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001---ef0501602Partition of Heavy and Alkali Metals during Sewage Sludge Incineration


Energy & Fuels 2006, 20, 583-590

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Partition of Heavy and Alkali Metals during Sewage Sludge Incineration
Jun Han, Minghou Xu,* and Hong Yao
State Key Laboratory of Coal Combustion, Huazhong UniVersity of Science and Technology, Wuhan, Hubei, 430074, People’s Republic of China

Masami Furuuchi, Takeo Sakano, and Premrudee Kanchanapiya
Graduate School of Natural Science and Technology, Kanazawa UniVersity, Kodatsuno, Kanazawa, Ishikawa 920-8667, Japan

Chikao Kanaoka
Ishikawa National College of Technology, Kitacyujo, Tsubata, Kahoku-gun, Ishikawa 929-0392, Japan ReceiVed May 27, 2005. ReVised Manuscript ReceiVed January 18, 2006

In this paper, two approaches are used to investigate the transformation of heavy and alkali metals during sewage sludge incineration: a thermodynamic equilibrium calculation and experiments using a bench-scale combustor. The distribution of heavy and alkali metals at different temperatures (800, 700, 600, 400, and 200 °C) is studied in these experiments. The chemical equilibrium calculation shows that vaporized metallic compounds form a condensed phase at different temperatures. The calculation results show that vaporized lead compounds begin to transfer to the solid phase at 500 °C. The main species of solid lead in the flue gas is PbCl2(s) below 500 °C. Zn and K have the same temperature range (300-500 °C), while the conversion temperature for As is 600 °C and that for Na and Cu is above 800 °C. The experimental results also prove the feasibility of separating heavy and alkali metals according to their gas-solid transformation temperature zones. On the basis of the experimental results, the optimum separation temperature is <400 °C for Pb and K, 600 °C for As, and 800 °C for Cu, Zn, and Na.

Introduction A large amount of sewage sludge is produced in the world every year. In China, the total production of sewage sludge was up to 450 000 000 tons in 2003.1 The problem of how to deal with the sludge is one of the most serious environmental issues. Through the years, sewage sludge has been successfully processed by various methods, such as incineration, disposal in landfill, use as a fertilizer and soil improvement material for land, and use in concrete manufacture. The forecast for trends in the disposal of sewage sludge in the EU to 2005 is shown in Figure 1 .2 Although landfilling and agriculture application are the disposal methods most widely used at present, these methods are expected to be used to a lesser extent in the future because of more stringent environmental standards and the implementation of policies to improve recycling. A technical guideline for the handling and disposal of urban waste released in 1992 in Germany3 requires that the organic content of any material deposited in landfill sites must be less than 5% from 2005, which obviously indicates that only incineration slag or ash is suitable for landfilling. Therefore, it is anticipated that incineration will
* Author to whom correspondence should be addressed. E-mail: mhxu@mail.hust.edu.cn. (1) People’s Republic of China Yearbook, 2003. (2) Werther, J.; Ogada, T. Sewage sludge combustion. Prog. Energy Combust. Sci. 1999, 25, 55-116. (3) S?rum, L.; Frandsen, F. J.; Hustad, J. E. On the fate of heavy metals in municipal solid waste combustion. Part II. From furnace to filter. Fuel 2004, 83, 1703-1710.

Figure 1. Sludge disposal routes in the EU to 2005.

be increasingly used because of its advantages: a large reduction in sludge volume, stabilization of toxic metals in bottom ash, and thermal destruction of toxic organic constituents. Furthermore, the calorific value of dry sludge is comparable to that of brown coal, which may be reused through incineration. Hence, incineration represents one of the methods with the most potential for sewage sludge disposal in the foreseeable future. However, incineration has obvious shortcomings, such as the emission of dioxins and heavy metals. In particular, the concentration of heavy metals in sewage sludge is critically high and the emission standards for toxic metals during sewage sludge incineration are now stricter than before. Thus, a study on how to control and reduce the emission of toxic metals during incineration is extremely urgent. Up to now, much research into the behavior and transformation of heavy metals during thermal

10.1021/ef0501602 CCC: $33.50 ? 2006 American Chemical Society Published on Web 02/10/2006

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processes has been conducted.3-11 However, this research only concentrated on how to stabilize or reduce the emission of heavy metals, primarily using sorbents.12-15 Few studies have considered the separation of heavy and alkali metals in the flue gas, especially the reuse and recycling of metals during waste thermal processes. As mentioned above, the concentrations of several heavy metals (Pb, Cu, and Zn) are very high in sewage sludge. Heavy metals can account for approximately 0.5-2.0% and in some cases even up to 4% of the total dry weight.16-19 Hence, it is possible and important to recycle or reuse some heavy metals from the fly ash during sewage-sludge incineration to concentrate the heavy metals in the fly ash. Dahl and Obernberger21 found that heavy metals could be separated when the fly ash from waste incineration was treated under some controlled conditions (temperature and atmosphere). The experimental results indicated that Cd could be effectively separated from the ash samples by thermal treatment at 900 °C, while that required for Zn was above 908 °C. Furthermore, the composition of the ash and the content of unburned carbon and sintering ash play an important role in the separation of heavy metals during ash treatment.17
(4) Nammari, D. R.; Hogland, W.; Marques, M.; Nimmermark, S.; Moutavtchi, V. Emissions from a controlled fire in municipal solid waste bales. Waste Manage. 2004, 24, 9-18. (5) Bakoglu, M.; Karademir, A.; Ayberk, S., Partitioning characteristics of targeted heavy metals in IZAYDAS hazardous waste incinerator. J. Hazard. Mater. 2003, B99, 89-105. (6) Thipse, S. S.; Dreizin, E. L. Metal partitioning in products of incineration of municipal solid waste. Chemosphere 2002, 46, 837849. (7) Delay, I.; Swithenbank, J.; Argent, B. B. Prediction of the distribution of alkali and trace elements between the condensed and gaseous phases generated during clinical waste incineration. J. Alloys Compd. 2001, 320, 282-295. (8) Jung, C. H.; Matsuto, T.; Tanaka, N.; Okada, T. Metal distribution in incineration residues of municipal solid waste (MSW) in Japan. Waste Manage. 2004, 24, 381-391. (9) Wang, K. S.; Chiang, K. Y.; Tsai, C. C.; Sun, C. J.; Tsai, C. C.; Lin, K. L. The effects of FeCl3 on the distribution of the heavy metals Cd, Cu, Cr, and Zn in a simulated multimetal incineration system. EnViron. Int. 2001, 26, 257-263. (10) Senior, C. L.; Bool, L. E.; Morency, J. R. Laboratory study of trace element vaporization from combustion of pulverized coal. Fuel Process. Technol. 2000, 63, 109-124. (11) Kirk, D. W.; Chan, C. C.; Marsh, H. Chromium behavior during thermal treatment of MSW fly ash. J. Hazard. Mater. 2002, B90, 39-49. (12) Liu, J.; Zheng, C.; Zeng, H.; Zhang, J.; Lu, X, Effects of solid adsorbents on the emission of heavy metals during coal combustion. EnViron. Sci. 2003, 5, 23-27 (in Chinese). (13) Chen, J. C.; Wey, M. Y.; Ou, W. Y. Capture of heavy metals by sorbents in incineration flue gas. Sci. Total EnViron. 1999, 228, 67-77. (14) Chen, J. C.; Wey, M. Y.; Lin, Y. C. The adsorption of heavy metals by different sorbents under various incineration condition. Chemosphere 1998, 13, 2617-2625. (15) Linak, W. P.; Srivastava, R. K.; Wendt, J. O. L. Sorbent capture of nickel, lead, and cadmium in laboratory swirl flame incineration. Combust. Flame 1995, 100, 241-250. (16) Gale, T. K. Mechanisms governing multi-species metals capture by kaolinite, hydrated lime and novel sorbents in high-temperature combustion environments. Ph.D. Dissertation, University of Arizona, Tucson, AZ, 2001. (17) Tyagi, R. D.; Couillard, D.; Tran, F. T. Comparative study of bacterial leaching of metals from sewage sludge in continuous stirred tank and air-lift reactors. Process Biochem. 1991, 26, 47-54. (18) Jain, D. K.; Tyagi, R. D. Leaching of heavy metals from anaerobic sewage sludge by sulfur-oxidizing bacteria. Enzyme Microb. Technol. 1992, 14, 376-383. (19) Sreekrishnan, T. R.; Tyagi, R. D.; Blais, J. F.; Campbell, P. G. C. Kinetics of heavy metal bioleaching from sewage sludgesI. Effects of process parameters. Water Res. 1993, 27, 1641-1651. (20) Blais, J. F.; Tyagi, R. D.; Auclair, J. C. Metals removal from sewage sludge by indigenous iron-oxidizing bacteria. J. EnViron. Sci. Health, Part A: Toxic/Hazard. Subst. EnViron. Eng. 1993, 28, 443-467. (21) Dahl, J.; Obernberger, I. Thermodynamic and experimental investigation on the possibilities of heavy metals recovery from contaminated biomass ashes by thermal treatment. In Proceedings of the 10th European Bioenergy Conference, Wurzburg, Germany, 1998; pp S241-S245.

Figure 2. Phase diagram for Pb and Zn during sewage sludge incineration.
Table 1. Properties of the Sewage Sludge Used element Cd Hd Nd Sd Cl percentage 0.97 0.47 0.682 0.21 0.03 element Al Ca Fe K Na P ppm 17 088 38 214 76 793 19 625 35 306 19 567 element As Pb Zn Cr Mn Cu ppm 527 1426 5010 508 4929 2719

Because metal compounds have different melting- and boiling-point temperatures, vapor-phase metallic compounds in flue gas will transform into the condensed phase when the temperature is below their dew-point temperature. Hence, it is possible to separate and recycle a proportion of the heavy metals in a special temperature zone using fly ash capture. However, the factors influencing the separation process other than the dewpoint temperature are not clearly known. In the present study, both thermodynamic equilibrium calculations and experiments at different temperatures (800, 700, 600, 400, and 200 °C) are conducted to identify the optimum temperature for separating and recycling some heavy metals. Thermodynamic Equilibrium Calculation. Some heavy metals, such as Pb, As, Na, and K, vaporize at high temperatures. Then, these metallic vapors transform into the solid phase as the temperature of the flue gas decreases to the appropriate dew point. However, these transformation processes are very complicated. It is difficult to clearly describe the distribution and speciation of heavy metals during incineration processes using experimental methods. Here, calculations for a thermodynamic equilibrium model (Factsage 5.3), instead of experiments, is used to predict the transformation and distribution of heavy and alkali metals. The input data for the calculations are taken from experiments as shown in Table 1, in which the elements considered in the model are C, H, O, N, Cl, S, Al, Ca, Fe, P, Na, K, As, Cr, Cu, Mn, Pb, and Zn. Figure 2 shows the distribution and transformation of Pb and Zn during sewage sludge incineration. For Pb, the major lead species is PbCl(g) when the temperature is above 1000 °C, while PbCl2(g) is the major compound in the temperature range 4001000 °C. These vaporized lead compounds start to transform

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Figure 5. Schematic diagram of the bench-scale combustor.

KFeCl2(s) is predicted to be the main solid species at low temperatures. The temperature range for conversion of K is suggested to be 400-500 °C, as shown in Figure 4. Experimental Section
Figure 3. Phase diagram for As and Cu during sewage sludge incineration.

Figure 4. Phase diagram for Na and K during sewage sludge incineration.

into the solid phase [PbCl2(s)] below 500 °C. For Zn, ZnCl2(g) is the major zinc compound at high temperatures and it can react with other components, such as sulfide or sulfur dioxide, in the flue gas to form solid-phase ZnS(s) if the temperature is below 500 °C. Figure 3 illustrates that As is in the condensed phase below 700 °C. All of the As compounds convert to the solid phase when the temperature is below 600 °C, while the temperature range for Cu is 700-1000 °C. The behavior of Na and K is shown in Figure 4. The major solid species of Na is NaCl(s), and the temperature range for the gas-solid transformation process is 600-800 °C. For K,

Experiments are carried out using an electrically heated benchscale combustor, as shown in Figure 5. The temperatures in the combustor are automatically controlled by thermocouples, and the wall temperature is kept at approximately 1400 °C by an electrical heater. The concentrations of SO2, NOx, CO, CO2, and O2 in the flue gas at the outlet of the chamber are monitored online by a gas analyzer. During experiments, the stoichiometric combustion air ratio is kept at approximately 1.2 in the combustion chamber. Dry sewage sludge is milled and sieved before the experiments to particles with a diameter of approximately 200 ?m. A feeder is located at the top of the combustor, and a feeding rate of approximately 5 g/min is used. The composition of the sewage sludge is summarized in Table 1. The sampling tube, which is made of quartz, consists of seven parts, as shown in Figure 6, with a diameter and length of 0.034 and 1 m, respectively. In addition, silica filters are used to capture the fly ash during experiments. The temperature of each part of the sampling tube is electrically heated and independently controlled by thermocouples. The temperature of the filter is the same as that of the sampling tube at each location. According to the thermodynamic calculation, the conversion temperature range for selected metals is 300-800 °C. Therefore, temperatures of 800, 700, 600, 400, and 200 °C are considered in these experiments. The experimental conditions are summarized in Table 2; The flow rate of the flue gas in the sampling device is approximately 0.11 m/s, and the residence time in the second and third filters is approximately 0.58 and 0.47 s, respectively. The mass ratio of the fly ash in the three filters is approximately 0.099:0.026:0.025. To capture all of the vapor-phase metallic compounds in the flue gas, the HNO3 and H2O solution is cooled using ice. Before the experiments, a blank test is performed to eliminate the effect of filter composition and other factors (HNO3 and H2O, experimental setup, etc.) on the concentration of heavy and alkali metals. The experimental conditions for the blank test are the same as for other experiments but without the addition of sewage sludge. After the test, the fly ash and filter are placed in a drying oven (105 °C) for 12 h to eliminate the water. The samples of fly ash in each filter and the fly ash deposited on the wall are digested separately in a HNO3-HF-HClO4 solution. The heavy and alkali metals are then analyzed by inductively coupled plasma-atomic absorption spectrometry (ICP-AAS). At the same time, the solution of HNO3 and H2O used to capture vapor-phase metals are also analyzed. In addition, particles of the fly ash from the filter are analyzed by scanning electron microscopy-energy dispersive X-ray (SEM-EDX) to identify their structure and composition.

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Figure 6. Outline of the sampling device.

Figure 7. Distribution of Pb at different temperatures.
Table 2. Experimental Conditions approximate temperature (°C) case A B C chamber 1000 1000 1000 first filter 800 800 800 second filter 700 600 400 third filter 200 200 200 PbCO3 PbCl2 PbO PbO2 PbSO4 PbS Table 3. Properties of Lead Compounds color white white yellow brown white gray melting point (°C) 315, decomposes 501 888 290, decomposes 1084 1120 boiling point (°C) 954 1470

Results and Discussion Pb Separation. The distribution of Pb in flue gas is shown in Figure 7. Most lead compounds are found in the third filter rather than the first one. The capture efficiency for Pb in the first filter (ηPb,1,C) is quite low, only 4.0% in case C (800 °C). Even the highest efficiency is only 10.5% (ηPb,1,B) in case B. The capture efficiency is defined as follows: As depicted in Figure 7, the effect of filters is very obvious. The capture efficiency (ηPb,3,k) varies from 27.8 to 44.3% in the experiments. In addition, it is evident that the efficiency of the second filter (ηPb,2,k ) 3.0-19.7%) is less than that of the third filter, which means that a large amount of lead is vaporized during incineration of the sewage sludge because of its low boiling-point temperatures, as listed in Table 3. These gaseous lead compounds are converted to the solid phase through chemical and physical transformation at temperatures below their dew points. Figure 7 also shows that the removal efficiency of the filters can be changed by varying the temperature. When the temperature of the second filter is kept at 700 °C, the removal efficiency is only 3.0% (ηPb,2,A). The efficiency is increased to 3.5% (ηPb,2,B) at 600 °C and then remarkably increased to 19.7% (ηPb,2,C) at 400 °C. A comparison shows that ηPb,3,C is approximately twice as high as ηPb,2,C, although the temperature

ηi,j,k )

mi,j,k mi,k + mi,water+HNO3,k + mi,wall,k

where ηi,j,k is the capture efficiency of element i in filter j for case k (A, B, or C), mi,j,k is the mass of element i in filter j for case k, mi,k is the mass of element i in the three filters for case k, mi,water+HNO3,k is the mass of element i in the solution of H2O and HNO3 for case k, and mi,wall,k is the mass of element i deposited on the walls for case k.

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Figure 8. Mass fraction of metallic compounds in the filters in case A.

Figure 9. SEM-EDX analysis of samples in the third filter in case C.

Figure 10. SEM-EDX analysis of samples in the second filter in case C.

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Figure 11. SEM-EDX analysis of samples in the first filter in case C.

lead compounds is below 400 °C, which is consistent with the results of the thermal equilibrium calculation shown in Figure 2, according to which 90% of the lead compounds are in the gas phase and gaseous lead compounds completely transform into the solid phase at 300 °C. In addition, the temperature of the second filter in case A is the highest (700 °C) in all of these experiments. Most of the vapor-phase lead compounds could not transform into the solid phase at such a high temperature. Thus, more vapor-phase lead compounds escape from the second filter with the flue gas. These lead compounds will then transform into the solid phase and be captured by the third filter, resulting in ηPb,3,A ) 44.27%, the highest value. The transformation process can be clearly observed from the SEM-EDX results. As shown in Figure 9, a high level of Pb is detected in the third filter sample, while no lead is found in either the first filter (Figure 10) or the second filter (Figure 11). Coarse particles in the flue gas are removed by the first filter. Moreover, the surface of the particles in the second and third filters is smooth, and the diameters are small (approximately

Figure 12. Distribution of Cu at different temperatures.

difference is only 200 °C. The mass fraction of various elements in the fly ash is shown in Figure 8. The fraction of lead in the fly ash collected by the third filter is the highest (26%) of the three, which is several-fold higher than that of lead ore. Hence, it is possible to recover Pb in the flue gas by collecting the fly ash. According to the experimental results, the conclusion can be drawn that the best separation and recovery temperature for

Figure 13. Mass fraction of metallic compounds in the filters in case C.

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Figure 14. Distribution of Zn at different temperatures.

1-3 ?m). Thus, we can speculate that these particles are formed by nucleation or condensation. Cu, Zn, and As Separation. Figure 12 illustrates the transformation of Cu in the flue gas. In contrast to lead, most Cu compounds in the flue gas are removed by the first filter, and the capture efficiency of the first filter is well above that of the other filters. Hence, it is not necessary to carry out any further experiments below 800 °C, and only case C is carried out. The efficiency of the first filter (ηCu,1,C) in capturing Cu is 37.8%, while the efficiencies of the second and third filters are 5.9% (ηCu,2,C) and 10.2% (ηCu,3,C), respectively. In addition, the

mass fraction of Cu in the fly ash of the first filter is 5.0%, which is higher than that in copper ore (0.9%). The mass fraction of various elements in case C is shown in Figure 13. Hence, the optimum temperature for Cu separation and reuse should be approximately 800 °C. It is likely that two mechanisms allow for highly effective scavenging of Cu in the flue gas by the first filter. One mechanism involves fine sludge particles being carried by the flue gas and escaping from the combustion chamber quickly; because Cu in the fine particles cannot be completely vaporized in such a short time in the combustion chamber, Cu is hence captured by the first filter. In the second mechanism, some Cu compounds could vaporize during sludge incineration, but because the boiling and melting points of these compounds are very high, they will transform into the solid phase at high temperatures. Thus, Cu compounds condensed into particles can be removed by the first filter. SEM-EDX analysis of samples in the first filter shows that the particle diameter is approximately 30 ?m and that the fly ash is porous. On the other hand, the surface of some particles is coated by other materials, which also proves that the particles in the first filter originated from two sources: fine sludge carried by the flue gas and the condensation of vapor-phase metallic compounds.

Figure 15. Distribution of As at different temperatures.

Figure 16. Distribution of Na at different temperatures.

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Figure 17. Distribution of K at different temperatures.

As shown in Figure 14, the optimum temperature is 800 °C for Zn, which is similar to that for Cu. The efficiency of the first filter is much higher than that of the other filters, and thus, only experimental case C is carried out. The temperature for As separation should be 600 °C, as shown in Figure 15, which is consistent with the result of the thermodynamic calculation. Because most arsenical compounds are highly volatile, a large amount of As will still exist in the gas phase, even below 200 °C. The results are not consistent with the calculation results shown in Figure 3. The reasons may be as follows. In the thermal equilibrium model, the database of compounds is not large enough, which may result in differences between the experiments and the calculation. In addition, the concentrations of the two elements are too low to be detected by SEM-EDX, as shown in Figures 9-11. Na and K Separation. Detailed experimental results for Na and K transformation are shown in Figures 16 and 17, respectively. The mass fraction of Na in the fly ash of the third filter in case C is 45%. However, ηNa,3,C is not the highest efficiency (16.2%) of the three filters: ηNa,1,C is the highest efficiency (35.1%), and filter 1 also shows a very high mass fraction (28.1%). Considering both the mass fraction in the fly ash and the capture efficiency of the filters, the optimum temperature for Na separation is 800 °C. For K, both the capture efficiency and mass ratio are highest for the third filter; thus, 200 °C is suitable for K. According to the SEM-EDX results shown in Figure 11, the contents of K, Na, and Cl are very high and the main K and Na species are suggested to be KCl and NaCl, respectively, which is also reported by French and Milne.22 In their experiments, KCl was the dominant transport form, while KOH was

a secondary volatile species when the Cl/K molar ratio was approximately 0.5. Conclusion Thermodynamic equilibrium calculations and an experimental investigation are carried out to identify the optimum separation temperature for specific heavy and alkali metals. A comparison of the calculated and experimental results shows that only some of the calculation results are reasonable. Consequently, an improvement of the model is necessary before the calculation can be used to explain the behavior of heavy metals during waste incineration. Nevertheless, results for thermal equilibrium calculations can provide useful information or guidance for the experiments. According to the experimental results, heavy and alkali metals can be separated at a specific temperature. The optimum temperature for Pb and K is below 400 °C, while that for As is 600 °C, and the best temperature for Cu, Zn, and Na should be 800 °C.
Acknowledgment. The authors would like to acknowledge the financial support provided by the National Natural Science Foundation of China (Grant number 50325621) and the National Key Basic Research and Development Program of China (Grant number 2002CB211602).
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(22) French, R. J.; Milne, T. A. Vapor phase release of alkali species in the combustion of biomass pyrolysis oils. Biomass Bioenergy 1994, 7, 315325.


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