Integrative Process for Simultaneous Removal of SO2 and NO

Sep 8, 2014 - Basic rules of NO oxidation by Fe 2+ /H 2 O 2 /AA directional decomposition system. Haiqian Zhao , Xingcun Gao , Chenghao Liu , Zhonghua...
1 downloads 20 Views 6MB Size
Download PDF
Article pubs.acs.org/EF

Integrative Process for Simultaneous Removal of SO2 and NO Utilizing a Vaporized H2O2/Na2S2O8 Yi Zhao,* Runlong Hao, Peng Zhang, and Sihan Zhou School of Environmental Science & Engineering, North China Electric Power University, 071003 Baoding, Hebei Province People’s Republic of China ABSTRACT: A novel integrative process for simultaneous removal of SO2 and NO from coal-fired flue gas was designed, in which, SO2 and NO were initially oxidized by a vaporized complex oxidant composed of hydrogen peroxide (H2O2) and sodium persulfate (Na2S2O8) (HN solution) and then absorbed by the Ca(OH)2 that followed. The effects of various reaction factors on the simultaneous removal were investigated, such as the concentration of Na2S2O8, the transition metal additives, the HN pH, the adding rate of HN, the residence time of flue gas, the reaction temperature, and the concentrations of coexistence gases O2, SO2, NO, and CO2. Under the optimal conditions, the simultaneous removal efficiencies of 98.9% for SO2 and 79.6% for NO were obtained. The reaction mechanism was speculated based on the characterization of removal products by scanning electron microscopy and X-ray diffraction and related literature references. Meanwhile, the macrokinetics of desulfurization and denitrification were also calculated.

1. INTRODUCTION The removal of SO2 and NOx, especially the insoluble NO, mainly emitted from coal-fired power plants, has been the subject of many studies in recent years because of many recent environmental problems that these pollutants have generated. As the largest coal-fired country in the world, China’s government has issued “Emission Standard of Air Pollutants for Thermal Power Plants” (GB13223-2011) to reduce NOx and SO2 emissions. In China, the most common technology for flue gas desulfurization (FGD) is the wet scrubbing technique, in which the lime/limestone process is a dominating technology. Among the flue gas denitrification technologies, the selective catalytic reduction (SCR) and selective noncatalytic reduction (SNCR) processes are frequently used, especially the SCR.1−4 Although this SCR-WFGD stage treatment technology succeeds in combined removal of SO2 and NO, it has the disadvantages of large occupying area and high running cost,5,6 so the development of new technologies and equipment of simultaneous removal has become the leading research direction in the air pollution control field,7 in which some advanced technologies such as electrochemistry,8,9 gas−solid phase adsorption,10,11 gas−solid phase catalysis,12,13 and liquid phase absorption14,15 were attempted for simultaneous removal of SO2 and NO, but most of them had technical or economic defects and cannot be developed to practicable technologies. The oxidation method is a promising way to simultaneously remove SO2 and NO, the core of which is to rapidly oxidize the insoluble NO, and a combined system of oxidizing NO to NO2 by vaporized oxidants and then absorbed by Ca(OH)2 seems to be an alternative option. The possible NO oxidants include chlorine-containing oxidants,7,16−18 ferrate(VI),19 manganese oxidants,20 ozone,21 and hydrogen peroxide,22−24 etc., but the high costs of NaClO2, ferrate(VI), and ozone (O3) and the secondary environmental problems owing to the toxic byproducts of manganese and chlorine generated in the disposal processes inhibit their applications. From the view of cost and © 2014 American Chemical Society

environmental impact, H2O2 is the most suitable reagent that can be employed to develop a promising and practical approach for simultaneous removal of SO2 and NO, but the key challenge is to synthesize a H2O2-based oxidant with a stronger oxidizability. It has been reported that persulfate (PS) and H2O2 have a synergy effect, from which, the active radicals of SO4• (2.5−3.1 V) and HO• (2.8 V) can be generated. In addition, PS has been used in the areas of total organic carbon (TOC) degradation and in situ chemical oxidation (ISCO) of subsurface contamination for a long time.25−35 Thus, we selected Na2S2O8 as an additive for H2O2 to prepare a H2O2-based complex oxidant (HN solution) to simultaneously remove SO2 and NO. To our knowledge, there were no reports in the field of simultaneous removal of SO2 and NO on the usage of Na2S2O8/H2O2. Moreover, for the purpose of water saving, a semidry integrative process for flue gas cleaning was also designed, in which SO2 and NO were initially oxidized by the vaporized HN and then absorbed by the Ca(OH)2 that followed . The effects of various reaction factors on the simultaneous removal were investigated experimentally, such as the concentration of PS, the transition metal additives, the HN pH, the HN adding rate, the residence time of flue gas, the reaction temperature, and the concentrations of coexistence gases. The reaction mechanism and macrokinetics of desulfurization and denitrification were also proposed at last.

2. EXPERIMENTAL SECTION 2.1. Reagents and Preparation of HN Solution. All reagents used were analytical reagents (Kermel Co., Tianjin, China). H2O2 of 30% (w/w) and Na2S2O8 of 98% (w/w) were used to prepare HN solution. Ca(OH)2 and anhydrous CaCl2 were employed as the absorbent and the dryer, respectively. The preparation order of HN Received: July 28, 2014 Revised: September 5, 2014 Published: September 8, 2014 6502

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

Figure 1. Schematic diagram of the experimental apparatus of the fixed bed: 1−5, CO2, N2, SO2, NO, and O2 gas cylinders; 6, flowmeters; 7, buffer bottle; 8, tee joint; 9, vaporization device; 10, thermal control electric heater; 11, HN solution; 12, peristaltic pump; 13, thermal couple; 14, reactor; 15, Ca(OH)2 powders; 16, thermostat oil bath; 17, dry tower; 18, flue gas analyzer. solution was that the fresh solutions of Na2S2O8 and H2O2 were added in a beaker in turn by using a pipet (10−1000 μL and 1−5 mL) and then shaken mildly. The pH of the HN solution was adjusted by H2SO4 solution of 1 mol/L and NaOH solution of 1 mol/L. 2.2. Experimental Apparatus and Procedures. The bench scale experiments mainly include the simulated flue gas generation, the HN vaporization, the integration of oxidation and absorption, and the tail gas detection, as shown in Figure 1. The simulated flue gas was generated from N2, SO2, NO, O2, and CO2 provided by the compressed cylinders (1−5; North Special Gas Co., Baoding, China). A peristaltic pump (12; BT100-1F, Longerpump, Baoding, China) was used to pump the HN solution (11) in the vaporization device (9) that was heated by a thermal control electric heater (10; ZDHW, Zhongxingweiye Co., Beijing). The inner temperature of the vaporization device was detected over time by a thermal couple (13; XMTD, Baoding) and in accordance with that of the reactor. The reactor was an U-type quartz tube (14) with a length of 30 cm and an inner diameter of 2.5 cm, heated by a thermostat oil bath (16; DC-RB, Duchuang Technology Co., Beijing). The tail gas was detected by a flue gas analyzer (18) (ECOM-J2KN, RBR Co., Isenlohn, Germany). During the experiments, the heater devices 10 and 16 were turned on to heat the relevant devices to the desired temperatures. Afterward, SO2 and NO and other gases were metered through the mass flow controllers (6) and then diluted by N2 to their desired concentrations, from which, the simulated flue gas was formed. Thereafter, the HN solution (11) was pumped by the peristaltic pump (12) and carried simultaneously by N2 to the vaporization device (9), where it was vaporized immediately. Then the oxidization and absorption reactions were carried out in the reactor (14, 15) that was filled with Ca(OH)2 powders supported on glass wool. The removal efficiencies of SO2 and NO were calculated based on the different values of the inlet and outlet concentrations of SO2 and NO. 2.3. Analytical Methods. The solution pH was tested by pH meter (PHS-3C, Youke, Shanghai). An X-ray diffraction (XRD, D8 ADVANCE type, BRUKER-AXS, Karlsruhe, Germany; 40 kV and 20 mA) was used to determine the components in fresh and spent Ca(OH)2, the scanning range was from 20 to 70° with a scanning velocity of 7° min−1. The surface patterns of fresh and spent Ca(OH)2 were observed using a scanning electron microscope (SEM, JEOL JSM-7500F, Tokyo, Japan).

Figure 2. Effect of Na2S2O8 concentration on simultaneous removal efficiencies. Simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3.

increases sharply from 52% to 76% with the concentration of Na2S2O8 varying from 0 to 200 mmol/L; afterward, the efficiency reaches a plateau. Therefore, in consideration of economy, the optimal concentration of Na2S2O8 in HN was determined as 200 mmol/L. The satisfactory removal of SO2 was attributed to the enhanced oxidation by vaporized HN and the effective absorption of Ca(OH)2. For the denitrification, the increase of efficiency was mainly due to the synergism of PS and H2O2, in which, SO4•− was generated by breaking the peroxide bond (O−O) of PS (eq 1) through the heat/H2O2 activations.30,33 In addition to that, HO• and HSO5− produced (eqs 2−4)32 via the reactions between SO4•− and H2O/H2O2 were also partly contributed to NO oxidation.30

3. RESULTS AND DISCUSSION 3.1. Effects of Various Additives. The mass concentration of H2O2 used in the tests was 30% (w/w), and the corresponding molar concentration was approximately 9.0 mol/L. Figure 2 describes the effect of Na2S2O8 concentration in HN on the simultaneous removal. It can be seen that the desulfurization efficiency is constant at 99% whatever the Na2S2O8 concentration. But for the denitrification, the addition of Na2S2O8 has an evident promotion, its removal efficiency

heat/H 2O2

S2 O82 − ⎯⎯⎯⎯⎯⎯⎯⎯⎯⎯→ 2SO4•− SO4

6503

•−

•

+ H 2O2 → HO +

(1)

HSO5−

(2)

SO4•− + H 2O ⇔ HO• + H+ + SO4 2 −

(3)

SO4•− + HO• → HSO5−

(4)

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

It has been known from the references that both of PS and H2O2 can be activated by the transition metals.34,35 For example, Liang et al.29,30 carried out the experiment of degradation of the organic pollutants by SO4•− generated by iron catalyzing PS, and Kolthoff et al.34 also demonstrated the cuprous had a catalytic effect on PS. Therefore, in order to enhance the denitrification efficiency, the effects of different transition metals on denitrification efficiency were investigated. As shown in Figure 3, the removal efficiencies are declined

Figure 4. Effect of adding rate of HN solution on simultaneous removal efficiencies. Simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, HN solution pH is 5.5, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3.

more HN solution addition. For NO, as the adding rate increased, the molar ratio of vaporized oxidants to NO was increased, which led to a promotion on NO oxidation. Nevertheless, as the adding rate further increased, the promotion was inhibited to some degree. This was due to reactants decompositions (eq 5) and radicals quenching (eqs 67) accelerating when more vaporized oxidants were present in the simulated flue gas.22,23,35 Furthermore, Figure 4 also shows that the molar ratio of vaporized oxidants to SO2 and NO is in a range of 1−6 when the adding rate varied from 10 to 60 μL/min. And with respect to the optimal adding rate of 30 μL/min, the corresponding molar ratio was 3.

Figure 3. Effects of transition metals on denitrification efficiency. Simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, reaction temperature is 383 K, and NO concentration is 550 mg/m3.

when Fe2+, Cu2+, and Mn2+ are added, which are inconsistent with the previous demonstrated conclusions.29,30,34,35 Known from the literature references,29,30,34,35 the transition metals can catalyze PS to generate SO4•−, but at the same time they can consume SO4•−, which depends on the character of the reaction system. In this experiment, as a consequence of the combination effect of H2O2 activation and heat catalysis, PS could effectively and completely produce SO4•−; thus the promotion on PS resulting from transition metal could be neglected. In contrast, the transition metal could adversely consume the just generated radicals to decrease the NO removal efficiency. 3.2. Effect of Adding Rate of HN Solution. Obviously, the rate and extent of an oxidation reaction depend on the oxidants’ concentrations that rely on the HN adding rate in our research. Therefore, the effects of the adding rate of HN on simultaneous removal were studied experimentally. It can be found in Figure 4 that the simultaneous removal efficiencies significantly increase from 89% to 99% for SO2 and from 40% to 76% for NO with the adding rate varying from 10 to 30 μL/ min, and then the simultaneous removal efficiencies are stable at 99% and 77%, respectively; in an adding rate range of 40−50 μL/min, after 60 μL/min, the simultaneous removal efficiencies declined slowly. Hence, from the consideration of economy, the optimal adding rate of HN was selected as 30 μL/min. For SO2, the increase of efficiency in the low adding rate range was due to the enhancement of SO2 oxidation and the increase of the moisture content on the surface of Ca(OH)2 resulting from

S2 O82 − + H 2O2 → 2H+ + 2SO4 2 − + O2

(5)

SO4•− + SO4•− → S2 O82 −

(6)

HO• + HO• → H 2O2

(7)

3.3. Effect of HN Solution pH. According to refs 28, 30, 33, and 36 we learned that the solution pH significantly affected the oxidation potentials and the speciation of H2O2 and PS. Therefore, experiments over a wide pH ranging from 2 to 7 were conducted. As shown in Figure 5, the simultaneous removal efficiencies are stable at 97% for SO2 and 76% for NO when the pH increases from 2 to 5.5. However, as the pH further increases from 5.5 to 7, the simultaneous removal efficiencies decline linearly to 92% and 60%, respectively. Therefore, the optimal HN pH was selected as 5.5. The followed reasons can account for the preceding phenomenon. First, H2O2 was decomposed as HO2− in the higher pH (eq 8),37,38 and then HO2− further promoted H2O2 decomposition (eq 9).39 Thus, the elevated pH was unfavorable to H2O2 existence. Secondary, the oxidizing ability of PS was reduced by increasing pH, which was due to SO4·− gradually transforming to •OH when the solution pH was over 5.5. Although •OH is an effective oxidant, the •OH-induced oxidation reaction is unselective. What’s worse, the •OH may be consumed by other coexisting species in HN.31,33 Thus, the 6504

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

flue gas temperatures of the inlet FGD-CFB are 120−170 °C and that of the outlet is about 70 °C.17 So a series of experiments of temperature dependence were carried out accordingly. It can be seen from Figure 6 that the

Figure 5. Effect of HN solution pH on simultaneous removal efficiencies. The simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3. Figure 6. Effect of reaction temperature on simultaneous removal efficiencies. The simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3.

weak acidic environmental condition was more favorable to NO oxidation, which was in accordance with the previously established conclusions; i.e., Chu et al.40 had investigated the synergy of SO42− and •OH in UV/S2O82−/H2O2 oxidizing iopromide, the results indicated that the highest decay rate was achieved at pH 4.34 and decreased with pH increasing. Another conclusion of PS oxidizing 1,1,1-trichloroethane (1,1,1-TCA) fast and thoroughly with the solution pH being 3 was drawn by Gu et al.,31 who also pointed out the acidic condition could catalyze SO4•− generation (eqs 10 and 11), and the process would be inhibited by the increasing pH such that when the solution pH increased from 7 to 7.5, the oxidation rate of TCA by PS decreased from 80.3% to 47.8%. Additionally, the influence of solution pH on PS oxidizing methyl-tert-butyl-ether (MTBE) was also investigated by Huang et al.,33 in which the phenomenon of oxidation rate decreasing sharply with pH increasing was observed. However, some other researchers had proposed the opposite views; for example, Block et al.36 found that the strong alkaline (pH over 13) could activate PS (eq 12) and , importantly, the activation process not only depended on the strong alkaline environment but also relied on the molar ratio of the pH modifier (such as KOH) and PS, inferring that the excess OH− was beneficial to PS activation, but not a little OH−. H 2O2 → H+ + HO2− −

desulfurization efficiency is over 98% constantly when the reaction temperature increases from 70 to 140 °C, and the highest efficiency of 99.5% is obtained at 110 °C. For NO, the efficiency increases sharply from 56.2% to 76% when the reaction temperature increases from 70 to 110 °C, after 120 °C, it declines slightly. Therefore, 110 °C could be considered as the optimal temperature. Apparently, the elevated temperature in the low-temperature range had a significant promotion on the simultaneous removal, which was due to that the evaporation rate of HN, the diffusions of oxidants, and oxidation products, and the chemical reaction rates between oxidants and SO2/NO were accelerated in the process. However, when the temperature was over 120 °C, H2O2 decomposition would be aggravated, which then would result in a slight decrease of NO removal. The phenomenon has been demonstrated in our previous work.41 3.5. Effect of Flue Gas Residence Time. It is important to investigate the effect of flue gas residence time because the contact time between oxidants and air pollutants has a great influence on the oxidation reactions, so the experiments with various flue gas residence times were carried out. As shown in Figure 7, the desulfurization efficiency is stable at 99% when the flue gas velocity is within 1.5 L/min; the corresponding residence time is less than 5.7 s, but it decreases linearly from 99% to 88% when the gas velocity increases from 1.5 to 3.0 L/ min; the corresponding residence time is in the range of 3.8− 5.7 s. A similar trend of denitrification changing with residence time is also observed: the efficiency is over 76% when the gas velocity is within 2.0 L/min; the corresponding residence time is less than 4.8 s; thereafter, the efficiency declines significantly. The reasons for the decreases were that the increase of the gas velocity at a constant adding rate of HN would decrease the oxidants concentration in the simulated flue gas and increase

(8) −

H 2O2 + HO2 → OH + O2 + H 2O

(9)

S2 O82 − + H+ → HS2 O8−

(10)

HS2 O8− → H+ + SO4•− + SO4 2 −

(11)

SO4•− + OH− → HO• + SO4 2 −

(12)

3.4. Effect of Reaction Temperature. In a typical coalfired power plant, the flue gas temperatures in the tail of the power boiler are approximately 800−900 °C; after passing the economizer, air preheater, and electrostatic precipitation, the 6505

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

velocity would decrease the contact time between SO2/NO and oxidants, leading to a decrease of oxidation efficiency. However, the excessive residence time of the flue gas will increase the volume capacity of the reactor and the running cost.42 Therefore, the optimal residence time was selected as 4.8 s with respect to the flue gas velocity of 2.0 L/min. 3.6. Effects of Coexistence Gases in Flue Gas. The experiments of simultaneous removal in the presence of coexistence gases of coal-fired flue gas were carried out. Parts a and b of Figure 8 show the effects of CO2 and O2: the simultaneous removal of SO 2 and NO were affected insignificantly by the variations of CO2 and O2 concentrations. Figure 8c describes the dependence of desulfurization on SO2 itself; it can be found that the variation of SO2 had a negligible effect. The reason was that Ca(OH)2, as an effective absorbent for SOx, can adapt various SO2 concentrations. However, the influence of SO2 on NO removal was either promotion or inhibition, depending on its concentration. When SO 2 concentration is in a range of 700−2400 mg/m3, the denitrification efficiency slightly increases from 76% to 81%, but as the SO2 concentration further increases, it declined slowly. The reason for the promotion was that SO3•− produced via the reactions of HSO3− and SO4•− and HO• (eqs 13−15) may enhance the NO oxidation.43 The inhibition caused by higher SO 2 concentration was a consequence of the competition reaction between NO and SO2 with the limited oxidants. Figure 8d shows the effect of NO: the desulfurization efficiency is constant at 99% when NO ranges from 300 to 1100 mg/m3; however, the denitrification efficiency declines

Figure 7. Effect of flue gas residence time on simultaneous removal efficiencies. The adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3.

the treating amount of SO2/NO through the reactor per unit time, leading to a decrease of the molar ratio of vaporized oxidants to SO2/NO. On the other hand, from the view of reaction kinetics, the oxidation reaction was significantly affected by the flue gas residence time; namely, the decline of flue gas residence time resulting from the increase of flue gas

Figure 8. Effects of coexistence gases in flue gas on simultaneous removal efficiencies. The simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3. 6506

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

from 78% to 64% when NO is between 500 and 1100 mg/m3, indicating that the increase of NO concentration has an evident inhibition on NO removal, but not for SO2 removal. According to the actual running conditions of coal-fired power plants in China, the concentration of NOx in flue gas is no more than 500 mg/m3 if the low NOx burner is equipped in the boiler, so the denitrification efficiency of 80% can be obtained. Therefore, the proposed process can not only meet the NOx emission standard for Chinese coal-fired power plants (lower than 100 mg/m3) but also replace the SCR system avoiding the NH3 escape and realize the goal of simultaneous removal of SO2 and NO. SO2 + H 2O → HSO3− + H+

(13)

SO4•− + HSO3− → HSO4 − + SO3•−

(14)

HO• + HSO3− → H 2O + SO3•−

(15)

3.7. Parallel Tests of Simultaneous Removal of SO2 and NO. Based on the above experiments, the optimal preparation conditions of HN and the reaction conditions were determined, in which the concentration ratio of H2O2 to Na2S2O8 was 9.0:200, the HN pH was 5.5, the adding rate of HN was 30 μL/min, the flue gas residence time was 4.8 s corresponding to the inlet velocity of 2.0 L/min, and the reaction temperature was 110 °C. Under these conditions, the average simultaneous removal efficiencies were 98.9% for SO2 and 79.6% for NO when the concentrations of SO2 and NO were 550 and 2800 mg/m 3 , respectively. The good reproducibility of the experimental data shown in Table 1 Table 1. Parallel Tests of Simultaneous Desulfurization and Denitrification SO2 NO

1

2

3

4

5

av

std err

98.2 79.2

99.1 80.3

99 80.7

98.6 79.1

99.4 78.5

98.9 79.6

0.42 0.81

indicates that the integrative process exhibits a stable and good performance on simultaneous removal of SO2 and NO, which can be as a beneficial reference for its application in industry. From the view of economy, it is estimated that the apparent cost of this integrative way is 7000 ¥/t, which is lower than that of the combination of WFGD−SCR, 10000−13000 ¥/t, in addition to the novel method having the following superiorities, such as the simplified system and devices, the lower operation and capital construction costs, and the smaller occupying area.19

Figure 9. SEM patterns of the spent Ca(OH)2 (II) and fresh Ca(OH)2 (I).

4. REACTION MECHANISM AND MACROKINETICS 4.1. Reaction Mechanism of Simultaneous Removal of SO2 and NO. Figure 9(I) shows the smooth surface of the fresh Ca(OH)2, by contrast, some new small clusters and grooves are observed in Figure 9(II), which may be due to the removal products such as CaSO4, CaSO3, Ca(NO3)2, and Ca(NO2)2 generating on the surface of Ca(OH)2. In order to verify the deduction, the removal products were analyzed by Xray diffractions. It can be seen from Figure 10 that, compared with the fresh sample, many new diffraction peaks appear in the XRD pattern of the spent absorbent, indicating that the complex reactions among SOx, NOx, Ca(OH)2, and vaporized HN occur. It also can be found that the characteristic peaks of CaSO4 and Ca(NO2)2 are a little stronger than those of CaSO3

Figure 10. XRD patterns of the spent Ca(OH)2 (II) and fresh Ca(OH)2 (I). 6507

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

Figure 11. Reaction orders of desulfurization and denitrification. The simulated flue gas inlet velocity is 2.0 L/min, residence time is 4.8 s, adding rate of HN solution is 30 μL/min, HN solution pH is 5.5, reaction temperature is 383 K, NO concentration is 550 mg/m3, and SO2 concentration is 1700 mg/m3.

and Ca(NO3)2, demonstrating that the major removal products are CaSO4 and Ca(NO2)2, and the minor ones are CaSO3 and Ca(NO3)2. According to the characterization of removal product and the related literature references,30−41,43 the reaction mechanism of simultaneous removal of SO2 and NO by vaporized HN was concluded as follows. First, the NO and SO2 were oxidized by the vaporized oxidants such as H2O2, PS, SO4•−, OH•, and HSO5− (eqs 16−24), of which the feasibility was demonstrated from the electrochemical perspective: the standard electrode potentials of H2O2 (1.770 V), PS (2.01 V), SO4•− (2.5−3.1 V), and •OH (2.80 V) are far higher than those of SO42−/H2SO3 (0.172 V), SO42−/SO2 (0.158 V), NO2/NO (1.049 V), NO3−/ NO (0.957 V), NO2−/NO (−0.460 V), and NO3−/NO2− (0.835 V).

Ca(OH)2 + 2N(V) → Ca(NO3)2 + H 2O

(28)

4.2. Macrokinetics of Desulfurization and Denitrification. 4.2.1. Reaction Orders of Desulfurization and Denitrification. The reaction orders of desulfurization and denitrification were calculated by using the initial rate method. First of all, the initial reaction rates of SO2 and NO depletions under various SO2 and NO concentrations were calculated (eqs 29 and 30). Then the reaction orders were obtained by using the differential form of the reaction rate (eqs 31 and 32). As shown in Figure 11a,b, the slopes are the reaction orders. After fitting, the reaction order of SO2 was 1.03216 (R2 was 0.97639) and that of NO was 0.93055 (R2 was 0.99474), which indicated that both of them were confirmed to the pseudo-first-order kinetics pattern in a 1−2 half-lifetime.32

H 2O2 + SO2 → H 2SO4

(16)

C = f (t )

H 2O2 + 2NO → 2H+ + 2NO2−

(17)

r0 = −

dC 0 df (t )t = 0 = dt dt

(30)

r0 = −

dC 0 = kC0 n dt

(31)

2S2 O82 − + SO2 + 2H 2O → 3SO4 2 − + 2SO4•− + 4H+ (18)

S2 O82 − + NO2− → SO4 2 − + SO4•− + NO2

(19)

SO4•− + NO2− → SO4 2 − + NO2

(20)

•

+

HO + NO → H + NO2 •

+

HO + NO2 → H + HSO5− HSO5−

−

NO3− −

+ SO2 → HSO4 + SO3 −

+ NO → HSO4 + NO2

(29)

log( −dC0/dt ) = log k + n log C0

where C is the concentrations of SO2 or NO, mg/m3; t is the reaction time, s; r0 is the initial reaction rates of SO2 or NO, mg·m−3·s−1; C0 is the initial concentrations of SO2 or NO, mg/ m3; k is the reaction rate constant, m3−n·mg−n·s−1; and n is the reaction order. 4.2.1. Apparent Activation Energies of Desulfurization and Denitrification. Parts a and b of Figure 12 show the pseudo-first-order reaction rate constants of desulfurization and denitrification under different reaction temperatures. As the reaction temperature increased from 363 to 413 K, the reaction rate constants for SO2 are 0.01633, 0.01947, 0.02500, 0.02787, 0.03211, and 0.03806 s−1 and those of NO are 0.00811, 0.00916, 0.01386, 0.01574, 0.01731, and 0.02076 s −1 . Obviously, the ln kobs decreases linearly with 1/T, therefore, which is fitted to an Arrhenius model (eq 33).

(21) (22) (23) (24)

Afterward, the reaction intermediates as well as almost all oxidation products such as HNO3, HNO2, H2SO4, H2SO3, NO2, and SO3 were absorbed by the hydrous Ca(OH)2 powders (eqs 25−28). Ca(OH)2 + S(IV) → CaSO3 + H 2O

(25)

Ca(OH)2 + S(VI) → CaSO4 + H 2O

(26)

Ca(OH)2 + 2N(III)/N(IV) → Ca(NO2 )2 + H 2O

(27)

(32)

⎛ E ⎞ k = A exp⎜ − a ⎟ ⎝ RT ⎠ 6508

(33)

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

It can be seen from Figure 12c that the obtained apparent activation energies are 22.33 and 23.75 kJ/mol for SO2 and NO, respectively. By contrast, the apparent activation energies of SO2 removal and NO removal by using NaClO2 solution in previous research were 66.25 kJ/mol for SO2 and 42.5 kJ/mol for NO.44 Additionally, those of SO2 and NO using H2O2/ NaOH solution were 42.65 and 27.8 kJ/mol.45 Obviously, both of them were higher than those obtained in our research, which was due to the fact that the contact area among the reactants and the molecular kinetic energy were greatly enlarged in the vaporization process, resulting in a decrease of reaction barrier.

5. CONCLUSION A complex oxidant made up of H2O2 and Na2S2O8 was prepared and the optimal reaction conditions for simultaneous removal of SO2 and NO were established; thus a novel flue gas purification process was proposed. The highest simultaneous removal efficiencies of SO2 and NO were 98.9% and 79.6%, respectively. The reaction mechanism was speculated based on the characterization of removal product by SEM and XRD and the related literature references. The reaction orders for desulfurization and denitrification were determined as 1.03216 and 0.93055, and the corresponding apparent activation energies were 22.33 and 23.75 kJ/mol, respectively.

■

AUTHOR INFORMATION

Notes

The authors declare no competing financial interest.

■

ACKNOWLEDGMENTS We appreciate the financial support by a grant from the National High Technology Research and Development Program of China (863 Program, No. 2013AA065403), from Hebei provincial Natural Science Foundation, People’s Republic of China (Grant B2011502027), from the Program for Changjiang Scholars and Innovative Research Team in University (Grant IRT1127), and from the Zhejiang Provincial Engineering Research Center of Industrial Boiler & Furnace Flue Gas Pollution Control, Hangzhou, People’s Republic of China (Grant 311202).

■

(1) Shafiee, S.; Topal, E. Energy Policy 2008, 36, 775−786. (2) Srivastava, R. K.; Jozewicz, W. S. J. Air Waste Manage. Assoc. 2001, 51, 1676−1688. (3) Chien, T. W.; Chu, H. J. Hazard. Mater. 2000, 80, 43−57. (4) Wendt, J. O. L.; Linak, W. P.; Groff, P. W.; Srivastava, R. K. AIChE. J. 2001, 47, 2603−2617. (5) Amin, N. A. S.; Chong, C. M. Chem. Eng. J. 2005, 113, 13−25. (6) Wang, J. D.; Wu, C. Q.; Chen, J. M.; Zhang, H. J. Chem. Eng. J. 2006, 121, 45−49. (7) Hutson, N. D.; Krzyzynska, R.; Srivastava, R. K. Ind. Eng. Chem. Res. 2008, 47, 5825−5831. (8) Chung, S. J.; Pillai, K. C.; Moon, I. S. Sep. Purif. Technol. 2009, 65, 156−163. (9) Govindan, M.; Chung, S. J.; Moon, Il. S. Ind. Eng. Chem. Res. 2012, 51, 2697−2703. (10) Davini, P. Carbon. 2001, 39, 2173−2179. (11) Tang, Q.; Zhang, Z. G.; Zhu, W. P.; Cao, Z. D. Fuel 2005, 84, 461−465. (12) Krishna, K.; Makkee, M. Appl. Catal., B 2005, 62, 35−44. (13) Sang, M. J.; Sang, D. K. Ind. Eng. Chem. Res. 2000, 39, 1911− 1916.

Figure 12. Determination of apparent activation energies. (a) Reaction rate constants of desulfurization in different reaction temperatures. (b) Reaction rate constants of denitrification in different reaction temperatures. (c) Apparent activation energies of desulfurization and denitrification.

Then the log type was deduced as follows (eq 34): ln k = ln A −

Ea RT

REFERENCES

(34)

where A is the preexponential factor; Ea is the apparent activation energy, kJ/mol; R is the universal gas constant; and T is the absolute temperature, K. 6509

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510

Energy & Fuels

Article

(14) Wang, L.; Zhao, W. R.; Wu, Z. B. J. Chem. Eng. 2007, 132, 227− 232. (15) Long, X. L.; Xiao, W. D.; Yuan, W. K. J. Hazard. Mater. B 2005, 123, 210−216. (16) Chu, H.; Chien, T. W.; Twu, B. W. J. Hazard. Mater. B 2001, 84, 241−252. (17) Wei, J. C.; Luo, Y. B.; Yu, B. J. Ind. Eng. Chem. 2009, 15, 16−22. (18) Zhao, Y.; Guo, T. X.; Chen, Z. Y.; Du, Y. R. Chem. Eng. J. 2010, 160, 42−47. (19) Zhao, Y.; Han, Y. H.; Ma, T. Z.; Guo, T. X. Environ. Sci. Technol. 2011, 45, 4060−4065. (20) Brogren, C.; Karlsson, H. T.; Bjerle, I. Chem. Eng. Technol. 1997, 20, 396−402. (21) Wei, L. S.; Zhou, J. H.; Wang, Z. H.; Cen, K. F. Ozone: Sci. Eng. 2006, 29, 207−214. (22) Liu, Y. X.; Zhang, J.; Sheng, C. D.; Zhang, Y. C.; Zhao, L. Chem. Eng. J. 2010, 162, 1006−1011. (23) Liu, Y. X.; Zhang, J. Ind. Eng. Chem. Res. 2011, 50, 3836−3841. (24) Cooper, C. D.; Clausen, C. A.; Pettey, L. Environ. Eng. 2002, 128, 68−72. (25) Eaton, A. D.; Clesceri, L. S.; Rice, E. W.; Greenberg, A. E. Standard Methods for the Examination of Water and Wastewater; APHA, AWWA, and WEF: Washington, DC, USA, 2005. (26) Jed, C.; Gretell, O.; John, C.; Kurt, D. P. Environ. Sci. Technol. 2010, 44, 9445−9450. (27) Osgerby, I. T. ISCO Technology Overview: Do You Really Understand the Chemistry?. In Contaminated Soils, Sediments and Water, Vol. 10; Calabrese, E. J., Kostecki, P. T., Dragun, J., Eds.; Springer: New York, 2006. (28) Furman, O. S.; Teel, A. L.; Watts, R. J. Environ. Sci. Technol. 2010, 44, 6423−6428. (29) Liang, C. J.; Huang, C. F.; Mohanty, N.; Lu, C. J.; Kurakalva, R. M. Ind. Eng. Chem. Res. 2007, 46, 6466−6479. (30) Liang, C. J.; Guo, Y. Y. Environ. Sci. Technol. 2010, 44, 8203− 8208. (31) Gu, X. G.; Lu, S. G.; Li, L.; Qiu, Z. F.; Sui, Q.; Lin, K. F.; Luo, Q. S. Ind. Eng. Chem. Res. 2011, 50, 11029−11036. (32) Waldemer, R. H.; Tratnyek, P. G.; Johnson, R. L.; Nurmi, J. T. Environ. Sci. Technol. 2007, 41, 1010−1015. (33) Huang, K. C.; Couttenye, R. A.; Hoag, G. E. Chemosphere 2002, 49, 413−420. (34) Kolthoff, I. M.; Woods, R. J. Am. Chem. Soc. 1966, 88, 1371− 1375. (35) Xu, X. H.; Ye, Q. F.; Tang, T. M.; Wang, D. H. J. Hazard. Mater. 2008, 158, 410−416. (36) Block, P. A.; Brown, R. A.; Robinson, D. Presented at the Fourth International Conference on the Remediation of Chlorinated and Recalcitrant Compounds, Monterey, CA, 2004. (37) Hua, Q.; Zhang, C.; Wang, Z.; Chen, Y.; Mao, K.; Zhang, X.; Xiong, Y.; Zhu, M. J. Hazard. Mater. 2008, 154, 795−803. (38) Yuan, F.; Hu, C.; Hu, X. X.; Qu, J. H. Water. Res. 2009, 43, 1766−1774. (39) Bohn, B.; Zetzsch, C. J. Phys. Chem. A 1997, 101, 1488−1493. (40) Chu, W.; Wang, Y. R.; Leung, H. F. Chem. Eng. J. 2011, 178, 154−160. (41) Zhao, Y.; Hao, R. L.; Guo, Q. J. Hazard. Mater. 2014, 280, 118− 126. (42) Sun, W. M. Power Environ. Protect. 2003, 19, 43−46 (In Chinese). (43) Chen, L.; Peng, X. Z.; Liu, J. H.; Li, J. J.; Wu, F. Ind. Eng. Chem. Res. 2012, 51, 13632−13638. (44) Liu, F. Experimental and Mechanism Study on Simultaneous Removal of SO2 and NOx Using Jet Bubble Reactor. Doctoral Dissertation, North China Electric Power University, Beijing, China, 2009. (45) Guo, T. X. Experimental Investigation on Simultaneous Removal of SO2 and NOx in Liquid Phase by New-Type Complex Absorbent. Doctoral Dissertation, North China Electric Power University, Beijinig, China, 2011. 6510

dx.doi.org/10.1021/ef501686j | Energy Fuels 2014, 28, 6502−6510