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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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Mercury removal by magnetic biochar derived from simultaneous
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activation and magnetization of sawdust
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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
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University of Science and Technology, Wuhan, 430074, China
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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
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TOC Art
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INTRODUCTION
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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
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, noble
21-23
, etc. Among these,
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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
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37-39
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pyrolysis processes, resulting in the complex of preparation process of magnetic biochar (MBC).
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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.
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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
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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,
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which was used as precursor for the MBC. The precursor was then pyrolyzed at the setting
EXPERIMENTAL SECTION
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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
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temperature in N2 flow, the MBC was obtained. The obtained products were denoted as FeRMBCT,
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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.
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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.
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Mercury removal experimental apparatus and procedures. The Hg0 removal performances of
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MBCs were investigated by a mercury adsorption fixed–bed system, as shown in Figure S1. The Hg0
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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
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about 2 g quartz sand was used. The height of sorbent in the reactor was about 10 mm. The flow of
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flue gas feed into the reactor was 1.2 L· min-1. To identify the mercury speciation in the flue gas, a
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mercury speciation conversion system was equipped. The impingers containing 10% KCl solution
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(side Ⅰ) or 0.5mol·L−1 SnCl2/HCl solution (side Ⅱ) was placed between the reactor and mercury
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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.
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On the other side, the Hg2+ was reduced to Hg0 by SnCl2 and the total concentration of mercury
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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
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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)
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0 Where Hg in0 and Hg out represent the Hg0 concentration at the inlet and outlet of reactor, and
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T Hg out represents the total mercury concentration at the reactor outlet. t represents the accumulated
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time of each set of experiment, and t=120 min in this work. Since Hg0 and Hg2+ in the outlet flue gas
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might be both adsorbed on the adsorbents, the ߟads covers the Hg0 and Hg2+ adsorbed on the
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adsorbents and the ߟoxi only represents the part of Hg2+ present in flue gas. The accumulate mercury
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adsorption capacity were calculated by equation (4), where Q represents the mercury adsorption
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capacity (µg⋅g-1), v represents the gas flow rate (m3⋅h-1), m represents the mass of sorbents (g), and t
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represents the adsorption time (h).
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The pseudo-first order model was applied to analyze the Hg0 capture process, where the
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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)
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Where qt and qe represents the amount of mercury adsorbed in the sorbent at time t and at
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equilibrium time (µg⋅g-1), respectively. k1 represents the rate constant of pseudo-first order equation
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(min-1). The value of qe and k1 could be obtained by fitting the mercury adsorption curve.
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The experimental conditions are summarized in Table S3. In set I and set II, the Hg0 removal
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performances of different MBCs were studied at 120 °C under N2+4% O2 atmosphere. In set Ⅲ, the
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Hg0 removal performances of the optimal sample were investigated at a wide temperature range of
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30~350 °C with an intention to obtain the optimal reaction temperature. In set IV, the Hg0 removal
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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
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brominated AC (Br-AC) used for mercury removal in power plants was selected for comparison. The
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ultimate analysis and the textural properties are described in SI (Table S4 and S5). To obtain the
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equilibrium mercury adsorption capacity by the pseudo-first order kinetic model, the 1000–min
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mercury removal experiments were performed in set IV.
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Sample characterization. The textural properties of various samples are listed in Table S6.
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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
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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
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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)
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and Fe3C were appeared on the MBCs. Based on the XRD results, the transformation of iron species 7
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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
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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)
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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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However, with the increase of FeCl3/sawdust impregnation mass ratio to 2.0, some aggregation of
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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
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the formation of C=O group(289.0–289.2 eV
), while the content of oxygen-rich functional
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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)
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(f)
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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
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Fe1.5MBC500 with little amount of Fe3O4, while Fe3O4 is the unique iron species in Fe1.5MBC600
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(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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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
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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
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Figure 4 Effect of reaction temperature on Hg0 removal efficiency
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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
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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
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addition, the Hg0 could be oxidized over the catalytic oxidation sites and then migrate to the adjacent
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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
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(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
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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
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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
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