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Experimental study on elemental mercury removal by diperiodatonickelate (IV) solution
Journal of Hazardous Materials 260 (2013) 383–388
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Experimental study on elemental mercury removal by diperiodatonickelate (IV) solution Yi Zhao a,∗ , Fangming Xue a , Xiaochu Zhao b , Tianxiang Guo a , Xiaolei Li a a b
School of Environmental Science and Engineering, North China Electric Power University, Baoding 071003, PR China Haidian Branch, Beijing Electric Power Supply Company, Beijing 100000, PR China
h i g h l i g h t s • Diperiodatonickelate (IV) solution was prepared and firstly used to removal of Hg0 . • Removal efficiencies of 86.2% were obtained in the presence of SO2 and NO. • The simultaneous removal mechanism of Hg0 was proposed.
a r t i c l e
i n f o
Article history: Received 19 February 2013 Received in revised form 19 May 2013 Accepted 21 May 2013 Available online 29 May 2013 Keywords: Diperiodatonickelate (IV) Hg0 removal Reaction mechanism Liquid phase
a b s t r a c t A novel method has been developed to remove elemental mercury from simulated flue gas by a diperiodatonickelate (IV) solution. The influencing factors, such as diperiodatonickelate (IV) concentration, reaction temperature, solution pH, the initial Hg0 concentration, SO2 concentration and NO concentration were investigated at a bubbling reactor. In the presence of SO2 and NO, removal efficiency of 86.2% for elemental mercury was obtained. Meanwhile, 56.2% of NO and 98% of SO2 were simultaneously removed, under the optimal experimental conditions, in which diperiodatonickelate (IV) concentration was 6 × 10−3 mol/L, reaction temperature was 50 ◦ C, the initial Hg0 concentration was 20 g/m3 and pH was 8.5. Moreover, based on the research results of the hydrolyzing products of IO4− , and the analysis of the removal products of Hg0 , the reaction mechanism of Hg0 removal was proposed. © 2013 Elsevier B.V. All rights reserved.
1. Introduction The amount of mercury in Earth’s biosphere is increasing gradually due to both natural and anthropogenic emissions. Worldwide mercury emissions from human activities are currently estimated to be 1000–6000 t/annum, which accounts for 30–55% of global atmospheric mercury emissions [1–3]. As a result, according to some reports, between 300,000 and 600,000 children are born every year with neurological problems, such as a decreased intelligence quotient. The globally increasing number of autism cases is partially attributed to exposure to mercury and distance from power plants [4,5]. The major mechanism by which anthropogenic mercury finds its way into the environment is the release from coal-fired power plants [1,6]. Throughout the coal combustion process, mercury speciates to the following three forms: elemental mercury (Hg0 ), oxidation mercury (Hg2+ ) and particulate-bound mercury (HgP ). Mercury removal rates among existing air pollution control devices (APCDs), ranging from 0 to 90%, vary significantly
∗ Corresponding author. Tel.: +86 0312 7522343; fax: +86 0312 7522192. E-mail address: [email protected] (Y. Zhao). 0304-3894/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2013.05.040
depending on coal type, fly ash properties and specific APCD configuration [7]. For example, Hg2+ can be removed by wet flue gas desulfurization (WFGD), and HgP is easily collected by electrostatic precipitators (ESP) or Fabric Filter (FF) [8]. Nevertheless, Hg0 is difficult to be captured by existing APCDs, because of its high volatility and low solubility in water [9]. Previous research has shown that Hg0 can be transported long distances in the atmosphere where its toxic effects can have global-scale impacts; for example, mercury has been shown to enter the food chain and interfere with ozone depletion in the arctic [10,11]. Hence, in recent years, Hg0 removal from flue gas has caused wide concern from both academia and industry. For Hg0 removal in liquid phase, the key point is to convert Hg0 to Hg2+ rapidly, the latter being dissolved easily in water. There have been several reports on the rapid oxidation of Hg0 [12,13], however, most of them have technical and economic disadvantages and have not the potential to develop into practicable technologies. Various oxidants such as sodium chlorite, permanganate, ozone, Mn/␥-Fe2 O3 , electrospun metal oxide-TiO2 nanofibers, membrane catalytic system, etc., are often used to oxidize Hg0 [14–18]. These classical chemicals either have lower economical efficiencies or may release several hazardous byproducts that can adversely affect environment. For example, sodium chlorite is considered as one
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Fig. 1. Experimental apparatus. (1) Steel bottle of SO2 ; (2) steel bottle of NO; (3) steel bottle of N2 ; (4) relief valve; (5) 60 mL flowmeter; (6) 1 L flowmeter; (7) 100 mL flowmeter; (8) thermostat water bath; (9) Hg0 generator; (10) buffer bottle; (11) bubble reactor; (12) tail gas absorption bottle; (13) tail gas outlet; (14) flue gas analyzer; (15) Hg0 analyzer; (16) computer; (17) sampling port; 18.1 mL injector; (A–C) two-way valve gate.
of the most effective reagents, but it is estimated that about 1.38 pounds of NaClO2 are needed to remove a pound of Hg0 . Therefore, the cost of absorbent is prohibitive. Higher Hg0 removal efficiency may also be obtained by using permanganate as an absorbent; this option is however too expensive. In addition, for two chemicals, a large number of heavy metal, manganese, and chlorine species that cause secondary pollution can remain in the removal products from process. Although ozone is an environmentally benign and effective absorbent, the energy consumption for the generation of ozone is excessive. Therefore, development of high efficiency, low pollution and economic technologies for Hg0 removal is still necessary and urgent. Because of its strong oxidizing power in alkaline medium, diperiodatonickelate (IV) (DPN) has extensively been used to oxidize some organic matters, such as l-asparagine [19], atenolol [20], 1,10-phenanthroline [21] and so on. To our knowledge, there are no reports about the application of DPN for removing Hg0 from flue gas. In this paper, a DPN solution was prepared according to the literatures [22,23] to remove Hg0 . Compared with the activated carbon injection technology, this absorbent has an obvious advantage in the economy. It is estimated that for 1 kg of Hg0 removal, the cost of the activated carbon is $55,114–154,320 [24], and that of DPN solution is approximately $5447 [22]. As far as environmental problem is concerned, Ni2+ , the main reduction product of DPN has the lower environmental effect than the activated carbon adsorbing mercury and can easily be removed by wastewater treatment device. Based on the operating principle of Chiyoda Thoroughbred121(CT-121) bubble reactor used in industry, the experiments on removal of Hg0 were carried out at a self-made bubble reactor. The influencing factors, such as DPN concentration, reaction temperature, the initial Hg0 concentration, solution pH, SO2 concentration and NO concentration were investigated. Furthermore, the presumed reaction mechanism of Hg0 removal was proposed. As a new effective Hg0 emission control process for flue gas cleaning, this work has great academic significance and an important reference for performing an experiment with higher volumes and flows or practical application.
2. Experimental materials and methods 2.1. Experimental equipment and reagents The experimental system is illustrated in Fig. 1, in which the key part of the experimental system is a bubble reactor (home-made) with 250 mL of the effective volume and 15.5 cm of height, on which a gas blanket of micron porous core fabric is located at 1.5 cm far from the bottom of reactor. Simulated flue gases such as SO2 , NO, and N2 were supplied from the compressed gas steel cylinder (North Special Gas Co., Ltd., China). Hg0 was generated from mercury permeation device (VICI Metronics Co., USA) heated in a thermostatic water bath with N2 as a carrier gas. Mass flow controller (LZB, Tianjin Flow Meter Co., Ltd., China) controlled the flow rate of each gas. All reagents used were analytical reagent (AR) (Tianjin Chemical Reagents Company); all solutions were prepared by high purity water (specific resistance > 18.25 M/cm); the DPN solution was prepared and calibrated referring to Ray’s method [23]. Tail gas absorption bottle filled with 10% (v/v) H2 SO4 -4% (w/w) KMnO4 , was used to absorb the rest of the pollutant gases. 2.2. Experimental methods During the experiments, SO2 , NO and N2 were metered through mass flow controller (7) and mixed with Hg0 vapor resulting from mercury osmotic tube in buffer bottle (10), in which SO2 , NO and Hg0 were diluted by N2 to the desired concentrations, from which the simulated flue gas formed. The oxidization and absorption reactions occurred when simulated flue gas with 1 L/min of the total flow entered into bubbling reactor (11). The analysis of Hg0 was carried out based on the US EPA method 101A [24], and the used Hg0 analyzer was QM201 cold atom fluorescence mercury detector (Suzhou, Qingan Instrument Company, China). The concentrations of SO2 and NO were measured by a Delta2000CD-IV flue gas analyzer (Heilbronn, MRU, Germany) (14). The removal efficiencies were calculated according to the concentrations of Hg0 , SO2 and NO before and after absorption. Then, the spent simulated flue gas
Y. Zhao et al. / Journal of Hazardous Materials 260 (2013) 383–388
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Fig. 2. Effects of the DPN concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; reaction temperature, 50 ◦ C; pH, 8.5; stabilization absorption time, 60 min.
Fig. 3. Effects of reaction temperature on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 5 × 10−4 mol/L; pH, 8.5; stabilization absorption time, 60 min.
was discharged into the atmosphere after being treated by tail gas absorption bottle (12). In addition, the reaction temperature was regulated by a digital control thermostatic water baths (Shanghai, Yichang Instrument Company, China). The solution pH was adjusted by mixed acid and alkali (phosphoric acid, acetic acid, boric acid and sodium hydroxide) buffer and measured by a pH meter (PHSJ-5, Shanghai Leici Instrument Company, China).
temperature ranging from 20 to 50 ◦ C. The highest removal efficiency was obtained at 50 ◦ C, thereafter, the removal efficiency decreased sharply. Accordingly, the best reaction temperature was selected as 50 ◦ C. In this experiment, Hg0 removal is a gas–liquid reaction that is affected significantly by the reaction temperature, in which the absorption reaction between Hg0 and DPN could be improved by increasing the reaction temperature [25,26]. Because Hg0 is difficult to dissolve in aqueous solution, Hg0 solubility will decrease with an increase of temperature according to Henry’s law, which may leads to the decrease of the efficiency [27]. 50 ◦ C may be a turning point of the promotion for removal reaction and inhibition due to the solubility of Hg0 decreasing. This result is consistent with that of the removal of Hg0 using potassium persulfate and potassium permanganate [28] had been added to discuss a decrease in removal efficiency after 50 ◦ C. In addition, the concentration of stronger power species, [Ni(H3 IO6 )2 (OH)2 ]2− might be decreased due to a change of the dissociation constant in Eqs. (3)–(5) after 50 ◦ C
2.3. Quality assurance and quality control All vessels used were made of borosilicate glass or polytetrafluoroethylene (PTFE) and cleaned thoroughly according to the US EPA method 101A. In order to ensure experimental stability, each experiment had five parallel tests. The correlation coefficient, R2 , of the calibration curve was higher than 0.995. Relative standard deviation (RSD) of each group parallel tests was lower than 0.75%. 3. Results and discussion
3.3. Effect of pH on Hg0 removal 3.1. Effect of DPN concentration on Hg0 removal The variation of removal efficiency for Hg0 removal as a function of the concentration of DPN is shown in Fig. 2. It can be seen from Fig. 2 that the removal efficiency enhances fast when the concentration of DPN is between 2 × 10−4 and 5 × 10−4 mol/L, however, the change occurs at 5 × 10−4 mol/L. The removal efficiency increases very slowly in DPN concentration ranging from 5 × 10−4 to 6 × 10−4 mol/L, because the total rate constant decrease when the DPN concentration is above 5 × 10−4 mol/L. Similar trends were obtained from Mulla and Nandibewoor’s research [25]. From the economy, the optimal DPN concentration was selected as 5 × 10−4 mol/L on Hg0 removal alone. Previous experiments carried out by Hiremath, Mulla and Liu [21,25,26,29] indicated that [Ni(H3 IO6 )2 (OH)2 ]2− was considered to be the active species and the reactions responding to [Ni(H3 IO6 )2 (OH)2 ]2− were the pseudo first order and double electron transfer mechanism. These results are benefit to establishing the optimal experimental conditions and indicating the reaction pathway.
The experiments for Hg0 removal were carried out at different pH ranging from 7.5 to 12.0. The results are shown in Fig. 4. It is observed from Fig. 4 that the Hg0 removal efficiency increase when pH is between 7.5 and 8.5, thereafter, the removal efficiencies of Hg0 decrease rapidly, which might be due to the change
3.2. Effect of reaction temperature on Hg0 removal Hg0 removal efficiency was measured at different reaction temperatures with keeping other conditions constant (Fig. 3). The removal efficiency of Hg0 increased slowly with reaction
Fig. 4. Effects of pH on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 5 × 10−4 mol/L; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
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Fig. 5. Effects of the initial Hg0 concentration on Hg0 removal. DPN concentration, 5 × 10−4 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
of the existing form of DPN under the different pH. It can be seen from Fig. 9 that the concentration of stronger power species, [Ni(H3 IO6 )2 (OH)2 ]2− achieves the highest when the pH is from 8.5 to 10.5, and decreases the lowest at pH 12. In this case, the week oxidizing species, [Ni(H2 IO6 )2 (OH)2 ]4− , is the main existing form of DPN, the previous evidences obtained from Mulla and Nandibewoor confirm this finding [26,29]. Therefore, there is a decrease in removal efficiency at pH 12. According to the results of experiments, the optimum solution pH was selected as 8.5.
Fig. 6. Effects of SO2 concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
HgSO3 + SO3 2− ↔ Hg(SO3 )2− 2
(2)
However, the removal efficiency decreased sharply when SO2 concentration increased continuously to 3590 mg/m3 . It meant that the higher concentration SO2 seemed to go against the removal of Hg0 , which might be due to the competing reaction of SO2 . In addition, during the reaction process, the removal efficiency of SO2 reached 98%.
3.4. Effect of the initial Hg0 concentration on Hg0 removal
3.6. Effects of NO concentration on Hg0 removal
The data of actual coal-fired power plants combustion processes indicates that the concentration of Hg0 varies with the content of mercury in burned coal and operating conditions of boilers. Experiments of Hg0 removal under different initial Hg0 concentrations were carried out to examine the adaptability of DPN for different coal species and operating conditions of boilers, as shown in Fig. 5. The removal efficiencies vary slightly when the concentrations of Hg0 ranges from 20 to 40 g/m3 , indicating that the method proposed adapts to different types of coal and combustion conditions when the removal efficiencies demand in between 74.1 and 69.2%.
As shown in Fig. 7, NO has a slight promoting effect for removing Hg0 efficiency from 74.1% to 78.2% when NO concentration increases from 0 to 413 mg/m3 . The improvement for Hg0 removal can be explained by that, in the absence of buffer solution, with an increase of NO concentration, the solution pH raised because the oxidation of NO by DPN solution was a reaction of releasing OH− [31], which was helpful to form the strong oxidizing species, [Ni(H3 IO6 )2 (OH)2 ]2− and to promote the removal efficiency. On the contrary, the removal efficiency of Hg0 fell off with a continued increase of NO concentration, which might be due to the competing reaction of NO. Under this experimental condition, the removal efficiency of NO was about 53.8%.
3.5. Effects of SO2 concentration on Hg0 removal According to the conclusion of Section 3.1, DPN concentration was determined to be 0.5 mmol/L on Hg0 removal alone. However, 0.5 mmol/L DPN was difficult to removal NO and SO2 because DPN was consumed quickly when NO and SO2 were added in simulated flue gas. Thus, the optimal DPN concentration for simultaneous removing NO, SO2 and Hg0 was investigated. According to the experimental results, DPN concentration used in Sections 3.5, 3.6 and 3.7 was selected as 6 × 10−3 mol/L. Fig. 6 shows the change of the removal efficiency of Hg0 with the SO2 concentration, from Fig. 6, one finds that the SO2 concentration has a dramatic effect on Hg0 removal. The removal efficiency of Hg0 increases from 74.1 to 83.5% when SO2 concentration increases from 0 to 1258 mg/m3 . The promoting action of SO2 for removing Hg0 may also be attribute to the complex of Hg·S(IV) [30]. In Eqs. (1) and (2), the removal product, Hg2+ reacted with SO3 2− to form HgSO3 , and then sequentially reacted with SO3 2− to form Hg(SO3 )2 2− which was more stable than HgSO3 . Therefore, the removal efficiency of Hg0 could be improved by an increase of SO2 concentration. Hg2+ + SO3 2− ↔ HgSO3
(1)
Fig. 7. Effects of NO concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
Y. Zhao et al. / Journal of Hazardous Materials 260 (2013) 383–388
Fig. 9. Concentration ratios of IO4 − and its dissociation products in [IO4 − ]ex .
Fig. 8. Experiments on simultaneous removal of SO2 , NO and Hg0 . Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min; SO2 concentration, 1258 mg/m3 ; NO concentration, 413 mg/m3 .
Table 1 Concentration of Hg2+ in the spent absorption solution.a
3.7. Experiments on simultaneous removal of SO2 , NO and Hg0 The experiments of removing SO2 , NO and Hg0 were carried out to verify the effect of simultaneous removal under the optimal conditions of Hg0 removal, in which reaction temperature was 50 ◦ C, DPN concentration was 6 mmol/L, solution pH was 8.5 and the initial Hg0 concentration was 20 g/m3 . As shown in Fig. 8, 86.2% of Hg0 , 56.2% of NO and 98% of SO2 are simultaneously removed from the simulated flue gas and the efficiencies are stable in the reaction time ranging from 10th to 60th minute. In addition, comparing Fig. 8 with Figs. 2–4 and Fig. 6, the removal efficiency of Hg0 increases obviously, which verifies further the promotion of SO2 and NO for Hg0 removal.
− 2− IO− 4 + OH + H2 O ⇔ H3 IO6 , − 3− IO− , 4 + 2OH ⇔ H2 IO6
(3)
ˇ2 = 1.62 × 106
(4)
ˇ3 = 4.68 × 108
2
2
− − [H3 IO6 2− ] = ˇ2 × [OH− ] × [IO− 4 ] = f2 ([OH ]) × [IO4 ] 2
− − − − [H2 IO3− 6 ] = ˇ3 × [OH ] × [IO4 ] = f3 ([OH ]) × [IO4 ]
[IO− 4] =
−(f2 ([OH− ]) + f3 ([OH− ]) + 1) +
3
4
5
Average
0 200
0 197
0 201
0 199
0 199
The concentration unit of the measurement values is ng/L.
2
f1 ([OH− ]) = ˇ1 × [OH− ] , −
− 2
f2 ([OH− ]) = ˇ2 × [OH− ],
[IO4 − ]ex
[Ni(H3 IO6 )2 (OH)2 ]2− + OH− ⇔ [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− +H2 O (11) In order to reveal the reaction processes of Hg0 removal, the concentrations of Hg2+ in the spent absorption solutions were measured, and the results are shown in Table 1 in which scan blank was taken from fresh DPN solution, and sample 1 was the spent absorption solution. The data of Table 1 means that the oxidation of Hg0 occurs during the reaction process.
(6) (7) (8) 2
(f2 ([OH− ]) + f3 ([OH− ]) + 1) + 4f1 ([OH− ]) × [IO− 4 ]ex 2f1 ([OH− ])
(10)
presents total concentration f3 ([OH ]) = ˇ3 × [OH ] and of periodate in solution. As shown in Fig. 9, the existence form of [IO4 − ]ex is significantly affected by solution pH. When the concentration of [IO4 − ]ex was 5 × 10−4 mol/L, the concentration ratios of IO4 − and its hydrolyzing products were calculated at different pH. When solution pH varies from 7.5 to 8.5, the main species of [IO4 − ]ex are IO4 − and H3 IO6 2− ; when solution pH varies from 8.5 to 11.0, that is H3 IO6 2− ; when solution pH varies from 11.0 to 12.0, H3 IO6 2− and H2 IO6 3− are the main existence form. So the main existence form of DPN should be [Ni(H3 IO6 )2 (OH)2 ]2− under the optimum solution pH (at 8.5), and the others were [Ni(H2 IO6 )(H3 IO6 )(OH)2 ]3− and H2 I2 O10 4− , the latter may be ignored [34] due to its very low concentration ratio. These results are consistent with the researches of Yang [35] and Chimatadar [19]. In the kinetic study of DPN by Harihar [20], the conclusion of [Ni(H3 IO6 )2 (OH)2 ]2− as the main reactive form of DPN was also proposed. Furthermore, it was reported that part of [Ni(H3 IO6 )2 (OH)2 ]2− could be changed into stronger power species, [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− , which was supported by the Hiremath’s studies on the observed fractional order in [OH− ] [19–21]:
The expressions of the concentration of IO4 − and its hydrolyzing products are shown in Eqs. (6)–(9), which are obtained from Eqs. (3)–(5): 2
2
0 198
where
(5)
− − [H2 I2 O10 4− ] = ˇ1 × [OH− ] × [IO− 4 ] = f1 ([OH ]) × [IO4 ]
1
Scan blank Sample 1
4− [IO− ] + [H3 IO6 2− ] + [H2 IO6 3− ] + [IO− 4 ]ex = [H2 I2 O10 4]
A key component, IO4− in the absorbent can be decomposed to a variety of species, which are closely related to the existing form of DPN. For the speculation of the removal mechanism of Hg0 , it is important to investigate the hydrolyzing products of IO4− . According to previous research results, the water soluble Ni(IV) periodate complex was reported to be [Ni(HIO6 )2 (OH)2 ]6− [32]. However, in an alkaline medium and at the pH used in our study, the dissociation reactions of IO4− and their dissociation constant in alkaline medium at 25 ◦ C are summarized as follows [33]: ˇ1 = 1.12 × 1015
Determination time
a
3.8. Removal mechanism discussion
− 4− 2IO− , 4 + 2OH ⇔ H2 I2 O10
387
(9)
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Based on the research results of the hydrolyzing products of IO4 − , and the analysis of the removal products of Hg0 , the reaction mechanism of Hg0 removal is inferred as follows: Hg(g) → Hg(l)
(12)
[Ni(H3 IO6 )2 (OH)2 ]2− + OH− ⇔ [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− +H2 O (13)
[Ni(H3 IO6 )2 (OH)2 ]2− + Hg(l) → [Ni(H3 IO6 )2 (OH)2 ]4− + Hg2+ (14) [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− + Hg(l) → [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]5− + Hg2+
(15)
4. Conclusions The factors influencing the removal of Hg0 from simulated flue gas by DPN, namely DPN concentration, reaction temperature, solution pH, initial Hg0 concentration, SO2 and NO concentration, are described for the first time in this study. The experimental results indicated that DPN had higher removal efficiency of Hg0 under the optimum experimental conditions. The reaction mechanism of Hg0 removal was proposed, in which [Ni(H3 IO6 )2 (OH)2 ]2− was the main existing and active form under the optimum experimental condition, and a partial [Ni(H3 IO6 )2 (OH)2 ]2− could be transformed to stronger power species, [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− . For Hg0 removal, it was attributed to the oxidation of Hg0 by these two species to Hg2+ . Acknowledgments We appreciate the financial supports of National Hightech Research and Development Projects (863 program, No. 2013AA65403), Program for Changjiang Scholars and Innovative Research Team in University (IRT1127) and Zhejiang Provincial Engineering Research Center of Industrial Boiler & Furnace Flue Gas Pollution Control, Hangzhou, PR China (311202). References [1] R. Yan, D.T. Liang, L. Tsen, Bench-scale experimental evaluation of carbon performance on mercury vapor adsorption, Fuel 83 (2004) 2401–2409. [2] S.H. Lee, Y.J. Rhim, S.P. Cho, Carbon-based novel sorbent for removing gas-phase mercury, Fuel 85 (2006) 219–226. [3] T.G. Lee, P. Biswas, E. Hedrick, Comparison of Hg0 capture efficiencies of three in situ generated sorbents, Aiche J. 47 (2001) 954–961. [4] Z. Barnea, T. Sachs, M. Chidambaram, Y. Sasson, A novel oxidative method for the absorption of Hg0 from flue gas of coal fired power plants using task specific ionic liquid scrubber, J. Hazard. Mater. 244–245 (2013) 495–500. [5] J. Wilcox, E. Rupp, S.C. Ying, D.H. Lim, A.S. Negreira, A. Kirchofer, F. Feng, K. Lee, Mercury adsorption and oxidation in coal combustion and gasification processes, Int. J. Coal Geol. 90–91 (2012) 4–20. [6] International Joint Commission and the Commission for Environmental Cooperation, Consultation on Emissions from Coal-Fired Electrical Utilities, Quebec, Canada, 2004, pp. 20–21. [7] National Bureau of Statistics of China, China Statistical Yearbook, Chinese Statistic Press, Beijing, 2007, pp. 261–265. [8] Y.J. Wang, N.F. Duan, L.G.S. Yang, L. Meng, Z.J. Huang, Experimental study on mercury removal by combined wet flue gas desulphurization with electrostatic precipitator, P. Chinese Soc. Electr. Eng. 28 (2008) 64–69 (in Chinese).
[9] K.L. Shawn, Y. Cao, Y. Duan, Factors affecting mercury speciation in a 100 MW coal-fired boiler with low-NOx burners, Energ. Fuel 19 (2005) 800–806. [10] P.A. Ariya, A.F. Khalizov, A. Gidas, Reaction of gaseous mercury with atomic and molecular halogens: kinetics, product studies, and atmospheric implication, J. Phys. Chem. A 106 (2002) 7310–7320. [11] N. Pirrone, G.J. Keeler, P.O. Warner, Trends of ambient concentrations and deposition fluxes of particulate trace metals in Detroit from 1982 to 1992, Sci. Total Environ. 162 (1995) 43–61. [12] Q.F. Ye, C.Y. Wang, D.H. Wang, G. Sun, X.H. Xu, Hg0 absorption in potassium persulfate solution, J. Zhejiang Univ. Sci. B. 7 (2006) 404–410. [13] X.H. Xu, Q.F. Ye, T.M. Tang, D.H. Wang, Hg0 oxidative absorption by K2 S2 O8 solution catalyzed by Ag+ and Cu2+ , J. Hazard. Mater. 158 (2008) 410–416. [14] National Bureau of Statistics of China, China Statistical Yearbook, Chinese Statistic Press, Beijing, 2007, pp. 552–557. [15] K.F. Cen, Technical report of clean coal-fired power generation, in: 2nd Seminar for the Clean Coal Technology of China Coal Society, Hangzhou, 2001. [16] S. Yang, Y. Guo, N. Yan, Z. Qu, J. Xie, C. Yang, J. Jia, Capture of gaseous elemental mercury from flue gas using a magnetic and sulfur poisoning resistant sorbent Mn/␥-Fe2 O3 at lower temperatures, J. Hazard. Mater. 186 (2011) 508–515. [17] Y. Yuan, Y. Zhao, H. Li, Y. Li, X. Gao, C. Zheng, J. Zhang, Electrospun metal oxideTiO2 nanofibers for elemental mercury removal from flue gas, J. Hazard. Mater. 227–228 (2012) 427–435. [18] Y. Guo, N. Yan, S. Yang, P. Liu, J. Wang, Z. Qu, J. Jia, Conversion of elemental mercury with a novel membrane catalytic system at low temperature, J. Hazard. Mater. 213–214 (2012) 62–70. [19] S.A. Chimatadar, T. Basavaraj, S.T. Nandibewoor, A study of the kinetics and mechanism of oxidation of L-tryptophan by diperiodatonickelate (IV) in aqueous alkaline medium, Russ. J. Phys. Chem. A 81 (2007) 1046–1053. [20] A.L. Harihar, M.R. Kembhavi, S.T. Nandibewoor, Kinetics and mechanism of the oxidation of 1, 10-phenanthroline by diperiodatonickelate (IV) in aqueous alkaline medium, Monatsh. Chem. 131 (2000) 1129–1137. [21] G.C. Hiremath, R.M. Mulla, S.T. Nandibewoor, Kinetic, mechanistic and spectral investigation of ruthenium (III) catalysed oxidation of atenolol by alkaline diperiodatonickelate (IV) (stopped flow technique), Catal Lett. 98 (1) (2004) 49–56. [22] A.G. Lappin, M.C. Laranjeira, O.L. Youdi, Kinetics and mechanism of reduction of a nickel (IV) complex by ascorbate ion, J. Chem. Soc. Dalton Trans. (1981) 721–725. [23] P. Ray, T. Moeller, Inorganic Synthesis, McGraw-Hill, New York, 1957. [24] U. S. EPA, Appendix to Method 1631: Determination of Particulate and Gaseous Mercury Emissions from Sewage Sludge Incinerators, U. S. EPA, Washington, DC, 1980. [25] R.M. Mulla, H.M. Gurubasavaraj, S.T. Nandibewoor, Kinetics of ruthenium (III)catalysed oxidation of paracetamol by diperiodatonickelate (IV) in aqueous alkaline medium (stopped flow technique), Applied Catalysis A: General 314 (2006) 208–215. [26] Y. Liu, Y. Zhang, Z. Liu, K. Deng, Graft copolymerization of butyl acrylate onto case in initiated by potassium diperiodatonickelate(IV) in alkaline medium, Eur. Polym. J. 38 (2002) 1619–1625. [27] Y. Zhao, S.T. Liu, J. Yao, X.Y. Ma, C.M. Chen, Removal of Hg0 with sodium chlorite solution and mass transfer reaction kinetics, Sci. China Ser. E. 53 (2010) 1258–1265. [28] Q.F. Ye, C.Y. Wang, X.H. Xu, D.C. Wang, Mass transfer-reaction of Hg0 absorption in potassium permanganate, J. Zhejiang Univ. -Sci. B. 41 (2007) 831–835. [29] C.V. Hiremath, D.C. Hiremath, S.T. Nandibewoor, Ruthenium(III) catalysed oxidation of gabapentin (neurontin) by diperiodatonickelate(IV) in aqueous alkaline medium: a kinetic and mechanistic study, J. Mol. Catal. A: Chem. 269 (2007) 246–253. [30] J.C. Chang, S.B. Ghrishi, Simulation and evaluation of elemental mercury concentration increase in flue gas across a wet scrubber, Environ. Sci. Technol. 37 (2003) 5763–5766. [31] D. Xia, C. He, L. Zhu, Y. Huang, H. Dong, M. Su, M.A. Asi, D. Bian, A novel wetscrubbing process using Fe(VI) for simultaneous removal of SO2 and NO, J. Environ. Monit. 13 (2011) 864–870. [32] R. Chang, Physical chemistry with applications to biological systems, Macmillan, New York, 1981, pp. 326. [33] A. Javeston, Hydrolysis of potassium periodate: ultracentrifugation, potentiometric traction, and Raman spectra, J. Chem. Soc. A (1969) 273–275. [34] C.E. Crouthamel, H.V. Meck, D.S. Martin, C.V. Banks, Spectrophotometric studies of dilute aqueous periodate solutions, J. Am. Chem. Soc. 73 (1951) 82–87. [35] C.Y. Yang, Z.J. Zhang, Z.L. Shi, A novel chemiluminescence reaction system for the determination of lincomycin with diperiodatonickelate (IV), Microchim. Acta 168 (2010) 293–298.
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Journal of Hazardous Materials journal homepage: www.elsevier.com/locate/jhazmat
Experimental study on elemental mercury removal by diperiodatonickelate (IV) solution Yi Zhao a,∗ , Fangming Xue a , Xiaochu Zhao b , Tianxiang Guo a , Xiaolei Li a a b
School of Environmental Science and Engineering, North China Electric Power University, Baoding 071003, PR China Haidian Branch, Beijing Electric Power Supply Company, Beijing 100000, PR China
h i g h l i g h t s • Diperiodatonickelate (IV) solution was prepared and firstly used to removal of Hg0 . • Removal efficiencies of 86.2% were obtained in the presence of SO2 and NO. • The simultaneous removal mechanism of Hg0 was proposed.
a r t i c l e
i n f o
Article history: Received 19 February 2013 Received in revised form 19 May 2013 Accepted 21 May 2013 Available online 29 May 2013 Keywords: Diperiodatonickelate (IV) Hg0 removal Reaction mechanism Liquid phase
a b s t r a c t A novel method has been developed to remove elemental mercury from simulated flue gas by a diperiodatonickelate (IV) solution. The influencing factors, such as diperiodatonickelate (IV) concentration, reaction temperature, solution pH, the initial Hg0 concentration, SO2 concentration and NO concentration were investigated at a bubbling reactor. In the presence of SO2 and NO, removal efficiency of 86.2% for elemental mercury was obtained. Meanwhile, 56.2% of NO and 98% of SO2 were simultaneously removed, under the optimal experimental conditions, in which diperiodatonickelate (IV) concentration was 6 × 10−3 mol/L, reaction temperature was 50 ◦ C, the initial Hg0 concentration was 20 g/m3 and pH was 8.5. Moreover, based on the research results of the hydrolyzing products of IO4− , and the analysis of the removal products of Hg0 , the reaction mechanism of Hg0 removal was proposed. © 2013 Elsevier B.V. All rights reserved.
1. Introduction The amount of mercury in Earth’s biosphere is increasing gradually due to both natural and anthropogenic emissions. Worldwide mercury emissions from human activities are currently estimated to be 1000–6000 t/annum, which accounts for 30–55% of global atmospheric mercury emissions [1–3]. As a result, according to some reports, between 300,000 and 600,000 children are born every year with neurological problems, such as a decreased intelligence quotient. The globally increasing number of autism cases is partially attributed to exposure to mercury and distance from power plants [4,5]. The major mechanism by which anthropogenic mercury finds its way into the environment is the release from coal-fired power plants [1,6]. Throughout the coal combustion process, mercury speciates to the following three forms: elemental mercury (Hg0 ), oxidation mercury (Hg2+ ) and particulate-bound mercury (HgP ). Mercury removal rates among existing air pollution control devices (APCDs), ranging from 0 to 90%, vary significantly
∗ Corresponding author. Tel.: +86 0312 7522343; fax: +86 0312 7522192. E-mail address: [email protected] (Y. Zhao). 0304-3894/$ – see front matter © 2013 Elsevier B.V. All rights reserved. http://dx.doi.org/10.1016/j.jhazmat.2013.05.040
depending on coal type, fly ash properties and specific APCD configuration [7]. For example, Hg2+ can be removed by wet flue gas desulfurization (WFGD), and HgP is easily collected by electrostatic precipitators (ESP) or Fabric Filter (FF) [8]. Nevertheless, Hg0 is difficult to be captured by existing APCDs, because of its high volatility and low solubility in water [9]. Previous research has shown that Hg0 can be transported long distances in the atmosphere where its toxic effects can have global-scale impacts; for example, mercury has been shown to enter the food chain and interfere with ozone depletion in the arctic [10,11]. Hence, in recent years, Hg0 removal from flue gas has caused wide concern from both academia and industry. For Hg0 removal in liquid phase, the key point is to convert Hg0 to Hg2+ rapidly, the latter being dissolved easily in water. There have been several reports on the rapid oxidation of Hg0 [12,13], however, most of them have technical and economic disadvantages and have not the potential to develop into practicable technologies. Various oxidants such as sodium chlorite, permanganate, ozone, Mn/␥-Fe2 O3 , electrospun metal oxide-TiO2 nanofibers, membrane catalytic system, etc., are often used to oxidize Hg0 [14–18]. These classical chemicals either have lower economical efficiencies or may release several hazardous byproducts that can adversely affect environment. For example, sodium chlorite is considered as one
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Fig. 1. Experimental apparatus. (1) Steel bottle of SO2 ; (2) steel bottle of NO; (3) steel bottle of N2 ; (4) relief valve; (5) 60 mL flowmeter; (6) 1 L flowmeter; (7) 100 mL flowmeter; (8) thermostat water bath; (9) Hg0 generator; (10) buffer bottle; (11) bubble reactor; (12) tail gas absorption bottle; (13) tail gas outlet; (14) flue gas analyzer; (15) Hg0 analyzer; (16) computer; (17) sampling port; 18.1 mL injector; (A–C) two-way valve gate.
of the most effective reagents, but it is estimated that about 1.38 pounds of NaClO2 are needed to remove a pound of Hg0 . Therefore, the cost of absorbent is prohibitive. Higher Hg0 removal efficiency may also be obtained by using permanganate as an absorbent; this option is however too expensive. In addition, for two chemicals, a large number of heavy metal, manganese, and chlorine species that cause secondary pollution can remain in the removal products from process. Although ozone is an environmentally benign and effective absorbent, the energy consumption for the generation of ozone is excessive. Therefore, development of high efficiency, low pollution and economic technologies for Hg0 removal is still necessary and urgent. Because of its strong oxidizing power in alkaline medium, diperiodatonickelate (IV) (DPN) has extensively been used to oxidize some organic matters, such as l-asparagine [19], atenolol [20], 1,10-phenanthroline [21] and so on. To our knowledge, there are no reports about the application of DPN for removing Hg0 from flue gas. In this paper, a DPN solution was prepared according to the literatures [22,23] to remove Hg0 . Compared with the activated carbon injection technology, this absorbent has an obvious advantage in the economy. It is estimated that for 1 kg of Hg0 removal, the cost of the activated carbon is $55,114–154,320 [24], and that of DPN solution is approximately $5447 [22]. As far as environmental problem is concerned, Ni2+ , the main reduction product of DPN has the lower environmental effect than the activated carbon adsorbing mercury and can easily be removed by wastewater treatment device. Based on the operating principle of Chiyoda Thoroughbred121(CT-121) bubble reactor used in industry, the experiments on removal of Hg0 were carried out at a self-made bubble reactor. The influencing factors, such as DPN concentration, reaction temperature, the initial Hg0 concentration, solution pH, SO2 concentration and NO concentration were investigated. Furthermore, the presumed reaction mechanism of Hg0 removal was proposed. As a new effective Hg0 emission control process for flue gas cleaning, this work has great academic significance and an important reference for performing an experiment with higher volumes and flows or practical application.
2. Experimental materials and methods 2.1. Experimental equipment and reagents The experimental system is illustrated in Fig. 1, in which the key part of the experimental system is a bubble reactor (home-made) with 250 mL of the effective volume and 15.5 cm of height, on which a gas blanket of micron porous core fabric is located at 1.5 cm far from the bottom of reactor. Simulated flue gases such as SO2 , NO, and N2 were supplied from the compressed gas steel cylinder (North Special Gas Co., Ltd., China). Hg0 was generated from mercury permeation device (VICI Metronics Co., USA) heated in a thermostatic water bath with N2 as a carrier gas. Mass flow controller (LZB, Tianjin Flow Meter Co., Ltd., China) controlled the flow rate of each gas. All reagents used were analytical reagent (AR) (Tianjin Chemical Reagents Company); all solutions were prepared by high purity water (specific resistance > 18.25 M/cm); the DPN solution was prepared and calibrated referring to Ray’s method [23]. Tail gas absorption bottle filled with 10% (v/v) H2 SO4 -4% (w/w) KMnO4 , was used to absorb the rest of the pollutant gases. 2.2. Experimental methods During the experiments, SO2 , NO and N2 were metered through mass flow controller (7) and mixed with Hg0 vapor resulting from mercury osmotic tube in buffer bottle (10), in which SO2 , NO and Hg0 were diluted by N2 to the desired concentrations, from which the simulated flue gas formed. The oxidization and absorption reactions occurred when simulated flue gas with 1 L/min of the total flow entered into bubbling reactor (11). The analysis of Hg0 was carried out based on the US EPA method 101A [24], and the used Hg0 analyzer was QM201 cold atom fluorescence mercury detector (Suzhou, Qingan Instrument Company, China). The concentrations of SO2 and NO were measured by a Delta2000CD-IV flue gas analyzer (Heilbronn, MRU, Germany) (14). The removal efficiencies were calculated according to the concentrations of Hg0 , SO2 and NO before and after absorption. Then, the spent simulated flue gas
Y. Zhao et al. / Journal of Hazardous Materials 260 (2013) 383–388
385
Fig. 2. Effects of the DPN concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; reaction temperature, 50 ◦ C; pH, 8.5; stabilization absorption time, 60 min.
Fig. 3. Effects of reaction temperature on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 5 × 10−4 mol/L; pH, 8.5; stabilization absorption time, 60 min.
was discharged into the atmosphere after being treated by tail gas absorption bottle (12). In addition, the reaction temperature was regulated by a digital control thermostatic water baths (Shanghai, Yichang Instrument Company, China). The solution pH was adjusted by mixed acid and alkali (phosphoric acid, acetic acid, boric acid and sodium hydroxide) buffer and measured by a pH meter (PHSJ-5, Shanghai Leici Instrument Company, China).
temperature ranging from 20 to 50 ◦ C. The highest removal efficiency was obtained at 50 ◦ C, thereafter, the removal efficiency decreased sharply. Accordingly, the best reaction temperature was selected as 50 ◦ C. In this experiment, Hg0 removal is a gas–liquid reaction that is affected significantly by the reaction temperature, in which the absorption reaction between Hg0 and DPN could be improved by increasing the reaction temperature [25,26]. Because Hg0 is difficult to dissolve in aqueous solution, Hg0 solubility will decrease with an increase of temperature according to Henry’s law, which may leads to the decrease of the efficiency [27]. 50 ◦ C may be a turning point of the promotion for removal reaction and inhibition due to the solubility of Hg0 decreasing. This result is consistent with that of the removal of Hg0 using potassium persulfate and potassium permanganate [28] had been added to discuss a decrease in removal efficiency after 50 ◦ C. In addition, the concentration of stronger power species, [Ni(H3 IO6 )2 (OH)2 ]2− might be decreased due to a change of the dissociation constant in Eqs. (3)–(5) after 50 ◦ C
2.3. Quality assurance and quality control All vessels used were made of borosilicate glass or polytetrafluoroethylene (PTFE) and cleaned thoroughly according to the US EPA method 101A. In order to ensure experimental stability, each experiment had five parallel tests. The correlation coefficient, R2 , of the calibration curve was higher than 0.995. Relative standard deviation (RSD) of each group parallel tests was lower than 0.75%. 3. Results and discussion
3.3. Effect of pH on Hg0 removal 3.1. Effect of DPN concentration on Hg0 removal The variation of removal efficiency for Hg0 removal as a function of the concentration of DPN is shown in Fig. 2. It can be seen from Fig. 2 that the removal efficiency enhances fast when the concentration of DPN is between 2 × 10−4 and 5 × 10−4 mol/L, however, the change occurs at 5 × 10−4 mol/L. The removal efficiency increases very slowly in DPN concentration ranging from 5 × 10−4 to 6 × 10−4 mol/L, because the total rate constant decrease when the DPN concentration is above 5 × 10−4 mol/L. Similar trends were obtained from Mulla and Nandibewoor’s research [25]. From the economy, the optimal DPN concentration was selected as 5 × 10−4 mol/L on Hg0 removal alone. Previous experiments carried out by Hiremath, Mulla and Liu [21,25,26,29] indicated that [Ni(H3 IO6 )2 (OH)2 ]2− was considered to be the active species and the reactions responding to [Ni(H3 IO6 )2 (OH)2 ]2− were the pseudo first order and double electron transfer mechanism. These results are benefit to establishing the optimal experimental conditions and indicating the reaction pathway.
The experiments for Hg0 removal were carried out at different pH ranging from 7.5 to 12.0. The results are shown in Fig. 4. It is observed from Fig. 4 that the Hg0 removal efficiency increase when pH is between 7.5 and 8.5, thereafter, the removal efficiencies of Hg0 decrease rapidly, which might be due to the change
3.2. Effect of reaction temperature on Hg0 removal Hg0 removal efficiency was measured at different reaction temperatures with keeping other conditions constant (Fig. 3). The removal efficiency of Hg0 increased slowly with reaction
Fig. 4. Effects of pH on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 5 × 10−4 mol/L; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
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Fig. 5. Effects of the initial Hg0 concentration on Hg0 removal. DPN concentration, 5 × 10−4 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
of the existing form of DPN under the different pH. It can be seen from Fig. 9 that the concentration of stronger power species, [Ni(H3 IO6 )2 (OH)2 ]2− achieves the highest when the pH is from 8.5 to 10.5, and decreases the lowest at pH 12. In this case, the week oxidizing species, [Ni(H2 IO6 )2 (OH)2 ]4− , is the main existing form of DPN, the previous evidences obtained from Mulla and Nandibewoor confirm this finding [26,29]. Therefore, there is a decrease in removal efficiency at pH 12. According to the results of experiments, the optimum solution pH was selected as 8.5.
Fig. 6. Effects of SO2 concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
HgSO3 + SO3 2− ↔ Hg(SO3 )2− 2
(2)
However, the removal efficiency decreased sharply when SO2 concentration increased continuously to 3590 mg/m3 . It meant that the higher concentration SO2 seemed to go against the removal of Hg0 , which might be due to the competing reaction of SO2 . In addition, during the reaction process, the removal efficiency of SO2 reached 98%.
3.4. Effect of the initial Hg0 concentration on Hg0 removal
3.6. Effects of NO concentration on Hg0 removal
The data of actual coal-fired power plants combustion processes indicates that the concentration of Hg0 varies with the content of mercury in burned coal and operating conditions of boilers. Experiments of Hg0 removal under different initial Hg0 concentrations were carried out to examine the adaptability of DPN for different coal species and operating conditions of boilers, as shown in Fig. 5. The removal efficiencies vary slightly when the concentrations of Hg0 ranges from 20 to 40 g/m3 , indicating that the method proposed adapts to different types of coal and combustion conditions when the removal efficiencies demand in between 74.1 and 69.2%.
As shown in Fig. 7, NO has a slight promoting effect for removing Hg0 efficiency from 74.1% to 78.2% when NO concentration increases from 0 to 413 mg/m3 . The improvement for Hg0 removal can be explained by that, in the absence of buffer solution, with an increase of NO concentration, the solution pH raised because the oxidation of NO by DPN solution was a reaction of releasing OH− [31], which was helpful to form the strong oxidizing species, [Ni(H3 IO6 )2 (OH)2 ]2− and to promote the removal efficiency. On the contrary, the removal efficiency of Hg0 fell off with a continued increase of NO concentration, which might be due to the competing reaction of NO. Under this experimental condition, the removal efficiency of NO was about 53.8%.
3.5. Effects of SO2 concentration on Hg0 removal According to the conclusion of Section 3.1, DPN concentration was determined to be 0.5 mmol/L on Hg0 removal alone. However, 0.5 mmol/L DPN was difficult to removal NO and SO2 because DPN was consumed quickly when NO and SO2 were added in simulated flue gas. Thus, the optimal DPN concentration for simultaneous removing NO, SO2 and Hg0 was investigated. According to the experimental results, DPN concentration used in Sections 3.5, 3.6 and 3.7 was selected as 6 × 10−3 mol/L. Fig. 6 shows the change of the removal efficiency of Hg0 with the SO2 concentration, from Fig. 6, one finds that the SO2 concentration has a dramatic effect on Hg0 removal. The removal efficiency of Hg0 increases from 74.1 to 83.5% when SO2 concentration increases from 0 to 1258 mg/m3 . The promoting action of SO2 for removing Hg0 may also be attribute to the complex of Hg·S(IV) [30]. In Eqs. (1) and (2), the removal product, Hg2+ reacted with SO3 2− to form HgSO3 , and then sequentially reacted with SO3 2− to form Hg(SO3 )2 2− which was more stable than HgSO3 . Therefore, the removal efficiency of Hg0 could be improved by an increase of SO2 concentration. Hg2+ + SO3 2− ↔ HgSO3
(1)
Fig. 7. Effects of NO concentration on Hg0 removal. Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min.
Y. Zhao et al. / Journal of Hazardous Materials 260 (2013) 383–388
Fig. 9. Concentration ratios of IO4 − and its dissociation products in [IO4 − ]ex .
Fig. 8. Experiments on simultaneous removal of SO2 , NO and Hg0 . Hg0 concentration, 20 g/m3 ; DPN concentration, 6 × 10−3 mol/L; pH, 8.5; reaction temperature, 50 ◦ C; stabilization absorption time, 60 min; SO2 concentration, 1258 mg/m3 ; NO concentration, 413 mg/m3 .
Table 1 Concentration of Hg2+ in the spent absorption solution.a
3.7. Experiments on simultaneous removal of SO2 , NO and Hg0 The experiments of removing SO2 , NO and Hg0 were carried out to verify the effect of simultaneous removal under the optimal conditions of Hg0 removal, in which reaction temperature was 50 ◦ C, DPN concentration was 6 mmol/L, solution pH was 8.5 and the initial Hg0 concentration was 20 g/m3 . As shown in Fig. 8, 86.2% of Hg0 , 56.2% of NO and 98% of SO2 are simultaneously removed from the simulated flue gas and the efficiencies are stable in the reaction time ranging from 10th to 60th minute. In addition, comparing Fig. 8 with Figs. 2–4 and Fig. 6, the removal efficiency of Hg0 increases obviously, which verifies further the promotion of SO2 and NO for Hg0 removal.
− 2− IO− 4 + OH + H2 O ⇔ H3 IO6 , − 3− IO− , 4 + 2OH ⇔ H2 IO6
(3)
ˇ2 = 1.62 × 106
(4)
ˇ3 = 4.68 × 108
2
2
− − [H3 IO6 2− ] = ˇ2 × [OH− ] × [IO− 4 ] = f2 ([OH ]) × [IO4 ] 2
− − − − [H2 IO3− 6 ] = ˇ3 × [OH ] × [IO4 ] = f3 ([OH ]) × [IO4 ]
[IO− 4] =
−(f2 ([OH− ]) + f3 ([OH− ]) + 1) +
3
4
5
Average
0 200
0 197
0 201
0 199
0 199
The concentration unit of the measurement values is ng/L.
2
f1 ([OH− ]) = ˇ1 × [OH− ] , −
− 2
f2 ([OH− ]) = ˇ2 × [OH− ],
[IO4 − ]ex
[Ni(H3 IO6 )2 (OH)2 ]2− + OH− ⇔ [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− +H2 O (11) In order to reveal the reaction processes of Hg0 removal, the concentrations of Hg2+ in the spent absorption solutions were measured, and the results are shown in Table 1 in which scan blank was taken from fresh DPN solution, and sample 1 was the spent absorption solution. The data of Table 1 means that the oxidation of Hg0 occurs during the reaction process.
(6) (7) (8) 2
(f2 ([OH− ]) + f3 ([OH− ]) + 1) + 4f1 ([OH− ]) × [IO− 4 ]ex 2f1 ([OH− ])
(10)
presents total concentration f3 ([OH ]) = ˇ3 × [OH ] and of periodate in solution. As shown in Fig. 9, the existence form of [IO4 − ]ex is significantly affected by solution pH. When the concentration of [IO4 − ]ex was 5 × 10−4 mol/L, the concentration ratios of IO4 − and its hydrolyzing products were calculated at different pH. When solution pH varies from 7.5 to 8.5, the main species of [IO4 − ]ex are IO4 − and H3 IO6 2− ; when solution pH varies from 8.5 to 11.0, that is H3 IO6 2− ; when solution pH varies from 11.0 to 12.0, H3 IO6 2− and H2 IO6 3− are the main existence form. So the main existence form of DPN should be [Ni(H3 IO6 )2 (OH)2 ]2− under the optimum solution pH (at 8.5), and the others were [Ni(H2 IO6 )(H3 IO6 )(OH)2 ]3− and H2 I2 O10 4− , the latter may be ignored [34] due to its very low concentration ratio. These results are consistent with the researches of Yang [35] and Chimatadar [19]. In the kinetic study of DPN by Harihar [20], the conclusion of [Ni(H3 IO6 )2 (OH)2 ]2− as the main reactive form of DPN was also proposed. Furthermore, it was reported that part of [Ni(H3 IO6 )2 (OH)2 ]2− could be changed into stronger power species, [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− , which was supported by the Hiremath’s studies on the observed fractional order in [OH− ] [19–21]:
The expressions of the concentration of IO4 − and its hydrolyzing products are shown in Eqs. (6)–(9), which are obtained from Eqs. (3)–(5): 2
2
0 198
where
(5)
− − [H2 I2 O10 4− ] = ˇ1 × [OH− ] × [IO− 4 ] = f1 ([OH ]) × [IO4 ]
1
Scan blank Sample 1
4− [IO− ] + [H3 IO6 2− ] + [H2 IO6 3− ] + [IO− 4 ]ex = [H2 I2 O10 4]
A key component, IO4− in the absorbent can be decomposed to a variety of species, which are closely related to the existing form of DPN. For the speculation of the removal mechanism of Hg0 , it is important to investigate the hydrolyzing products of IO4− . According to previous research results, the water soluble Ni(IV) periodate complex was reported to be [Ni(HIO6 )2 (OH)2 ]6− [32]. However, in an alkaline medium and at the pH used in our study, the dissociation reactions of IO4− and their dissociation constant in alkaline medium at 25 ◦ C are summarized as follows [33]: ˇ1 = 1.12 × 1015
Determination time
a
3.8. Removal mechanism discussion
− 4− 2IO− , 4 + 2OH ⇔ H2 I2 O10
387
(9)
388
Y. Zhao et al. / Journal of Hazardous Materials 260 (2013) 383–388
Based on the research results of the hydrolyzing products of IO4 − , and the analysis of the removal products of Hg0 , the reaction mechanism of Hg0 removal is inferred as follows: Hg(g) → Hg(l)
(12)
[Ni(H3 IO6 )2 (OH)2 ]2− + OH− ⇔ [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− +H2 O (13)
[Ni(H3 IO6 )2 (OH)2 ]2− + Hg(l) → [Ni(H3 IO6 )2 (OH)2 ]4− + Hg2+ (14) [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− + Hg(l) → [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]5− + Hg2+
(15)
4. Conclusions The factors influencing the removal of Hg0 from simulated flue gas by DPN, namely DPN concentration, reaction temperature, solution pH, initial Hg0 concentration, SO2 and NO concentration, are described for the first time in this study. The experimental results indicated that DPN had higher removal efficiency of Hg0 under the optimum experimental conditions. The reaction mechanism of Hg0 removal was proposed, in which [Ni(H3 IO6 )2 (OH)2 ]2− was the main existing and active form under the optimum experimental condition, and a partial [Ni(H3 IO6 )2 (OH)2 ]2− could be transformed to stronger power species, [Ni(H3 IO6 )(H2 IO6 )(OH)2 ]3− . For Hg0 removal, it was attributed to the oxidation of Hg0 by these two species to Hg2+ . Acknowledgments We appreciate the financial supports of National Hightech Research and Development Projects (863 program, No. 2013AA65403), Program for Changjiang Scholars and Innovative Research Team in University (IRT1127) and Zhejiang Provincial Engineering Research Center of Industrial Boiler & Furnace Flue Gas Pollution Control, Hangzhou, PR China (311202). References [1] R. Yan, D.T. Liang, L. Tsen, Bench-scale experimental evaluation of carbon performance on mercury vapor adsorption, Fuel 83 (2004) 2401–2409. [2] S.H. Lee, Y.J. Rhim, S.P. Cho, Carbon-based novel sorbent for removing gas-phase mercury, Fuel 85 (2006) 219–226. [3] T.G. Lee, P. Biswas, E. Hedrick, Comparison of Hg0 capture efficiencies of three in situ generated sorbents, Aiche J. 47 (2001) 954–961. [4] Z. Barnea, T. Sachs, M. Chidambaram, Y. Sasson, A novel oxidative method for the absorption of Hg0 from flue gas of coal fired power plants using task specific ionic liquid scrubber, J. Hazard. Mater. 244–245 (2013) 495–500. [5] J. Wilcox, E. Rupp, S.C. Ying, D.H. Lim, A.S. Negreira, A. Kirchofer, F. Feng, K. Lee, Mercury adsorption and oxidation in coal combustion and gasification processes, Int. J. Coal Geol. 90–91 (2012) 4–20. [6] International Joint Commission and the Commission for Environmental Cooperation, Consultation on Emissions from Coal-Fired Electrical Utilities, Quebec, Canada, 2004, pp. 20–21. [7] National Bureau of Statistics of China, China Statistical Yearbook, Chinese Statistic Press, Beijing, 2007, pp. 261–265. [8] Y.J. Wang, N.F. Duan, L.G.S. Yang, L. Meng, Z.J. Huang, Experimental study on mercury removal by combined wet flue gas desulphurization with electrostatic precipitator, P. Chinese Soc. Electr. Eng. 28 (2008) 64–69 (in Chinese).
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