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Mercury removal from flue gas by magnetic iron-copper oxide modified porous char derived from biomass materials
Fuel 256 (2019) 115977
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Mercury removal from flue gas by magnetic iron-copper oxide modified porous char derived from biomass materials
T
Wei Yang, Ying Li, Shuo Shi, Hui Chen, Ye Shan, Yangxian Liu
⁎
School of Energy and Power Engineering, Jiangsu University, Zhenjiang, Jiangsu 212013, China
ARTICLE INFO
ABSTRACT
Keywords: Elemental mercury Flue gas Iron-copper mixed oxides Biomass porous char Magnetic catalyst
In this work, a magnetic iron-copper oxide modified porous char derived from biomass materials was prepared by microwave activation and ultrasonic-assisted impregnation method to remove Hg0 from flue gas. The effects of operating parameters and flue gas components on Hg0 removal performance as well as physicochemical properties of samples were studied. The reaction mechanism and the regeneration performance of samples were also investigated. The experimental results display that the biomass-based porous char derived from microwave and steam activation exhibits a larger pore volume, higher surface area and well-developed pore structure. The ultrasonic treatment promotes the dispersion of active components on the surface of samples. The CuFe0.3/ WSWU10(500) sample exhibits good Hg0 removal performance, and about 90.58% of Hg0 is captured from flue gas at 130 °C. Increasing SO2 concentration inhibits Hg0 removal. The presence of O2 and NO promotes the removal of Hg0. Low concentration of water vapor (> 4%) improves Hg0 removal performance, while excessive water vapor plays an opposite role. The presence of copper/iron active components, chemisorbed oxygen and lattice oxygen over the surface of adsorbent is conducive to the removal of Hg0. The modified sample exhibits a good regeneration performance.
1. Introduction Mercury has received more and more global attention due to its bioaccumulation, mobility and toxicity in both food chains and ecosystems [1]. Coal-fired utility boilers release large amounts of mercury into the atmosphere each year, which is considered to be the largest source of anthropogenic mercury emissions. In the United States, mercury emissions from coal-fired utility boilers account for approximately one-third of the 150 tons of mercury emissions [2]. Coal combustion contributes about 38% of anthropogenic mercury emissions in China [3]. Mercury in coal-fired flue gas is mainly composed of gaseous mercury species and particulate-bound mercury (Hgp). Gaseous mercury species are mainly involved in oxidized mercury (Hg2+) and elemental mercury (Hg0) [4]. The existing pollution control devices such as wet flue gas desulfurization device (WFGD) and particulate control device can be used to capture oxidized mercury and particulate-bound mercury, respectively. However, due to its low solubility in water and high volatility, the elemental mercury (Hg0) in flue gas is difficult to be effectively captured by the above-mentioned existing pollution control devices [5]. Therefore, the main focus of mercury emission control is the effective removal of elemental mercury (Hg0) from flue gas. Up to date, a number of mercury control methods have been developed for ⁎
the capture of elemental mercury from coal-fired flue gas, such as catalytic oxidation, sorbent adsorption, chemical oxidation, photochemical oxidation, photocatalytic oxidation and so on [6–11]. The activated carbon (AC) injection method, which is performed in the upstream of the particulate control device, is recognized as one of the most effective techniques for elemental mercury removal. However, higher removal costs and C/Hg0 ratios limit the widespread application of this method [12]. Moreover, the development of mercury removal technologies is also extremely limited by the separation and recovery of spent sorbents/catalysts from fly ash. Biochar, a byproduct from pyrolysis of biomass under oxygen-limited environment, has been used to remove elemental mercury from flue gas [13]. With lower cost, it could be a promising substitute for activated carbon in the field of mercury removal [14]. However, lower mercury removal performance and undeveloped pore structure limit the application of biomass char. Thus, some chemical or physical modification methods are applied to improve the mercury capture performance. The methods mainly involve the introduction of active sites on the surface of biochar, such as halides, sulfur, metal oxides, etc [15–17]. However, it is still a huge challenge to separate the spent modified biochar from fly ash. Therefore, it is necessary to develop some easily separable biochars to control mercury emissions.
Corresponding author. E-mail address: [email protected] (Y. Liu).
https://doi.org/10.1016/j.fuel.2019.115977 Received 16 May 2019; Received in revised form 22 July 2019; Accepted 6 August 2019 Available online 13 August 2019 0016-2361/ © 2019 Elsevier Ltd. All rights reserved.
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In order to separate the spent sorbents from fly ash, the addition of the magnetic materials such as γ-Fe2O3 and Fe3O4 to sorbents may be an efficient way [18,19]. Some related results [19–24] suggest that the γFe2O3 and Fe3O4 can capture Hg0 from flue gas to a certain extent, but its Hg0 removal capacity is relatively low. Many studies have shown that copper oxide (CuOx) is a promising Hg0 capture material due to its higher stability and catalytic activity for Hg0 removal at low temperature [25]. Du et al. [26] prepared a CuOx-based sample and found that it exhibited a good performance for Hg0 removal. In addition, Yang et al. [27] found that CuO can effectively improve the Hg0 removal capacity. Thus, the incorporation of copper oxides into Fe-based sorbents may contribute to the improvement of mercury removal efficiency. Related research shows that the activation of microwave and steam can greatly improve the pore structure of biochars [15]. Moreover, ultrasonic treatment is beneficial to the dispersion of active ingredients on the surface of samples [13]. Therefore, in this work, a novel magnetic iron-copper oxide modified biomass-based porous char derived from the activation of microwave and steam was prepared by using an ultrasonic- assisted impregnation method, and was used to control Hg0 emissions from flue gas. The physicochemical properties of the samples were investigated by various characterization methods such as XPS, BET, SEM, XRD and magnetism analysis. The effects of Cu/Fe molar ratios, reaction temperatures, loading values, calcination temperatures on Hg0 capture efficiency were analyzed, as well as the effects of flue gas constituents, microwave activation and ultrasonic-assisted impregnation. Furthermore, both the reaction mechanism and the regeneration performance of the samples for Hg0 removal were also researched. The results of this research will provide some references for the development of this new technology in the field of Hg0 emission control.
a water bath at 40 °C for 30 min. Then, the mixture was dried for 10 h at 95 °C and calcined in a temperature-controlled tubular furnace at a certain temperature for 60 min under the protection of nitrogen atmosphere. The obtained samples were denoted as CuFex/WSWUy(z). Where ultrasonically treatment is represented as “U”, the molar ratio of Cu/Fe is represented as “x” (0.1, 0.3 and 0.5), the mass ratio of CuFe/ WSW is represented as “y” (5%, 10% and 15%), and the calcination temperature (°C) is represented as “z” (300 °C, 500 °C and 700 °C), respectively. The samples in the absence of ultrasonic treatment are denoted as CuFex/WSWy(z). 2.2. Sample characterization The crystalline structure of samples was measured via an X-ray diffractometer (XRD, D8 ADVANCE, Bruker, Germany). The surface characteristics of samples were researched via a Hitachi S-4800 instrument from Japan (Scanning Electron Microscopy, SEM). The chemical states of some elements on the samples were determined via the ESCALAB250 spectrometer instrument (X-ray photoelectron spectroscopy, XPS). The pore structures/specific surface areas of all samples were obtained based on a N2 desorption/adsorption method at 77 K using the Micromeritics Tristar II 3020 analyzer from USA (Brunauer Emmett Teller, BET, Micromeritics Instrument Crop.). The magnetism of sample was also measured based on VSM device (Vibrating Sample Magnetometer, SQULD VSM, Quantum). 2.3. Experimental procedures and setup The experimental setup diagram for testing the Hg0 capturing capacity of the samples is depicted in Fig. 1. The testing system roughly consists of the following parts: a flue gas generation system, a flue gas and Hg0 online detection device, a fixed bed reactor and an exhaust gas purification device. The flue gas generation system roughly consists of a water steam generator, a Hg0 vapor generator, a gas mixer, some gas cylinders. The detection system roughly involves a Hg0 online analyzer (QM201H, China) and a flue gas analyzer (MRU-VARIOPLUS, Germany). The length and inner diameter of the fixed bed reactor were 4.0 cm and 2.3 cm, respectively. The exhaust gas purification device was an activated carbon trap that can effectively capture the residual pollutants from the tail gas. The simulated flue gas was mainly composed of 5% O2, 400 ppm NO, 600 ppm SO2, 4% (vol) H2O, 65 μg/m3 Hg0 and N2 (as the balanced gas). The water vapor and Hg0 vapor were provided based on a water steam generator and a Hg0 vapor generator, respectively. For a test, the mixture of 35 mg sample and 0.5 g quartz sand was used and the total flow rate of flue gas passing through the reactor was kept at 0.8 L/min, which corresponded to the GHSV (hourly space velocity of gas) of about 57000 h−1. The concentration of Hg0 and each component (especially SO2, NO, O2) in flue gas at the inlet/outlet of the reactor was measured by an online Hg0 analyzer and a flue gas analyzer, respectively. All lines of the experimental system were warmed to 105 °C via a heating tape to prevent the condensation of Hg0 vapor and water vapor. The exhaust gas generated in each experiment was introduced into a treatment system for efficient purification to avoid pollution. Before the mercury removal experiment, a blank test was first carried out in the fixed bed reactor to evaluate the stability of the Hg0 concentration in the simulated flue gas. The efficiency of Hg0 removal can be calculated using the following equation:
2. Experimental 2.1. Preparation of samples Wheat straw (from Yancheng of Jiangsu province, China) was washed, dried, pulverized and sieved to the size of less than 300 μm (< 50 mesh). The results of proximate and ultimate analysis of biomass materials (wheat straw) are listed in Table 1. The raw materials were placed in a temperature-controlled tubular furnace at 600 °C and pyrolyzed for 20 min under the protection of nitrogen (N2) (200 mL/min). The obtained biochar derived from wheat straw was noted as WS. 1.0 g of the wheat straw char (WS) was put into a quartz glass tube with the protection of nitrogen (400 mL/min). The quartz glass tube was placed vertically in a microwave stove (700 W, Glanz Equipment Co., Ltd., China). The water steam (2 mL/h) generated by a water vapor generator was introduced into the quartz glass tube for 1 h. The biochar samples activated by water steam and microwave were denoted as WSW. The activated samples were further modified via an ultrasonic-assisted impregnation method, which is depicted as follows: Firstly, suitable amounts of iron nitrate (Fe(NO3)3) (Sinopharm Chemical Reagent Co., Ltd., China (Purity, > 98.5%)) and copper nitrate (Cu(NO3)2) (Xilong Science Co., Ltd., China (Purity, > 99.0%)) were dissolved in deionized water. Secondly, a certain amount of the activated samples was immersed into the above solution for 60 min with continuous stirring at 30 °C. The above mixture was further ultrasonically treated in Table 1 The proximate/ultimate analysis of wheat straw. Proximate analysis (wt%)
=
Ultimate analysis (wt%)
Var
Mar
Aar
FCar
Car
Nar
Oar
Har
Sar
65.73
5.82
4.47
23.98
40.51
0.92
42.74
5.46
0.08
Hg in
Hg out
Hg in
× 100%
(1)
where η (%) was defined as the removal efficiency of Hg0 for samples. Hgin (μg/m3) and Hgout (μg/m3) were defined as concentration of Hg0 before and after adsorption of the samples, respectively. 2
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Fig. 1. The diagram of experimental apparatus for capturing Hg0.
3. Results and discussion
3.1.2. SEM analysis The SEM micrographs of raw biochar (WS), activated sample (WSW) and modified samples are depicted in Fig. 2. It can be seen from Fig. 2 that raw biochar (WS) exhibits a relatively poor pore structure, while the sample WSW activated by steam and microwave has well-developed pore structure. Compared with WSW, the surface of the modified samples becomes rougher and the pore structure has been blocked to some extent. For the sample without ultrasonic treatment (CuFe0.3/ WSW10(500)), a large amount of particulate matter aggregated on the surface. However, the ultrasonic treated sample (CuFe0.3/ WSWU10(500)) shows a better pore structure because agglomerates are cleared by ultrasonic processing. This indicates that ultrasonic-assisted impregnation promotes the dispersion of Cu-Fe mixed oxides on the surface of samples. The analysis of pore structure (BET analysis) in Table 2 supports these results.
3.1. Characterization of materials 3.1.1. BET analysis The physical properties of raw samples and modified samples including pore volume, BET surface area and average pore size are summarized in Table 2. The sample (WSW) activated by microwave and steam shows the highest pore volume (0.291 cm3/g) and BET surface area (286.494 m2/g), while the raw material WS shows a relative poor pore volume (0.217 cm3/g) and BET surface area (154.920 m2/g). The result indicates that activation of microwave and steam has a great influence on the structural properties of materials [28]. Compared with the activated sample WSW, the introduction of copper-iron mixed oxide reduces the specific surface areas and pore volumes of the modified materials in different extents. This might be resulted from the blockage that was caused by the introduction of copper-iron mixed oxides into the partial micropore of the activated materials [29]. In addition, the pore volume and specific surface area of the ultrasonic-treated sample (CuFe0.3/WSWU10(500)) are larger than that of sample without ultrasonic treatment (CuFe0.3/WSW10(500)), which indicates that ultrasonic treatment can contribute to the rise of the pore volumes and BET surface areas of the samples.
3.1.3. XRD analysis The XRD patterns of raw WSW and modified samples are shown in Fig. 3. As shown in Fig. 3(a), for the ultrasonic treated sample (CuFe0.3/WSWU10(500)), only two weak peaks of iron oxides are observed, while the peaks from copper oxides are not detected. The result demonstrates that copper oxides and most iron oxides are highly dispersed on the surface of sample (WSW) in the form of poorly crystalline state or amorphous [30]. Moreover, some characteristic peaks, corresponding to iron oxides and copper oxides, are detected on the CuFe0.3/WSW10(500) sample. These results suggest that ultrasonic treatment is beneficial to the dispersion of metal oxides over the surface of samples. The XRD pattern of samples with different molar ratios of Cu/Fe is exhibited in Fig. 3(b). As exhibited in Fig. 3(b), the characteristic peaks of iron oxides and copper oxides are clearly observed on the samples with Cu/Fe molar ratios of 0.1 and 0.5, respectively. For the sample with Cu/Fe molar ratio of 0.3 (CuFe0.3/WSWU10(500)), only two diffraction peaks for iron oxides are detected. Moreover, no diffraction peaks for copper oxides are detected on the surface of the sample CuFe0.3/WSWU10(500). The results show that when the molar ratio of Cu/Fe is 0.3, the iron-copper mixed oxide exhibits good interaction and
Table 2 The physical properties of raw and modified materials. Samples
BET area (m2/g)
Pore volume (cm3/g)
Average pore size (nm)
WS WSW CuFe0.3/WSWU5(500) CuFe0.3/WSWU10(500) CuFe0.3/WSWU15(500) CuFe0.1/WSWU10(500) CuFe0.5/WSWU10(500) CuFe0.3/WSW10(500)
154.920 286.494 272.444 253.492 247.027 191.388 252.850 223.312
0.217 0.291 0.274 0.259 0.253 0.209 0.297 0.250
5.608 4.060 4.038 4.409 4.133 4.377 4.603 4.417
3
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Fig. 2. SEM micrographs for samples: (a) WS, (b) WSW, (c) CuFe0.3/WSW10(500), (d) CuFe0.3/WSWU10(500).
important impact on the removal of Hg0. According to the results of BET analysis in Table 2, ultrasonic treatment increases the BET surface area and pore volume of the modified samples, which is conducive to the removal of Hg0. In addition, SEM analysis in Fig. 2 and XRD analysis in Fig. 3(a) reveal that ultrasonic treatment improves the dispersion of Cu-Fe mixed oxides on the surface of samples and also avoids the occurrence of aggregation, thus promoting the removal of Hg0. Some related studies [13,32] also suggest that the ultrasound treatment promotes the Hg0 removal from flue gas, which is consistent with the present research results.
excellent dispersion on the surface of sample, which is conducive to the capture of Hg0 [31]. Fig. 3(c) exhibits the structure and phase composition of samples at different calcining temperatures. For the sample calcined at 300 °C, the peaks for iron oxides and Cu2O are observed on the surface of the sample. Compared with the sample calcined at 300 °C, only two peaks of iron oxides are observed on the CuFe0.3/WSWU10(500) sample (calcined at 500 °C), while the peaks of copper oxides are not detected. Moreover, with the rise of calcining temperature to 700 °C, some diffraction peaks of Fe3C are also detected. This results indicate that calcination temperature plays an important role in the generation of metal oxide components and crystalline phases on the surface of samples.
3.2.2. Effect of Cu/Fe molar ratio on Hg0 removal Fig. 5(b) illustrates the influence of different Cu/Fe molar ratio on Hg0 removal performance over samples. As depicted in Fig. 5(b), with the Cu/Fe molar ratio changing from 0.1 to 0.3, the average efficiency of Hg0 removal rises from 79.61% to 90.58%. However, when the Cu/ Fe molar ratio further changes from 0.3 to 0.5, the average Hg0 removal efficiency drops from 90.58% to 86.26%. According to the results of XRD analysis in Fig. 3(b), for samples with Cu/Fe molar ratios of 0.1 and 0.5, some diffraction peaks of iron and copper oxides are clearly detected. However, for the sample with Cu/Fe molar ratio of 0.3, no characteristic peaks of copper oxides are observed. So when the molar ratio of Cu/Fe is 0.3, the copper-iron mixed oxide shows a good dispersion and a strong interaction [33]. Related studies [33,34] indicate that the interaction between iron oxides and copper oxides produces more active sites on the surface of samples, thus promoting the removal of Hg0. Therefore, based on the above experimental and analytical results, 0.3 is selected as the optimal Cu/Fe molar ratio for the further research.
3.1.4. Magnetism analysis The magnetism analysis of the sample is displayed in Fig. 4. It is seen from the Fig. 4 that the CuFe0.3/WSWU10(500) sample exhibits good magnetization characteristics. Good magnetization characteristics can effectively avoid the permanent magnetization of samples in an external magnetic field. Therefore, when the external magnetic field disappears, the sample particles will exhibit a non-aggregated dispersion state. Thus, the spent sample after Hg0 capture can be easily collected from the fly ash. 3.2. Effects of operating parameters on Hg0 removal 3.2.1. Effect of ultrasonic treatment on Hg0 removal The influence of ultrasonic treatment on Hg0 removal performance of samples is researched, and the results are depicted in Fig. 5(a). As seen in Fig. 5(a), the average Hg0 removal efficiencies of WS, WSW, CuFe0.3/WSW10(500) and CuFe0.3/WSWU10(500) are 13.55%, 58.89%, 84.78% and 90.58%, respectively. The Hg0 removal efficiency of ultrasonic treating sample (CuFe0.3/WSWU10(500)) is significantly higher than that of sample without ultrasonic treatment (CuFe0.3/ WSW10(500)). The result indicates that the ultrasonic treatment has an
3.2.3. Effect of calcination temperature on Hg0 removal The influence of calcination temperature on Hg0 removal performance of modified samples is investigated, and the results are illustrated in Fig. 5(c). As observed in Fig. 5(c), the calcination temperature 4
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Fig. 3. The XRD pattern of raw and modified samples.
[32,35] suggest that Cu2O exhibits a lower catalytic oxidation activity compared to CuO, thus inhibiting the removal of Hg0. Moreover, some characteristic peaks, corresponding to Fe3C, are also detected at the calcination temperature of 700 °C. The result [14] indicates that a portion of iron oxides can react with amorphous carbon to form Fe3C, which reduces the concentration of active components on the surface of modified sample, thereby being able to inhibit the removal of Hg0. Therefore, 500 °C is selected as the optimal calcination temperature for the further research. 3.2.4. Effect of loading value on Hg0 removal Fig. 5(d) illustrates the influence of loading value on Hg0 removal performance over modified samples. It can be seen that, compared with the WSW sample, the introduction of Cu-Fe mixed oxides significantly improves the Hg0 removal of modified samples. As the loading value advances from 0 to 5% at first and then further advances from 5 to 10%, the average Hg0 removing efficiency rises from 58.89% to 79.21% at first and then further rises from 79.21% to 90.58%. Related results [36,37] suggest that the content of active components on the surface of modified sample will be increased through increasing the loading value (Cu-Fe mixed oxides), thereby being able to enhance the Hg0 removal. For all that, the sample with the loading value of 15% has more Cu-Fe mixed oxides than the other modified samples, its Hg0 removing efficiency is only 88.01%. This indicates that excessive Cu-Fe mixed oxide is not conducive to Hg0 removal. According to the BET analysis in Table 2, with increasing loading value, the pore volumes and BET surface areas of sample significantly decrease. Relevant results [32,35,38] indicate that excessive loading value will lead to the blockage of the partial pore structure, which is not
Fig. 4. The magnetism analysis of the CuFe0.3/WSWU10(500) sample.
has a great influence on Hg0 removal performance of modified samples. It can be seen that, as the calcination temperature increases from 300 °C to 500 °C, the average Hg0 removal efficiency of modified sample greatly increases from 81.49% to 90.58%. But the Hg0 removal efficiency decreases from 90.58% to 87.19% when the calcination temperature further rises from 500 °C to 700 °C. This result illustrates that the modified sample exhibits the highest Hg0 removal efficiency when the calcination temperature is 500 °C. According to the results of XRD analysis in Fig. 3(c), two diffraction peaks, belonging to Cu2O, are observed at the calcination temperature of 300 °C. Related results 5
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Fig. 5. The effects of (a) ultrasonic treatment, (b) Cu/Fe molar ratio, (c) calcination temperature, (d) loading value, (e) reaction temperature on removal efficiency of Hg0. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%; H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3 and reaction temperature, 130 °C.
conducive to the physisorption of Hg0. The gaseous Hg0 is first captured on the surface of samples by physisorption to form adsorbed Hg0 in the process of Hg0 removal. And then the adsorbed Hg0 will be removed by oxidation or chemisorption on the active sites of samples. Thus, the reduction of physisorption performance will further inhibit the oxidation or chemisorption process [14]. Therefore, 10% is chosen as an optimized loading value for further research.
3.2.5. Effect of reaction temperature on Hg0 removal The effect of reaction temperature on Hg0 removal performance of the optimal sample (CuFe0.3/WSWU10(500)) is investigated, and the results are displayed in Fig. 5(e). It can be seen that the CuFe0.3/ WSWU10(500) sample shows a lower Hg0 removal efficiency at 70 °C, and only about 77.23% of Hg0 is removed from flue gas. As the reacting temperature advances from 100 °C to 130 °C, the average Hg0 removal efficiency of sample rises from 81.88% to 90.58%. The results reveal that higher temperature promotes the removal of Hg0. Related research 6
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Fig. 6. The effects of flue gas components (O2, NO, SO2, H2O) on Hg0 removal efficiency. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%; H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3; Sample: CuFe0.3/WSWU10(500); Reaction temperature, 130 °C.
[35,39,40] suggests that more energy can be obtained through increasing temperature, thus accelerating the chemical reaction rate of Hg0 removal. However, when the temperature increases from 130 °C to 160 °C, the average efficiency of Hg0 removal drops from 90.58% to 86.60%. Some relevant results [14,17,41] indicate that excessive temperature will weaken the physisorption performance, and even will cause the desorption of adsorbed Hg0 on the sample, thereby hindering the removal of Hg0 from flue gas. Therefore, 130 °C is identified as the optimal temperature for Hg0 removing reactions in this work.
3.3.2. Effect of NO concentration Fig. 6(b) shows the influences of NO concentration on Hg0 removing performance. It is found from Fig. 6(b) that NO has a positive influence on the Hg0 removing performance of the CuFe0.3/WSWU10(500) sample. Without NO, the average Hg0 removing efficiency of the sample is 85.28%. As the NO concentrations change from 200 ppm to 800 ppm, the average efficiency of Hg0 removal over CuFe0.3/WSWU10(500) sample rises from 88.77% to 93.96%. This shows that the Hg0 removal capacity of sample has been enhanced through increasing the concentration of NO. The results of some researchers [41,44,45] suggested that NO could be oxidized by chemisorbed oxygen or/and lattice oxygen on the surface of catalyzer to produce some new active species, such as NO+ and NO2. Thus, when these newly formed active species are generated during the Hg0 removal, and they can react with adsorbed Hg0 to produce HgO and Hg(NO3)2, thereby being able to promote the removal of Hg0.
3.3. Effects of flue gas components on Hg0 removal 3.3.1. Effect of O2 concentration Fig. 6(a) exhibits the impact of O2 concentration on Hg0 removing performance. As exhibited in Fig. 6(a), the average Hg0 removing efficiency of the CuFe0.3/WSWU10(500) sample is 81.34% in the absence of O2. When 2% O2 is introduced into flue gas, the average efficiency of Hg0 removal rises from 81.34% to 87.08%, and further rises to 93.31% in the presence of 10% O2. Related data proves that the addition of O2 can promote the removal of Hg0. Results of the previous researchers [19,42,43] had suggested that the consumed chemisorbed oxygen and lattice oxygen would be regenerated and replenished by introducing the gaseous oxygen into flue gas during the process of Hg0 removal. Thus, a rise in the concentration of gaseous oxygen is beneficial to enhance the Hg0 removal capacity of samples.
3.3.3. Effect of SO2 concentration Fig. 6(c) displays the impact of SO2 concentration on performance of Hg0 removal. As seen in Fig. 6(c), the introduction of SO2 is not conductive to the removal of Hg0. When SO2 is absent, the CuFe0.3/ WSWU10(500) sample exhibits the highest Hg0 removal efficiency, and about 95.98% of Hg0 is removed. However, when SO2 is introduced into flue gas, and as the SO2 concentration increases from 0 to 1800 ppm, the average Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample decreases greatly from 95.98% to 81.85%. The inhibitive effect of SO2 concentration on Hg0 removal efficiency may be caused by the 7
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following reasons. On the one hand, the competitive adsorption between Hg0 and SO2 reduces the content of active sites for removing the gaseous Hg0 [44,46]. On the other hand, some metal sulfates produced by the reaction of metal oxides and SO2 can cause the deactivation and poisoning of the sample [29,47]. Thus, a rise in the concentration of SO2 reduces the removal efficiency of Hg0.
by the reaction Eq. (2) [19].
Hg0 (g)
(2)
Hg0 (ad)
Secondly, the generation process of lattice oxygen ([O]) on the sample can be expressed by the following reaction Eqs. (3)–(5) [31,32,52].
3.3.4. Effect of water vapor concentration Fig. 6(d) displays the effect of water vapor concentration on Hg0 removing performance. Compared with the flue gas without water vapor, the addition of 4% water vapor can promote the removal of Hg0. Relevant research [48] indicates that the hydroxyl radicals derived from water molecule decomposition can react with the adsorbed Hg0 to generate HgO, thus promoting the removal of Hg0. But when 12% water vapor is added into flue gas, the removal efficiency of Hg0 sharply drops from 90.58% to 71.58%. Firstly, the competition between water vapor and Hg0 for the same active sites is not conducive to removing the Hg0 [19,44]. Secondly, the water vapor can also combine with Cu-Fe mixed oxides and SO3 to generate sulfates, which easily causes the deactivation of the sample, thereby inhibiting the removal of Hg0 [29,41]. In addition, the water film, formed on the surface of the sorbent, also inhibits the removal of Hg0 [49].
2CuO
Cu2 O+ [O]
(3)
Fe2 O3
2FeO + [O]
(4)
Fe3 O4
3FeO + [O]
(5) 0
Thirdly, the adsorbed Hg on the surface of sample can combine with lattice oxygen to generate HgO. In addition, the adsorbed Hg0 can also be oxidized by chemisorbed oxygen to produce HgO. The specific reaction steps can be expressed by the following reaction Eqs. (6)–(8) [32,52].
Hg0 (ad) + [O]
HgO(ad)
(6)
Hg0 (ad) + O*
HgO(ad)
(7)
HgO(ad)
(8)
HgO(g)
Finally, the chemisorbed oxygen (O*) and lattice oxygen ([O]), which are consumed in the process of Hg0 removal, will be efficiently replenished and regenerated by introducing gaseous oxygen into the flue gas. The reaction process can be described by the following Eqs. (9)–(12) [19,31,52].
0
3.4. Reaction mechanism of Hg removal To investigate the reaction mechanism of Hg0 removal, the chemical states of some elements on the fresh and spent samples were tested by XPS. Fig. 7(a) and (b) present the Hg 4f XPS analysis spectra of the fresh and spent samples. As seen in Fig. 7(a), only a strong peak appears at 103.1 eV on the fresh sample, which usually corresponds to Si 2p [19]. For the spent sample, in addition to the peak belonging to Si 2p, a new spectral peak at 104.1 eV is observed, which is assigned to Hg2+ [19]. The results suggest that Hg0 adsorbed on the surface of sample is existed in the form of divalent mercury. Fig. 7(c) and (d) show the O 1s XPS analysis spectra of the fresh and spent samples. It can be seen that there are three obvious overlapping peaks for the O 1s in the fresh and spent samples. The peak at 530.4/ 530.6 eV belongs to the lattice oxygen (denoted as [O]) [14,49]. The peak value of 531.5/531.6 eV corresponds to the chemisorbed oxygen (denoted as O*) [5,41]. Moreover, the peak at 533.3/532.7 eV has been attributed to C-O [4,14]. After the Hg0 removal tests, the content of lattice oxygen increases from 19.66% to 25.01%, while the content of chemisorbed oxygen decreases from 35.19% to 26.68%. These data indicate that chemisorbed oxygen plays an important role in the removal of Hg0. The spectral peaks of Fe 2p in the fresh and spent samples are shown in Fig. 7(e) and (f). Two main spectrum peaks corresponding to Fe 2p3/2 and Fe 2p1/2 are detected at 711.6 eV and 725.1 eV, respectively [50]. The Fe 2p3/2 peak consists of one Fe2+ peak (710.2 eV) and two Fe3+ peaks (711.3 eV and 713.1 eV) [14,19]. The Fe 2p1/2 peak at 725.1 eV is also attributed to Fe3+ [50]. After the Hg0 removal experiments, the content of Fe2+ increases from 2.70% to 17.64%, while the content of Fe3+ decreases from 97.30% to 82.36%. These results suggest that the transition from Fe3+ to Fe2+ occurs in the process of Hg0 removal. The Cu 2p XPS spectra for the fresh and spent samples are shown in Fig. 7(g) and (h). It can be seen that the Cu 2p1/2 band at 954.0–956.0 eV, the higher band of Cu 2p3/2 at 933.0–934.5 eV, and the shake-up peak at 939.0–942.0 eV are attributed to Cu2+ [25,51,52]. The lower band of Cu 2p3/2 at 932.2 eV belongs to Cu+ [26,32,53]. As shown in Fig. 7(g) and (h), the content of Cu+ in the spent sample increases from 8.98% (for the fresh sample) to 21.29%. Conversely, the content of Cu2+ in the spent sample decreases from 91.02% (for the fresh sample) to 78.71%. The results indicate that Cu2+ participates in the removal process of Hg0. Based on the above analysis and experiment results, the Hg0 removal mechanism of modified samples is proposed preliminarily. Firstly, the gas-phase Hg0 in flue gas is trapped on the surface of sample to generate the adsorbed Hg0. The formation process can be described
O2
(9)
2O*
Cu2 O+ 1/2O2
2CuO
(10)
2FeO + 1/2O2
Fe2 O3
(11)
3FeO + 1/2O2
Fe3 O4
(12)
3.5. Adsorption kinetics In this work, two adsorption kinetic models (the pseudo-first model and the pseudo-second model) are applied to study the removal mechanism and adsorption kinetic process of modified porous chars. The pseudo-first model is mainly used to describe the process of external mass transfer. The pseudo-second model can be used to describe the process of chemisorption. The pseudo-first model can be expressed by the following equation [4]:
lg(qe
qt ) = lg qe
k1 t 2.303
(13)
The pseudo-second model can be expressed by the following equation [54]:
t 1 1 = + t qt qe k2 qe2
(14)
where t (min) represents the reaction time. qe and qt (μg/g) represent the uptake amount of samples at time t and equilibrium time, respectively. k1 (1/min) and k2 (g/(μg·min)) represent the contents of the pseudo-first model and the pseudo-second model, respectively. The fitting curves and fitting parameters are shown in Fig. 8 and Table 3, respectively. As shown in Fig. 8(a) and Table 3, the correlation coefficients (R2), obtained by the pseudo-first model, are relatively low (< 0.96). This result indicates that the external mass transfer is not the rate-controlling step for the Hg0 removal process. Thus, the pseudo-first model cannot be used to describe the Hg0 removal process of modified porous chars [4]. It can be seen from Fig. 8(b) and Table 3 that the the pseudo-second model can fit the experimental data very well. In Table 3, compared with the pseudo-first model, the correlation coefficients (R2) of the pseudo-second model are higher than 0.96. These 8
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Fig. 7. The XPS analysis of the fresh and spent samples over the spectral regions of Hg 4f, O 1s, Fe 2p and Cu 2p.
results show that the pseudo-second model can be used to express the Hg0 adsorption process of modified samples. Therefore, the chemisorption process is the key rate-controlling step for Hg0 removal [54,55,56].
3.6. Comparison and regeneration of the modified porous char Table S1 listed the comparison of Hg0 adsorption capacity between the modified porous char and other activated carbon sorbents/catalysts. 9
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Fig. 8. Kinetic analysis of modified porous chars: (a) the pseudo-first model, (b) the pseudo-second model.
As shown in Table S1, compared with activated carbon sorbents/catalysts, the modified porous char (CuFe0.3/WSWU10(500)) exhibits the best Hg0 removal performance, and its Hg0 adsorption capacity is 2276.45 μg/g. In addition, the modified porous char in this paper mainly comes from cheap agricultural wastes, which is beneficial to reduce the cost of Hg0 removal. Therefore, the modified porous char (CuFe0.3/WSWU10(500)) has good development and application prospects due to its high Hg0 adsorption capacity and low cost. In addition to high Hg0 adsorption capacity, the regeneration performance of modified porous char after Hg0 removal is also very important for its future application and development. Therefore, in this work, the spent CuFe0.3/WSWU10(500) sample was first heated at 400 °C for 60 min under the protection of nitrogen, and then regenerated at 250 °C for 30 min under air atmosphere. The Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample in six cycles is shown in Fig. 9. It can be seen from Fig. 9 that the sample still has a high Hg0 removal efficiency after six regeneration cycles. After the sixth cycle regeneration, the Hg0 removal performance of the sample is basically stable, and the Hg0 removal efficiency is about 76.33%. These results suggest that the CuFe0.3/WSWU10(500) sample exhibits good regeneration characteristics.
Fig. 9. The Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample in six cycles. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%, H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3 and reaction temperature, 130 °C.
4. Conclusions and about 90.58% of Hg0 is removed from flue gas at 130 °C. Increasing the concentration of O2 and NO can clearly promote the removal of Hg0. Increasing SO2 concentration inhibits Hg0 removal. Low concentration of water vapor improves Hg0 removal performance, while high concentration of water vapor plays an opposite role. The copper/ iron active components, chemisorbed oxygen and lattice oxygen play an important role in the process of Hg0 removal. The modified sample exhibits good regeneration performance.
In this article, a novel magnetic iron-copper oxide modified biomass-based porous char was synthesized by microwave activation and ultrasonic-assisted impregnation method and was used to control Hg0 emissions from flue gas. The effects of operating parameters and flue gas components on Hg0 removal were investigated. The reaction mechanism and the regeneration performance of samples were also studied. The results indicate that the ultrasound treatment promotes the Hg0 removal from flue gas. The microwave and steam activation greatly improves the pore structure of porous char. The CuFe0.3/ WSWU10(500) sample exhibits an excellent Hg0 removal performance, Table 3 Kinetic parameters of modified porous chars. Samples
CuFe0.3/WSWU5(500) CuFe0.3/WSWU10(500) CuFe0.3/WSWU15(500)
Pseudo-first model
Pseudo-second model 2
qe (μg/g)
k1 (1/min)
R
425.48 928.47 604.48
0.0037 0.0015 0.0019
0.9577 0.9139 0.9197
10
qe (μg/g)
k2 (g/(μg·min))
R2
1027.81 2276.45 1592.17
1.2384 × 10−6 0.2729 × 10−6 0.5559 × 10−6
0.9786 0.9962 0.9646
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Acknowledgements
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11
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Full Length Article
Mercury removal from flue gas by magnetic iron-copper oxide modified porous char derived from biomass materials
T
Wei Yang, Ying Li, Shuo Shi, Hui Chen, Ye Shan, Yangxian Liu
⁎
School of Energy and Power Engineering, Jiangsu University, Zhenjiang, Jiangsu 212013, China
ARTICLE INFO
ABSTRACT
Keywords: Elemental mercury Flue gas Iron-copper mixed oxides Biomass porous char Magnetic catalyst
In this work, a magnetic iron-copper oxide modified porous char derived from biomass materials was prepared by microwave activation and ultrasonic-assisted impregnation method to remove Hg0 from flue gas. The effects of operating parameters and flue gas components on Hg0 removal performance as well as physicochemical properties of samples were studied. The reaction mechanism and the regeneration performance of samples were also investigated. The experimental results display that the biomass-based porous char derived from microwave and steam activation exhibits a larger pore volume, higher surface area and well-developed pore structure. The ultrasonic treatment promotes the dispersion of active components on the surface of samples. The CuFe0.3/ WSWU10(500) sample exhibits good Hg0 removal performance, and about 90.58% of Hg0 is captured from flue gas at 130 °C. Increasing SO2 concentration inhibits Hg0 removal. The presence of O2 and NO promotes the removal of Hg0. Low concentration of water vapor (> 4%) improves Hg0 removal performance, while excessive water vapor plays an opposite role. The presence of copper/iron active components, chemisorbed oxygen and lattice oxygen over the surface of adsorbent is conducive to the removal of Hg0. The modified sample exhibits a good regeneration performance.
1. Introduction Mercury has received more and more global attention due to its bioaccumulation, mobility and toxicity in both food chains and ecosystems [1]. Coal-fired utility boilers release large amounts of mercury into the atmosphere each year, which is considered to be the largest source of anthropogenic mercury emissions. In the United States, mercury emissions from coal-fired utility boilers account for approximately one-third of the 150 tons of mercury emissions [2]. Coal combustion contributes about 38% of anthropogenic mercury emissions in China [3]. Mercury in coal-fired flue gas is mainly composed of gaseous mercury species and particulate-bound mercury (Hgp). Gaseous mercury species are mainly involved in oxidized mercury (Hg2+) and elemental mercury (Hg0) [4]. The existing pollution control devices such as wet flue gas desulfurization device (WFGD) and particulate control device can be used to capture oxidized mercury and particulate-bound mercury, respectively. However, due to its low solubility in water and high volatility, the elemental mercury (Hg0) in flue gas is difficult to be effectively captured by the above-mentioned existing pollution control devices [5]. Therefore, the main focus of mercury emission control is the effective removal of elemental mercury (Hg0) from flue gas. Up to date, a number of mercury control methods have been developed for ⁎
the capture of elemental mercury from coal-fired flue gas, such as catalytic oxidation, sorbent adsorption, chemical oxidation, photochemical oxidation, photocatalytic oxidation and so on [6–11]. The activated carbon (AC) injection method, which is performed in the upstream of the particulate control device, is recognized as one of the most effective techniques for elemental mercury removal. However, higher removal costs and C/Hg0 ratios limit the widespread application of this method [12]. Moreover, the development of mercury removal technologies is also extremely limited by the separation and recovery of spent sorbents/catalysts from fly ash. Biochar, a byproduct from pyrolysis of biomass under oxygen-limited environment, has been used to remove elemental mercury from flue gas [13]. With lower cost, it could be a promising substitute for activated carbon in the field of mercury removal [14]. However, lower mercury removal performance and undeveloped pore structure limit the application of biomass char. Thus, some chemical or physical modification methods are applied to improve the mercury capture performance. The methods mainly involve the introduction of active sites on the surface of biochar, such as halides, sulfur, metal oxides, etc [15–17]. However, it is still a huge challenge to separate the spent modified biochar from fly ash. Therefore, it is necessary to develop some easily separable biochars to control mercury emissions.
Corresponding author. E-mail address: [email protected] (Y. Liu).
https://doi.org/10.1016/j.fuel.2019.115977 Received 16 May 2019; Received in revised form 22 July 2019; Accepted 6 August 2019 Available online 13 August 2019 0016-2361/ © 2019 Elsevier Ltd. All rights reserved.
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In order to separate the spent sorbents from fly ash, the addition of the magnetic materials such as γ-Fe2O3 and Fe3O4 to sorbents may be an efficient way [18,19]. Some related results [19–24] suggest that the γFe2O3 and Fe3O4 can capture Hg0 from flue gas to a certain extent, but its Hg0 removal capacity is relatively low. Many studies have shown that copper oxide (CuOx) is a promising Hg0 capture material due to its higher stability and catalytic activity for Hg0 removal at low temperature [25]. Du et al. [26] prepared a CuOx-based sample and found that it exhibited a good performance for Hg0 removal. In addition, Yang et al. [27] found that CuO can effectively improve the Hg0 removal capacity. Thus, the incorporation of copper oxides into Fe-based sorbents may contribute to the improvement of mercury removal efficiency. Related research shows that the activation of microwave and steam can greatly improve the pore structure of biochars [15]. Moreover, ultrasonic treatment is beneficial to the dispersion of active ingredients on the surface of samples [13]. Therefore, in this work, a novel magnetic iron-copper oxide modified biomass-based porous char derived from the activation of microwave and steam was prepared by using an ultrasonic- assisted impregnation method, and was used to control Hg0 emissions from flue gas. The physicochemical properties of the samples were investigated by various characterization methods such as XPS, BET, SEM, XRD and magnetism analysis. The effects of Cu/Fe molar ratios, reaction temperatures, loading values, calcination temperatures on Hg0 capture efficiency were analyzed, as well as the effects of flue gas constituents, microwave activation and ultrasonic-assisted impregnation. Furthermore, both the reaction mechanism and the regeneration performance of the samples for Hg0 removal were also researched. The results of this research will provide some references for the development of this new technology in the field of Hg0 emission control.
a water bath at 40 °C for 30 min. Then, the mixture was dried for 10 h at 95 °C and calcined in a temperature-controlled tubular furnace at a certain temperature for 60 min under the protection of nitrogen atmosphere. The obtained samples were denoted as CuFex/WSWUy(z). Where ultrasonically treatment is represented as “U”, the molar ratio of Cu/Fe is represented as “x” (0.1, 0.3 and 0.5), the mass ratio of CuFe/ WSW is represented as “y” (5%, 10% and 15%), and the calcination temperature (°C) is represented as “z” (300 °C, 500 °C and 700 °C), respectively. The samples in the absence of ultrasonic treatment are denoted as CuFex/WSWy(z). 2.2. Sample characterization The crystalline structure of samples was measured via an X-ray diffractometer (XRD, D8 ADVANCE, Bruker, Germany). The surface characteristics of samples were researched via a Hitachi S-4800 instrument from Japan (Scanning Electron Microscopy, SEM). The chemical states of some elements on the samples were determined via the ESCALAB250 spectrometer instrument (X-ray photoelectron spectroscopy, XPS). The pore structures/specific surface areas of all samples were obtained based on a N2 desorption/adsorption method at 77 K using the Micromeritics Tristar II 3020 analyzer from USA (Brunauer Emmett Teller, BET, Micromeritics Instrument Crop.). The magnetism of sample was also measured based on VSM device (Vibrating Sample Magnetometer, SQULD VSM, Quantum). 2.3. Experimental procedures and setup The experimental setup diagram for testing the Hg0 capturing capacity of the samples is depicted in Fig. 1. The testing system roughly consists of the following parts: a flue gas generation system, a flue gas and Hg0 online detection device, a fixed bed reactor and an exhaust gas purification device. The flue gas generation system roughly consists of a water steam generator, a Hg0 vapor generator, a gas mixer, some gas cylinders. The detection system roughly involves a Hg0 online analyzer (QM201H, China) and a flue gas analyzer (MRU-VARIOPLUS, Germany). The length and inner diameter of the fixed bed reactor were 4.0 cm and 2.3 cm, respectively. The exhaust gas purification device was an activated carbon trap that can effectively capture the residual pollutants from the tail gas. The simulated flue gas was mainly composed of 5% O2, 400 ppm NO, 600 ppm SO2, 4% (vol) H2O, 65 μg/m3 Hg0 and N2 (as the balanced gas). The water vapor and Hg0 vapor were provided based on a water steam generator and a Hg0 vapor generator, respectively. For a test, the mixture of 35 mg sample and 0.5 g quartz sand was used and the total flow rate of flue gas passing through the reactor was kept at 0.8 L/min, which corresponded to the GHSV (hourly space velocity of gas) of about 57000 h−1. The concentration of Hg0 and each component (especially SO2, NO, O2) in flue gas at the inlet/outlet of the reactor was measured by an online Hg0 analyzer and a flue gas analyzer, respectively. All lines of the experimental system were warmed to 105 °C via a heating tape to prevent the condensation of Hg0 vapor and water vapor. The exhaust gas generated in each experiment was introduced into a treatment system for efficient purification to avoid pollution. Before the mercury removal experiment, a blank test was first carried out in the fixed bed reactor to evaluate the stability of the Hg0 concentration in the simulated flue gas. The efficiency of Hg0 removal can be calculated using the following equation:
2. Experimental 2.1. Preparation of samples Wheat straw (from Yancheng of Jiangsu province, China) was washed, dried, pulverized and sieved to the size of less than 300 μm (< 50 mesh). The results of proximate and ultimate analysis of biomass materials (wheat straw) are listed in Table 1. The raw materials were placed in a temperature-controlled tubular furnace at 600 °C and pyrolyzed for 20 min under the protection of nitrogen (N2) (200 mL/min). The obtained biochar derived from wheat straw was noted as WS. 1.0 g of the wheat straw char (WS) was put into a quartz glass tube with the protection of nitrogen (400 mL/min). The quartz glass tube was placed vertically in a microwave stove (700 W, Glanz Equipment Co., Ltd., China). The water steam (2 mL/h) generated by a water vapor generator was introduced into the quartz glass tube for 1 h. The biochar samples activated by water steam and microwave were denoted as WSW. The activated samples were further modified via an ultrasonic-assisted impregnation method, which is depicted as follows: Firstly, suitable amounts of iron nitrate (Fe(NO3)3) (Sinopharm Chemical Reagent Co., Ltd., China (Purity, > 98.5%)) and copper nitrate (Cu(NO3)2) (Xilong Science Co., Ltd., China (Purity, > 99.0%)) were dissolved in deionized water. Secondly, a certain amount of the activated samples was immersed into the above solution for 60 min with continuous stirring at 30 °C. The above mixture was further ultrasonically treated in Table 1 The proximate/ultimate analysis of wheat straw. Proximate analysis (wt%)
=
Ultimate analysis (wt%)
Var
Mar
Aar
FCar
Car
Nar
Oar
Har
Sar
65.73
5.82
4.47
23.98
40.51
0.92
42.74
5.46
0.08
Hg in
Hg out
Hg in
× 100%
(1)
where η (%) was defined as the removal efficiency of Hg0 for samples. Hgin (μg/m3) and Hgout (μg/m3) were defined as concentration of Hg0 before and after adsorption of the samples, respectively. 2
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Fig. 1. The diagram of experimental apparatus for capturing Hg0.
3. Results and discussion
3.1.2. SEM analysis The SEM micrographs of raw biochar (WS), activated sample (WSW) and modified samples are depicted in Fig. 2. It can be seen from Fig. 2 that raw biochar (WS) exhibits a relatively poor pore structure, while the sample WSW activated by steam and microwave has well-developed pore structure. Compared with WSW, the surface of the modified samples becomes rougher and the pore structure has been blocked to some extent. For the sample without ultrasonic treatment (CuFe0.3/ WSW10(500)), a large amount of particulate matter aggregated on the surface. However, the ultrasonic treated sample (CuFe0.3/ WSWU10(500)) shows a better pore structure because agglomerates are cleared by ultrasonic processing. This indicates that ultrasonic-assisted impregnation promotes the dispersion of Cu-Fe mixed oxides on the surface of samples. The analysis of pore structure (BET analysis) in Table 2 supports these results.
3.1. Characterization of materials 3.1.1. BET analysis The physical properties of raw samples and modified samples including pore volume, BET surface area and average pore size are summarized in Table 2. The sample (WSW) activated by microwave and steam shows the highest pore volume (0.291 cm3/g) and BET surface area (286.494 m2/g), while the raw material WS shows a relative poor pore volume (0.217 cm3/g) and BET surface area (154.920 m2/g). The result indicates that activation of microwave and steam has a great influence on the structural properties of materials [28]. Compared with the activated sample WSW, the introduction of copper-iron mixed oxide reduces the specific surface areas and pore volumes of the modified materials in different extents. This might be resulted from the blockage that was caused by the introduction of copper-iron mixed oxides into the partial micropore of the activated materials [29]. In addition, the pore volume and specific surface area of the ultrasonic-treated sample (CuFe0.3/WSWU10(500)) are larger than that of sample without ultrasonic treatment (CuFe0.3/WSW10(500)), which indicates that ultrasonic treatment can contribute to the rise of the pore volumes and BET surface areas of the samples.
3.1.3. XRD analysis The XRD patterns of raw WSW and modified samples are shown in Fig. 3. As shown in Fig. 3(a), for the ultrasonic treated sample (CuFe0.3/WSWU10(500)), only two weak peaks of iron oxides are observed, while the peaks from copper oxides are not detected. The result demonstrates that copper oxides and most iron oxides are highly dispersed on the surface of sample (WSW) in the form of poorly crystalline state or amorphous [30]. Moreover, some characteristic peaks, corresponding to iron oxides and copper oxides, are detected on the CuFe0.3/WSW10(500) sample. These results suggest that ultrasonic treatment is beneficial to the dispersion of metal oxides over the surface of samples. The XRD pattern of samples with different molar ratios of Cu/Fe is exhibited in Fig. 3(b). As exhibited in Fig. 3(b), the characteristic peaks of iron oxides and copper oxides are clearly observed on the samples with Cu/Fe molar ratios of 0.1 and 0.5, respectively. For the sample with Cu/Fe molar ratio of 0.3 (CuFe0.3/WSWU10(500)), only two diffraction peaks for iron oxides are detected. Moreover, no diffraction peaks for copper oxides are detected on the surface of the sample CuFe0.3/WSWU10(500). The results show that when the molar ratio of Cu/Fe is 0.3, the iron-copper mixed oxide exhibits good interaction and
Table 2 The physical properties of raw and modified materials. Samples
BET area (m2/g)
Pore volume (cm3/g)
Average pore size (nm)
WS WSW CuFe0.3/WSWU5(500) CuFe0.3/WSWU10(500) CuFe0.3/WSWU15(500) CuFe0.1/WSWU10(500) CuFe0.5/WSWU10(500) CuFe0.3/WSW10(500)
154.920 286.494 272.444 253.492 247.027 191.388 252.850 223.312
0.217 0.291 0.274 0.259 0.253 0.209 0.297 0.250
5.608 4.060 4.038 4.409 4.133 4.377 4.603 4.417
3
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Fig. 2. SEM micrographs for samples: (a) WS, (b) WSW, (c) CuFe0.3/WSW10(500), (d) CuFe0.3/WSWU10(500).
important impact on the removal of Hg0. According to the results of BET analysis in Table 2, ultrasonic treatment increases the BET surface area and pore volume of the modified samples, which is conducive to the removal of Hg0. In addition, SEM analysis in Fig. 2 and XRD analysis in Fig. 3(a) reveal that ultrasonic treatment improves the dispersion of Cu-Fe mixed oxides on the surface of samples and also avoids the occurrence of aggregation, thus promoting the removal of Hg0. Some related studies [13,32] also suggest that the ultrasound treatment promotes the Hg0 removal from flue gas, which is consistent with the present research results.
excellent dispersion on the surface of sample, which is conducive to the capture of Hg0 [31]. Fig. 3(c) exhibits the structure and phase composition of samples at different calcining temperatures. For the sample calcined at 300 °C, the peaks for iron oxides and Cu2O are observed on the surface of the sample. Compared with the sample calcined at 300 °C, only two peaks of iron oxides are observed on the CuFe0.3/WSWU10(500) sample (calcined at 500 °C), while the peaks of copper oxides are not detected. Moreover, with the rise of calcining temperature to 700 °C, some diffraction peaks of Fe3C are also detected. This results indicate that calcination temperature plays an important role in the generation of metal oxide components and crystalline phases on the surface of samples.
3.2.2. Effect of Cu/Fe molar ratio on Hg0 removal Fig. 5(b) illustrates the influence of different Cu/Fe molar ratio on Hg0 removal performance over samples. As depicted in Fig. 5(b), with the Cu/Fe molar ratio changing from 0.1 to 0.3, the average efficiency of Hg0 removal rises from 79.61% to 90.58%. However, when the Cu/ Fe molar ratio further changes from 0.3 to 0.5, the average Hg0 removal efficiency drops from 90.58% to 86.26%. According to the results of XRD analysis in Fig. 3(b), for samples with Cu/Fe molar ratios of 0.1 and 0.5, some diffraction peaks of iron and copper oxides are clearly detected. However, for the sample with Cu/Fe molar ratio of 0.3, no characteristic peaks of copper oxides are observed. So when the molar ratio of Cu/Fe is 0.3, the copper-iron mixed oxide shows a good dispersion and a strong interaction [33]. Related studies [33,34] indicate that the interaction between iron oxides and copper oxides produces more active sites on the surface of samples, thus promoting the removal of Hg0. Therefore, based on the above experimental and analytical results, 0.3 is selected as the optimal Cu/Fe molar ratio for the further research.
3.1.4. Magnetism analysis The magnetism analysis of the sample is displayed in Fig. 4. It is seen from the Fig. 4 that the CuFe0.3/WSWU10(500) sample exhibits good magnetization characteristics. Good magnetization characteristics can effectively avoid the permanent magnetization of samples in an external magnetic field. Therefore, when the external magnetic field disappears, the sample particles will exhibit a non-aggregated dispersion state. Thus, the spent sample after Hg0 capture can be easily collected from the fly ash. 3.2. Effects of operating parameters on Hg0 removal 3.2.1. Effect of ultrasonic treatment on Hg0 removal The influence of ultrasonic treatment on Hg0 removal performance of samples is researched, and the results are depicted in Fig. 5(a). As seen in Fig. 5(a), the average Hg0 removal efficiencies of WS, WSW, CuFe0.3/WSW10(500) and CuFe0.3/WSWU10(500) are 13.55%, 58.89%, 84.78% and 90.58%, respectively. The Hg0 removal efficiency of ultrasonic treating sample (CuFe0.3/WSWU10(500)) is significantly higher than that of sample without ultrasonic treatment (CuFe0.3/ WSW10(500)). The result indicates that the ultrasonic treatment has an
3.2.3. Effect of calcination temperature on Hg0 removal The influence of calcination temperature on Hg0 removal performance of modified samples is investigated, and the results are illustrated in Fig. 5(c). As observed in Fig. 5(c), the calcination temperature 4
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Fig. 3. The XRD pattern of raw and modified samples.
[32,35] suggest that Cu2O exhibits a lower catalytic oxidation activity compared to CuO, thus inhibiting the removal of Hg0. Moreover, some characteristic peaks, corresponding to Fe3C, are also detected at the calcination temperature of 700 °C. The result [14] indicates that a portion of iron oxides can react with amorphous carbon to form Fe3C, which reduces the concentration of active components on the surface of modified sample, thereby being able to inhibit the removal of Hg0. Therefore, 500 °C is selected as the optimal calcination temperature for the further research. 3.2.4. Effect of loading value on Hg0 removal Fig. 5(d) illustrates the influence of loading value on Hg0 removal performance over modified samples. It can be seen that, compared with the WSW sample, the introduction of Cu-Fe mixed oxides significantly improves the Hg0 removal of modified samples. As the loading value advances from 0 to 5% at first and then further advances from 5 to 10%, the average Hg0 removing efficiency rises from 58.89% to 79.21% at first and then further rises from 79.21% to 90.58%. Related results [36,37] suggest that the content of active components on the surface of modified sample will be increased through increasing the loading value (Cu-Fe mixed oxides), thereby being able to enhance the Hg0 removal. For all that, the sample with the loading value of 15% has more Cu-Fe mixed oxides than the other modified samples, its Hg0 removing efficiency is only 88.01%. This indicates that excessive Cu-Fe mixed oxide is not conducive to Hg0 removal. According to the BET analysis in Table 2, with increasing loading value, the pore volumes and BET surface areas of sample significantly decrease. Relevant results [32,35,38] indicate that excessive loading value will lead to the blockage of the partial pore structure, which is not
Fig. 4. The magnetism analysis of the CuFe0.3/WSWU10(500) sample.
has a great influence on Hg0 removal performance of modified samples. It can be seen that, as the calcination temperature increases from 300 °C to 500 °C, the average Hg0 removal efficiency of modified sample greatly increases from 81.49% to 90.58%. But the Hg0 removal efficiency decreases from 90.58% to 87.19% when the calcination temperature further rises from 500 °C to 700 °C. This result illustrates that the modified sample exhibits the highest Hg0 removal efficiency when the calcination temperature is 500 °C. According to the results of XRD analysis in Fig. 3(c), two diffraction peaks, belonging to Cu2O, are observed at the calcination temperature of 300 °C. Related results 5
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Fig. 5. The effects of (a) ultrasonic treatment, (b) Cu/Fe molar ratio, (c) calcination temperature, (d) loading value, (e) reaction temperature on removal efficiency of Hg0. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%; H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3 and reaction temperature, 130 °C.
conducive to the physisorption of Hg0. The gaseous Hg0 is first captured on the surface of samples by physisorption to form adsorbed Hg0 in the process of Hg0 removal. And then the adsorbed Hg0 will be removed by oxidation or chemisorption on the active sites of samples. Thus, the reduction of physisorption performance will further inhibit the oxidation or chemisorption process [14]. Therefore, 10% is chosen as an optimized loading value for further research.
3.2.5. Effect of reaction temperature on Hg0 removal The effect of reaction temperature on Hg0 removal performance of the optimal sample (CuFe0.3/WSWU10(500)) is investigated, and the results are displayed in Fig. 5(e). It can be seen that the CuFe0.3/ WSWU10(500) sample shows a lower Hg0 removal efficiency at 70 °C, and only about 77.23% of Hg0 is removed from flue gas. As the reacting temperature advances from 100 °C to 130 °C, the average Hg0 removal efficiency of sample rises from 81.88% to 90.58%. The results reveal that higher temperature promotes the removal of Hg0. Related research 6
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Fig. 6. The effects of flue gas components (O2, NO, SO2, H2O) on Hg0 removal efficiency. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%; H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3; Sample: CuFe0.3/WSWU10(500); Reaction temperature, 130 °C.
[35,39,40] suggests that more energy can be obtained through increasing temperature, thus accelerating the chemical reaction rate of Hg0 removal. However, when the temperature increases from 130 °C to 160 °C, the average efficiency of Hg0 removal drops from 90.58% to 86.60%. Some relevant results [14,17,41] indicate that excessive temperature will weaken the physisorption performance, and even will cause the desorption of adsorbed Hg0 on the sample, thereby hindering the removal of Hg0 from flue gas. Therefore, 130 °C is identified as the optimal temperature for Hg0 removing reactions in this work.
3.3.2. Effect of NO concentration Fig. 6(b) shows the influences of NO concentration on Hg0 removing performance. It is found from Fig. 6(b) that NO has a positive influence on the Hg0 removing performance of the CuFe0.3/WSWU10(500) sample. Without NO, the average Hg0 removing efficiency of the sample is 85.28%. As the NO concentrations change from 200 ppm to 800 ppm, the average efficiency of Hg0 removal over CuFe0.3/WSWU10(500) sample rises from 88.77% to 93.96%. This shows that the Hg0 removal capacity of sample has been enhanced through increasing the concentration of NO. The results of some researchers [41,44,45] suggested that NO could be oxidized by chemisorbed oxygen or/and lattice oxygen on the surface of catalyzer to produce some new active species, such as NO+ and NO2. Thus, when these newly formed active species are generated during the Hg0 removal, and they can react with adsorbed Hg0 to produce HgO and Hg(NO3)2, thereby being able to promote the removal of Hg0.
3.3. Effects of flue gas components on Hg0 removal 3.3.1. Effect of O2 concentration Fig. 6(a) exhibits the impact of O2 concentration on Hg0 removing performance. As exhibited in Fig. 6(a), the average Hg0 removing efficiency of the CuFe0.3/WSWU10(500) sample is 81.34% in the absence of O2. When 2% O2 is introduced into flue gas, the average efficiency of Hg0 removal rises from 81.34% to 87.08%, and further rises to 93.31% in the presence of 10% O2. Related data proves that the addition of O2 can promote the removal of Hg0. Results of the previous researchers [19,42,43] had suggested that the consumed chemisorbed oxygen and lattice oxygen would be regenerated and replenished by introducing the gaseous oxygen into flue gas during the process of Hg0 removal. Thus, a rise in the concentration of gaseous oxygen is beneficial to enhance the Hg0 removal capacity of samples.
3.3.3. Effect of SO2 concentration Fig. 6(c) displays the impact of SO2 concentration on performance of Hg0 removal. As seen in Fig. 6(c), the introduction of SO2 is not conductive to the removal of Hg0. When SO2 is absent, the CuFe0.3/ WSWU10(500) sample exhibits the highest Hg0 removal efficiency, and about 95.98% of Hg0 is removed. However, when SO2 is introduced into flue gas, and as the SO2 concentration increases from 0 to 1800 ppm, the average Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample decreases greatly from 95.98% to 81.85%. The inhibitive effect of SO2 concentration on Hg0 removal efficiency may be caused by the 7
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following reasons. On the one hand, the competitive adsorption between Hg0 and SO2 reduces the content of active sites for removing the gaseous Hg0 [44,46]. On the other hand, some metal sulfates produced by the reaction of metal oxides and SO2 can cause the deactivation and poisoning of the sample [29,47]. Thus, a rise in the concentration of SO2 reduces the removal efficiency of Hg0.
by the reaction Eq. (2) [19].
Hg0 (g)
(2)
Hg0 (ad)
Secondly, the generation process of lattice oxygen ([O]) on the sample can be expressed by the following reaction Eqs. (3)–(5) [31,32,52].
3.3.4. Effect of water vapor concentration Fig. 6(d) displays the effect of water vapor concentration on Hg0 removing performance. Compared with the flue gas without water vapor, the addition of 4% water vapor can promote the removal of Hg0. Relevant research [48] indicates that the hydroxyl radicals derived from water molecule decomposition can react with the adsorbed Hg0 to generate HgO, thus promoting the removal of Hg0. But when 12% water vapor is added into flue gas, the removal efficiency of Hg0 sharply drops from 90.58% to 71.58%. Firstly, the competition between water vapor and Hg0 for the same active sites is not conducive to removing the Hg0 [19,44]. Secondly, the water vapor can also combine with Cu-Fe mixed oxides and SO3 to generate sulfates, which easily causes the deactivation of the sample, thereby inhibiting the removal of Hg0 [29,41]. In addition, the water film, formed on the surface of the sorbent, also inhibits the removal of Hg0 [49].
2CuO
Cu2 O+ [O]
(3)
Fe2 O3
2FeO + [O]
(4)
Fe3 O4
3FeO + [O]
(5) 0
Thirdly, the adsorbed Hg on the surface of sample can combine with lattice oxygen to generate HgO. In addition, the adsorbed Hg0 can also be oxidized by chemisorbed oxygen to produce HgO. The specific reaction steps can be expressed by the following reaction Eqs. (6)–(8) [32,52].
Hg0 (ad) + [O]
HgO(ad)
(6)
Hg0 (ad) + O*
HgO(ad)
(7)
HgO(ad)
(8)
HgO(g)
Finally, the chemisorbed oxygen (O*) and lattice oxygen ([O]), which are consumed in the process of Hg0 removal, will be efficiently replenished and regenerated by introducing gaseous oxygen into the flue gas. The reaction process can be described by the following Eqs. (9)–(12) [19,31,52].
0
3.4. Reaction mechanism of Hg removal To investigate the reaction mechanism of Hg0 removal, the chemical states of some elements on the fresh and spent samples were tested by XPS. Fig. 7(a) and (b) present the Hg 4f XPS analysis spectra of the fresh and spent samples. As seen in Fig. 7(a), only a strong peak appears at 103.1 eV on the fresh sample, which usually corresponds to Si 2p [19]. For the spent sample, in addition to the peak belonging to Si 2p, a new spectral peak at 104.1 eV is observed, which is assigned to Hg2+ [19]. The results suggest that Hg0 adsorbed on the surface of sample is existed in the form of divalent mercury. Fig. 7(c) and (d) show the O 1s XPS analysis spectra of the fresh and spent samples. It can be seen that there are three obvious overlapping peaks for the O 1s in the fresh and spent samples. The peak at 530.4/ 530.6 eV belongs to the lattice oxygen (denoted as [O]) [14,49]. The peak value of 531.5/531.6 eV corresponds to the chemisorbed oxygen (denoted as O*) [5,41]. Moreover, the peak at 533.3/532.7 eV has been attributed to C-O [4,14]. After the Hg0 removal tests, the content of lattice oxygen increases from 19.66% to 25.01%, while the content of chemisorbed oxygen decreases from 35.19% to 26.68%. These data indicate that chemisorbed oxygen plays an important role in the removal of Hg0. The spectral peaks of Fe 2p in the fresh and spent samples are shown in Fig. 7(e) and (f). Two main spectrum peaks corresponding to Fe 2p3/2 and Fe 2p1/2 are detected at 711.6 eV and 725.1 eV, respectively [50]. The Fe 2p3/2 peak consists of one Fe2+ peak (710.2 eV) and two Fe3+ peaks (711.3 eV and 713.1 eV) [14,19]. The Fe 2p1/2 peak at 725.1 eV is also attributed to Fe3+ [50]. After the Hg0 removal experiments, the content of Fe2+ increases from 2.70% to 17.64%, while the content of Fe3+ decreases from 97.30% to 82.36%. These results suggest that the transition from Fe3+ to Fe2+ occurs in the process of Hg0 removal. The Cu 2p XPS spectra for the fresh and spent samples are shown in Fig. 7(g) and (h). It can be seen that the Cu 2p1/2 band at 954.0–956.0 eV, the higher band of Cu 2p3/2 at 933.0–934.5 eV, and the shake-up peak at 939.0–942.0 eV are attributed to Cu2+ [25,51,52]. The lower band of Cu 2p3/2 at 932.2 eV belongs to Cu+ [26,32,53]. As shown in Fig. 7(g) and (h), the content of Cu+ in the spent sample increases from 8.98% (for the fresh sample) to 21.29%. Conversely, the content of Cu2+ in the spent sample decreases from 91.02% (for the fresh sample) to 78.71%. The results indicate that Cu2+ participates in the removal process of Hg0. Based on the above analysis and experiment results, the Hg0 removal mechanism of modified samples is proposed preliminarily. Firstly, the gas-phase Hg0 in flue gas is trapped on the surface of sample to generate the adsorbed Hg0. The formation process can be described
O2
(9)
2O*
Cu2 O+ 1/2O2
2CuO
(10)
2FeO + 1/2O2
Fe2 O3
(11)
3FeO + 1/2O2
Fe3 O4
(12)
3.5. Adsorption kinetics In this work, two adsorption kinetic models (the pseudo-first model and the pseudo-second model) are applied to study the removal mechanism and adsorption kinetic process of modified porous chars. The pseudo-first model is mainly used to describe the process of external mass transfer. The pseudo-second model can be used to describe the process of chemisorption. The pseudo-first model can be expressed by the following equation [4]:
lg(qe
qt ) = lg qe
k1 t 2.303
(13)
The pseudo-second model can be expressed by the following equation [54]:
t 1 1 = + t qt qe k2 qe2
(14)
where t (min) represents the reaction time. qe and qt (μg/g) represent the uptake amount of samples at time t and equilibrium time, respectively. k1 (1/min) and k2 (g/(μg·min)) represent the contents of the pseudo-first model and the pseudo-second model, respectively. The fitting curves and fitting parameters are shown in Fig. 8 and Table 3, respectively. As shown in Fig. 8(a) and Table 3, the correlation coefficients (R2), obtained by the pseudo-first model, are relatively low (< 0.96). This result indicates that the external mass transfer is not the rate-controlling step for the Hg0 removal process. Thus, the pseudo-first model cannot be used to describe the Hg0 removal process of modified porous chars [4]. It can be seen from Fig. 8(b) and Table 3 that the the pseudo-second model can fit the experimental data very well. In Table 3, compared with the pseudo-first model, the correlation coefficients (R2) of the pseudo-second model are higher than 0.96. These 8
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Fig. 7. The XPS analysis of the fresh and spent samples over the spectral regions of Hg 4f, O 1s, Fe 2p and Cu 2p.
results show that the pseudo-second model can be used to express the Hg0 adsorption process of modified samples. Therefore, the chemisorption process is the key rate-controlling step for Hg0 removal [54,55,56].
3.6. Comparison and regeneration of the modified porous char Table S1 listed the comparison of Hg0 adsorption capacity between the modified porous char and other activated carbon sorbents/catalysts. 9
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Fig. 8. Kinetic analysis of modified porous chars: (a) the pseudo-first model, (b) the pseudo-second model.
As shown in Table S1, compared with activated carbon sorbents/catalysts, the modified porous char (CuFe0.3/WSWU10(500)) exhibits the best Hg0 removal performance, and its Hg0 adsorption capacity is 2276.45 μg/g. In addition, the modified porous char in this paper mainly comes from cheap agricultural wastes, which is beneficial to reduce the cost of Hg0 removal. Therefore, the modified porous char (CuFe0.3/WSWU10(500)) has good development and application prospects due to its high Hg0 adsorption capacity and low cost. In addition to high Hg0 adsorption capacity, the regeneration performance of modified porous char after Hg0 removal is also very important for its future application and development. Therefore, in this work, the spent CuFe0.3/WSWU10(500) sample was first heated at 400 °C for 60 min under the protection of nitrogen, and then regenerated at 250 °C for 30 min under air atmosphere. The Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample in six cycles is shown in Fig. 9. It can be seen from Fig. 9 that the sample still has a high Hg0 removal efficiency after six regeneration cycles. After the sixth cycle regeneration, the Hg0 removal performance of the sample is basically stable, and the Hg0 removal efficiency is about 76.33%. These results suggest that the CuFe0.3/WSWU10(500) sample exhibits good regeneration characteristics.
Fig. 9. The Hg0 removal efficiency of the CuFe0.3/WSWU10(500) sample in six cycles. Experimental conditions: NO concentration, 400 ppm; SO2 concentration, 600 ppm; O2 concentration, 5%, H2O concentration, 4% (vol); Hg0 concentration, 65 μg/m3 and reaction temperature, 130 °C.
4. Conclusions and about 90.58% of Hg0 is removed from flue gas at 130 °C. Increasing the concentration of O2 and NO can clearly promote the removal of Hg0. Increasing SO2 concentration inhibits Hg0 removal. Low concentration of water vapor improves Hg0 removal performance, while high concentration of water vapor plays an opposite role. The copper/ iron active components, chemisorbed oxygen and lattice oxygen play an important role in the process of Hg0 removal. The modified sample exhibits good regeneration performance.
In this article, a novel magnetic iron-copper oxide modified biomass-based porous char was synthesized by microwave activation and ultrasonic-assisted impregnation method and was used to control Hg0 emissions from flue gas. The effects of operating parameters and flue gas components on Hg0 removal were investigated. The reaction mechanism and the regeneration performance of samples were also studied. The results indicate that the ultrasound treatment promotes the Hg0 removal from flue gas. The microwave and steam activation greatly improves the pore structure of porous char. The CuFe0.3/ WSWU10(500) sample exhibits an excellent Hg0 removal performance, Table 3 Kinetic parameters of modified porous chars. Samples
CuFe0.3/WSWU5(500) CuFe0.3/WSWU10(500) CuFe0.3/WSWU15(500)
Pseudo-first model
Pseudo-second model 2
qe (μg/g)
k1 (1/min)
R
425.48 928.47 604.48
0.0037 0.0015 0.0019
0.9577 0.9139 0.9197
10
qe (μg/g)
k2 (g/(μg·min))
R2
1027.81 2276.45 1592.17
1.2384 × 10−6 0.2729 × 10−6 0.5559 × 10−6
0.9786 0.9962 0.9646
Fuel 256 (2019) 115977
W. Yang, et al.
Acknowledgements
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