of O2- increased from 36.3% to 47.5% (Table S9). This suggested that O*

335

participated in the adsorption of Hg0. The increase of O2- in spent Fe1.5MBC600 should be due to the

336

appearance of O2- in HgO, which was generated by reaction (9). From the above analysis, the

337

adsorption mechanism of Hg0 over Fe3O4 could be described as follows: (7)

Hg0(g)→Hg(ads)

(8)

Hg(ad)+Fe3O4→Hg−Fe3O4 Hg(ads)+O*→HgO 338 339

(9)

(2) The role of oxygen–rich functional groups Previous studies had shown that the oxygen–rich functional groups could promote the

340

adsorption/oxidation of Hg0

35, 49

341

Figure 7 (c), and the relative content of each carbonaceous functional groups were summarized in

342

Table S7. There are three peaks for the C 1s spectra in both the fresh and spent Fe1.5MBC600: 248.8

343

ev assigned to C=C, 286.3 ev assigned to C–O, and 289.0 ev assigned to C=O. After Hg0 adsorption

344

experiment, the content of C=O group in Fe1.5MBC600 significantly decreased from 20.1% to 9.2%.

345

In contrast, the content of C–O group increased from 12.7% to 20.7%. This suggested that the C=O

346

group was involved in the Hg0 adsorption/oxidation and was converted into C–O group during the

347

Hg0 removal process. This could be interpreted by the reactions (10) and (11).

. The C 1s spectra in the fresh and spent Fe1.5MBC600 are showed in

Hg0→Hg2++2e-

(10)

C=O+e-→C–O

(11) 18


Page 18 of 26

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348

During the Hg0 removal process, Hg0 could be firstly adsorbed on the sorbent surface (reaction

349

(7)) and was oxidized to Hg2+ by losing two electrons. Therefore, the oxidation of Hg0 will be

350

promoted by reaction (11), since the Fe1.5MBC600 surface could act as an electrode to accept electron

351

for Hg0 oxidation. The C=O group will act as electron acceptors, facilitating the electron transfer

352

process for Hg0 oxidation

353

group, resulting in the increase of C–O group content in spent Fe1.5MBC600. Thus, the existence of

354

significant amounts of oxygen–rich functional groups on the MBC surface is beneficial for the

355

adsorption and oxidation of Hg0.

(a)

31, 50

. After accepted electron, the C=O group will be converted into C–O

(b)

(c)

356

Figure 7 XPS spectra of fresh and spent Fe1.5MBC600 over the spectral regions of (a) Fe 2p, (b) C 1s,

357

(c) O 1s.

358

In summary, novel MBCs, prepared by one step pyrolysis of FeCl3−laden sawdust, showed

359

excellent Hg0 removal performance at a wide reaction temperature window (120−250 °C). Compared

360

to a commercial Br−AC used for Hg removal in power plants, the Fe1.5MBC600 presented better Hg0

361

adsorption capacity and adsorption rate. Meanwhile, the mechanism of Hg0 removal over MBC was

362

investigated. The Fe3+(t) coordination and lattice oxygen in Fe3O4 and C=O group in MBC could

363

both act as active adsorption/oxidation sites for Hg0. The spent MBC can be easily separated from fly

364

ash for recycle because of its excellent magnetic property. Thus, future research will focus on the 19



365

regeneration performance of deactivated MBC. Meanwhile, the effects of impurities like HCl, SO2,

366

NO, and water vapor in flue gas on Hg0 removal performance will be investigated as well, since this

367

is an important aspects in the utilization of the sorbents in realistic flue gas.

368

369

Supporting Information. Information regarding catalysts preparation, characterization of catalysts,

370

the experimental apparatus, Figures S1–S5, Tables S1-S9. This material is available free of charge

371

via the Internet at http://pubs.acs.org.

372

373

Corresponding Author

374

*Yongchun Zhao, Phone: 86–27–87542417. Fax: 86–27–87545526. E–mail: [email protected]

375

*Junying Zhang, Phone: 86–27–87542417. Fax: 86–27–87545526. E–mail: [email protected]

376

377

This research was supported by the National Key Basic Research Program (973) of China

378

(No.2014CB238904), the National Key Technologies R&D Program (2016YFB0600604), and the

379

National Natural Science Foundation of China (NSFC) No.51376074, 51206192, U1510201). Author

380

would like to thank anonymous reviewers for their critical comments. Author would also like to

381

thank analytical and testing center in Huazhong University of Science and Technology for the

382

assistance during experiments.

383

REFERENCES

384

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Mercury Removal by Magnetic Biochar Derived from Simultaneous

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