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Fixed-bed studies of the interactions between mercury and coal combustion fly ash
Fuel Processing Technology 82 (2003) 197 – 213 www.elsevier.com/locate/fuproc
Fixed-bed studies of the interactions between mercury and coal combustion fly ash Grant E. Dunham a, Raymond A. DeWall a, Constance L. Senior b,* a b
Energy and Environmental Research Center, 15 North 23rd Street, Grand Forks, ND 58203, USA Reaction Engineering International, 77 West 200 South, Suite 210, Salt Lake City, UT 84101, USA
Abstract Sixteen different fly ash samples, generated from both pilot-scale and full-scale combustion systems, were exposed to a simulated flue gas containing either elemental mercury or HgCl2 in a bench-scale reactor system at the Energy and Environmental Research Center to evaluate the interactions and determine the effects of temperature, mercury species, and ash type on adsorption of mercury and oxidation of elemental mercury. The fly ash samples were characterized for surface area, loss on ignition, and forms of iron in the ash. While many of the ash samples oxidized elemental mercury, not all of the samples that oxidized mercury also captured elemental mercury. However, no capture of elemental mercury was observed without accompanying oxidation. Generally, oxidation of elemental mercury increased with increasing amount of magnetite in the ash. However, one high-carbon subbituminous ash with no magnetite showed considerable mercury oxidation that may have been due to unburned carbon. Surface area as well as the nature of the surface appeared to be important for oxidation and adsorption of elemental mercury. The capacity of the ash samples for HgCl2 was similar to that for elemental mercury. There was a good correlation between the capacity for HgCl2 and the surface area; capacity decreased with increasing temperature. D 2003 Elsevier Science B.V. All rights reserved. Keywords: Mercury control; Coal-fired power plant; Fly ash
1. Introduction On December 14, 2001, the U.S. Environmental Protection Agency (EPA) announced that it would regulate mercury emissions from coal-fired utility boilers with regulations being proposed in 2003 and implementation expected by 2007. Before electric utilities can implement emission minimization strategies for mercury, they must have an accurate * Corresponding author. 0378-3820/03/$ - see front matter D 2003 Elsevier Science B.V. All rights reserved. doi:10.1016/S0378-3820(03)00070-5
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means of predicting emissions in all effluent streams for the range of fuels and operating conditions commonly utilized. Results from sampling mercury from full-scale coal-fired boilers as part of EPA’s Information Collection Request (ICR) [1] indicate a wide range of mercury concentrations and splits between elemental, oxidized, and particulate-bound mercury. The existing air pollution control devices (APCDs) on a coal-fired boiler can remove a significant amount of mercury. The split between gaseous elemental mercury and gaseous oxidized mercury affects how much mercury is removed in wet scrubbers as well as how much mercury is adsorbed by fly ash. The partitioning between gaseous and particulate-bound mercury at the back end of a boiler determines how much mercury will be removed by the particulate control device (PCD). Ash plays a role in both the adsorption of mercury and the oxidation of elemental mercury in flue gas. Some insight into the mechanism or mechanisms can be gained by looking at the body of research concerning mercury and activated carbon sorbents. Both elemental mercury and HgCl2 are adsorbed by activated carbon at temperatures characteristic of the PCD. Activated carbon capacities at saturation are much less than monolayer coverage [2], which suggests that only certain sites on the surface are active for mercury adsorption. The equilibrium capacity is a function of the concentration of mercury in the gas [2,3] and has an inverse dependence on temperature [2]. The adsorption of HgCl2 on activated carbon is not greatly affected by flue gas composition [3,4]. The adsorption of Hg0, however, is highly dependent on gas composition and involves interaction with the acid gases (NOx, SOx, HCl). Based on fixed-bed studies, researchers have concluded that the adsorption of elemental mercury on activated carbon occurs by chemisorption rather than physisorption [5,6]. The result of adsorption of Hg0 on carbon is an oxidized compound (and not elemental mercury) on the surface of the carbon, based on X-ray absorption fine structure spectroscopy (XAFS) [7] and desorption studies [2,8]. Unburned carbon has been suspected of adsorbing mercury for both eastern and western bituminous and subbituminous coals. Often, as a consequence of low-NOx burners or lowNOx combustion systems, pulverized coal boilers can produce high levels of unburned carbon when burning bituminous coals [9] or, less commonly, subbituminous coals [10]. Mercury has been found to concentrate in the carbon-rich fraction of fly ash [11,12]. Recently, data on mercury in ash and carbon content were collected from sampling campaigns on full-scale coal-fired power plants [13]. Data from nine different plants were examined. Surprisingly good correlations of Hg removal with C/Hg ratio were observed for a range of coal and boiler types. There was a clear effect of temperature in comparing data from hot-side electrostatic precipitators (ESPs) with that from cold-side ESPs. Thus, for bituminous coals, this suggested that mercury can be adsorbed by the unburned carbon in the ash. Unfortunately, the largest database of mercury speciation and removal in existing coalfired power plant APCDs, the EPA’s Information Collection Request (ICR) database, does not include any information on the carbon content of the ash. A recent attempt to estimate the carbon content from a subset of ICR data [14] concluded that increased carbon in the fly ash was positively correlated with increased mercury removal across cold-side ESPs, for both bituminous and subbituminous coal. However, it is not possible to generalize to all coal types and APCDs and conclude that high carbon in ash will always give high levels of particulate-bound mercury.
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Norton et al. [15] carried out fixed-bed tests on two fly ash samples taken from ESP hoppers in full-scale boilers. The ash samples were from a Powder River Basin (PRB) coal and a Pittsburgh seam bituminous coal. In spite of large differences in mineralogy, the two samples did not show large differences in their ability to oxidize elemental mercury. Acid gases (HCl, NO2, and SO2) promoted oxidation of elemental mercury. The presence of NO appeared to suppress oxidation. In the Pittsburgh seam bituminous coal, the magnetic portion (i.e., iron-rich) of the ash did not appear to be more catalytically active than the nonmagnetic phases; however, differences in surface area between the magnetic and nonmagnetic portions made it difficult for the authors to conclude this with certainty. Lee et al. [16] studied simulated ash mixtures and actual ash samples in another fixedbed experiment using simulated flue gas. They identified NOx and HCl as gases that were important for the oxidation of elemental mercury. Both copper oxide (CuO) and iron oxide (Fe2O3) were identified as active compounds for oxidation of elemental mercury in the simulated ash mixtures. In the presence of NOx, even alumina and silica became active in promoting mercury oxidation. Of the fly ash samples tested, a Pittsburgh seam bituminous ash had significant catalytic activity with respect to mercury oxidation, especially in the presence of NOx. The other fly ash samples were also more catalytically active in the presence of NOx, as compared to HCl. In an entrained flow reactor, using exhaust gas from coal combustion, Zhuang et al. [17] also observed that elemental mercury could be oxidized and adsorbed on the surface of fly ash particles. Experiments carried out with pure oxide particles in the entrained flow reactor showed that iron oxide was effective at oxidizing elemental mercury. Fixed-bed studies can provide insight into the mechanisms for mercury adsorption on fly ash. The specific goal of this work was to perform bench-scale tests to evaluate the interactions of mercury with ash and ash components at temperatures characteristic of cold-side particulate collection devices.
2. Experimental 2.1. Sample characterization Ash was collected from two pilot-scale rigs, the University of Arizona selfsustained combustor [18] and the University of North Dakota EERC particulate test combustor (PTC) [19]. The coals burned included eastern bituminous (Ohio blend, Pittsburgh), western subbituminous (Wyodak, Absaloka, Cordero Rojo), and North Dakota lignite. At the University of Arizona, a downflow combustor with a firing rate of 17 kW was used to generate ash samples. Combustion was self-sustained in the furnace, which was 6 m long and had an internal diameter of 0.15 m. Ash samples were collected at a temperature of 200 jC using a special collection system consisting of a heated filter. The filter was installed in a horizontal section between the exit of the hot zone and the baghouse. Because the filter was upstream of an ash trap, the largest ash particles might not have been collected.
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The Pittsburgh (Blacksville) bituminous coal was burned in the PTC, and the ash was collected at 180 jC in a fabric filter pulse-jet baghouse equipped with Ryton bags. Sampling with EPA Method 29 was conducted during this test. The total mercury concentration at the inlet of the baghouse was 6.9 with 0.7 Ag/m3 equivalent mercury on the fly ash. Based on the Method 29 results, 85% of the mercury in the flue gas was oxidized. However, this method has since been shown to have high bias for oxidized mercury. The Absaloka and Cordero Rojo subbituminous coals were also burned in the PTC, but the ash was collected in the advanced hybrid particulate collector (AHPC), which is a combined ESP and baghouse equipped with GORETEXR membrane filter bags. No gas-phase mercury sampling was associated with these ash samples. In addition to the pilot-scale ash samples, a number of ash samples were taken from the hoppers of particulate collection devices at full-scale power plants. Five ash samples (AA, MA, DMA, GA, DA) were taken from the ESP hoppers of full-scale power plants burning eastern bituminous coals [9]. Additional information on the source of these ashes is provided in Table 1. A full-scale ash sample obtained from the ESP hopper at Coal Creek station located in North Dakota was evaluated. This was the same parent coal as from the North Dakota lignite sample burned at the University of Arizona. The Coal Creek station had two tangentially pc-fired boilers with an ESP upstream of a wet scrubber. The Valmy station ash was from a station burning a western bituminous coal in two units with low-NOx burners. The ash was collected from Unit 2, which was equipped with a spray dryer followed by a reverse-gas baghouse. Based on the concentration of mercury in the coal, the estimated mercury flue gas concentration was 4.2 Ag/m3. The mercury concentration on the base ash was 0.557 Ag/g, which translates to roughly 3.8 Ag/m3 of mercury in the flue gas, assuming a particulate loading of 2.5 gr/scf (6.86 g/m3). Based on these concentrations, the ash captured roughly 91% of the mercury in the flue gas. Sampling at the station indicated nearly all of the mercury in the flue gas was captured by the fly ash. A sample of ash from the Comanche station (formerly Public Service of Colorado) was collected by ADA Technologies on January 23, 1998 when the plant was burning a PRB coal from the Belle Ayr Mine. The station had two combustion units. Unit 1 was a drybottom, tangential-fired boiler with no NOx controls. Unit 2 was a dry-bottom, wall-fired boiler using overfire air for NOx control. Both units were equipped with a fabric filter dust collector. It was not specified from which unit the sample was collected. According to information in the literature [21], the mercury removal across the reverse-gas baghouse was 34 –78% at 138 jC (280jF) and less than 30% at 165– 175 jC (330 –350jF). The Table 1 Characteristics of Eastern Bituminous coal/ash samples from full-scale power plants Parameter
GA
DMA
DA
AA
MA
Hg in coal (Ag/g) Cl in coal (Ag/g) Coal ash content (%) Hg removal by ash (%) Boiler type
0.18 1300 13.6 6.8 Pulverized Coal
0.09 1600 10.9 7.8 Pulverized Coal
0.12 1100 13.9 24.0 Pulverized Coal
0.09 1100 18.5 10.1 Cyclone
0.09 2300 10.1 42.4 Stoker
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concentration of mercury on the fly ash indicated a significant amount of mercury capture by the fly ash during combustion. This is consistent with the reported data that showed less than 2 Ag/m3 vapor-phase mercury in the flue gas. The mercury concentration in the coal indicated approximately 5.4 Ag/m3 total mercury in the flue gas. A western bituminous coal was burned at the Valmont plant in a dry-bottom, tangentialfired boiler equipped with low-NOx burners. Particulate control was achieved with a fabric filter baghouse. Ash samples were collected from this plant as part of ICR sampling activity. Very little mercury was present in the coal. The Black Thunder subbituminous coal was burned in the Big Stone plant in Milbank, SD. The unit was a cyclone-fired boiler with overfire air for NOx control. The ash was collected in a demonstration AHPC. No mercury sampling was associated with this sample. Loss on ignition (LOI) and surface area were measured for the ash samples; surface area was computed from the nitrogen Brunauer – Emmett – Teller (B.E.T.) isotherm equation. As a part of this program, three pilot-scale ashes (Ohio, Wyodak and North Dakota lignite) were also characterized by the University of Utah. The investigation of particle morphology was carried out with a scanning electron microscope. In addition to measuring the LOI and surface area, 57Fe Mo¨ssbauer spectroscopy was carried out by Dr. Frank Huggins at the University of Kentucky for selected samples in order to quantify the state of iron in the ash. This was done because previous work [16] has suggested that iron oxides in fly ash may be responsible for some portion of catalytic oxidation of mercury. Mo¨ssbauer spectroscopy is a powerful, element-specific technique for iron in complex materials. In the area of coal characterization, it has been used for identification of iron minerals, quantitative estimates of pyritic sulfur, oxidation and weathering studies, and conversion of iron minerals during coal combustion, liquefaction, and gasification. Details of Mo¨ssbauer spectroscopy as applied to coal and coal ash have been discussed elsewhere [20]. 2.2. Fixed-bed apparatus and procedures A bench-scale system has been developed at the EERC for evaluating mercury sorbents and fly ashes in a fixed-bed configuration. A detailed description of this system has been previously reported [21]. The system, diagrammed in Fig. 1, was composed of three subsystems: the mercury/gas delivery system, the fixed-bed system, and the mercury measurement system. Mass flow controllers were used to maintain constant gas flow for most of the flue gas constituents. The mercury and mercuric chloride permeation tubes were maintained at a constant temperature and continuously purged with nitrogen to ensure constant outlet mercury concentrations. The output from each of the permeation tubes was periodically verified by taking EPA Method 101A samples at the inlet of the fixed-bed system. A simulated flue gas was used for all tests. The flue gas compositions for some of the tests were based on the proximate/ultimate analysis data for the parent coal. The flue gas blends are presented in Table 2. The flue gas mixture used for the other ash samples was the baseline mix used for evaluating sorbents for mercury control (in previous work at the EERC) and is also presented in Table 2.
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Fig. 1. Fixed-bed reactor system.
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Table 2 Flue gas mixtures for bench-scale tests Flue gas component
Ohio blend
North Dakota lignite
Wyodak
Pittsburgh, AA, DA, DMA, GA, MA
Valmy, Comanche, Dale Valmont, Black Thunder, Cordero Rojo, Absaloka
O2 (%) CO2 (%) H2O (%) SO2 (ppmv) NO (ppmv) NO2 (ppmv) HCl (ppmv)
3.3 14.0 8.0 1900 300 15 60
3.0 14.0 14.7 840 300 15 5
3.0 14.0 12.0 300 300 15 5
4.0 14.0 6.0 1700 550 27 50
6.0 12.0 8.0 1600 300 20 50
The ash samples were evaluated in a fixed bed consisting of a Teflon-coated 2.5-in.diameter filter holder. A quartz filter nominally loaded with 1.5 g of fly ash made up the actual fixed bed. The fixed-bed assembly was maintained at the desired temperature inside an oven that could be controlled to F 1 K. A slipstream of the flue gas could be sent to the continuous mercury monitors (CMMs) either from the inlet or the outlet of the fixed bed. The mercury measurement system consisted of a conditioning/conversion system and several CMMs. The conditioning/conversion system conditioned the sample gas by removing moisture and acid gases before sending it to the CMMs. In order to determine the amounts of elemental and oxidized forms of mercury, the conditioning/conversion system either reduced all forms of mercury to elemental mercury for a measurement of total mercury concentration or removed the oxidized forms of mercury from the flue gas stream for a measurement of elemental mercury concentration. The difference between the total mercury and Hg0 was assumed to be some form of Hg2 +. Three different CMMs were used to measure outlet mercury concentrations from the bench-scale system: a Semtech 2000, a PS Analytical Sir Galahad, and a Tekran Model 2537A. Fig. 2 is a plot of the outlet mercury concentration as a function of time from one run where all three instruments were used to measure the mercury concentration. The agreement among the instruments was very good. The results from the tests are presented as mercury concentration (normalized to a percent of the inlet mercury concentration) as a function of time. In addition to having good agreement between instruments, the results from the system have been shown to be very repeatable [21]. The test protocol for each run was essentially the same. At the start of each test, the concentration of total mercury at the outlet of the fixed bed was measured until 100% breakthrough was reached (outlet mercury concentration equals the inlet mercury concentration). At this point, the conditioning/conversion system was switched to measure the elemental mercury concentration at the outlet of the fixed bed (noted on the plots as Hg0). If no oxidation was taking place, the concentration would remain at 100% of the inlet elemental mercury concentration; however, if any of the elemental mercury was being oxidized across the filter, it would result in a drop in the outlet concentration of elemental mercury. After establishing the outlet elemental mercury concentration, the system was switched to measure the total inlet mercury concentration. When measuring the elemental mercury concentration at the outlet of the fixed bed while injecting HgCl2, the expected
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Fig. 2. Example of mercury measurements from three continuous mercury analyzers.
concentration was zero. If it was higher than zero, this indicated that some of the HgCl2 had been reduced to elemental mercury.
3. Results 3.1. Sample characterization Table 3 shows the characteristics of the pilot-scale ash. The bituminous ash samples had approximately 5 wt.% LOI and surface areas of 1.5 and 2.0 m2/g. The Wyodak ash had a very high LOI and surface area not typical of low-rank coals burned at full scale. The other two subbituminous ashes had more typical properties. SEM analyses of three of the pilotscale ash samples showed that the Ohio and North Dakota ash samples contained large, irregular-shaped particles, and no ultrafine aggregates, whereas the Wyodak sample was found to contain a lot of ultrafine aggregates, as shown in Fig. 3. The surface area information and the SEM pictures suggested that the Wyodak sample contained some soot and that the other two pilot-scale samples from the University of Arizona combustor did Table 3 Characteristics of pilot-scale ash Coal
Coal rank
Test rig
LOI, dry (wt.%)
SA (m2/g)
Hg, as received (Ag/g)
Ohio blend ND lignite Wyodak Pittsburgh Absaloka Cordero Rojo
Bituminous Lignite Subbituminous Bituminous Subbituminous Subbituminous
Arizona Arizona Arizona EERC PTC EERC PTC EERC PTC
6.79 2.86 4.45 4.3 1.34 0.21
1.4 0.68 9.33 1.95 3.04 0.74
0.276 0.536 1.29 0.14 0.223 0.183
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Fig. 3. SEM photo of Wyodak ash from University of Arizona combustor (24,900 times magnification).
not. The presence of soot would account for the extraordinarily high surface area of the Wyodak ash and was probably the result of the specific pilot-scale combustion system in which the ash was generated. While such high carbon (and surface area) is not typical of ash from full-scale systems burning subbituminous coal, this unusual ash makes for an interesting study on the potential for soot to affect mercury in flue gas. Table 4 provides a summary of the properties of the full-scale ash samples. The LOI of the bituminous ashes spanned a wide range from 1 to 45 wt.%, which was probably indicative of different firing systems and coal rank. Some of these ash samples also had very high surface areas. The LOI of the western ashes (bituminous and subbituminous) Table 4 Characteristics of full-scale ash Fly ash
Coal rank
LOI, dry (%)
Surface area (m2/g)
Mercury, Ag of Hg/g of ash
Comanche Valmy Coal Creek Valmont Black Thunder AA MA DA GA DMA
Subbituminous Western bituminous Lignite Western bituminous Subbituminous Eastern bituminous Eastern bituminous Eastern bituminous Eastern bituminous Eastern bituminous
0.90 4.50 0.01 0.88 1.38 37.12 44.4 6.22 0.67 1.5
1.35 5.98 NMa 1.49 3.0 4.36 3.92 6.17 0.17 0.96
1.05 0.557 0.0067 NM 2.13 0.282 0.060 0.114 0.002 0.051
a
NM = Not measured.
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was less than 5 wt.%. The surface areas of these ashes were between 1 and 6 m2/g, with the exception of the ash from the Coal Creek plant. In general, the full-scale samples had a wide range of carbon contents (measured as LOI); some of the ash from bituminous coals had very high carbon. The surface area of the full-scale ash samples did not appear to follow the carbon content. For example, ash samples AA and MA had very high LOI values, but the surface area of these samples was not commensurately high. The Mo¨ssbauer data (Table 5) indicate that the iron in all the samples was present in either glass or iron oxide phases. Two different subsets of samples were identified depending on whether or not the calcium ferrite phases were significant. In the bituminous ash samples, iron was found in oxide phases or in glassy phases, whereas in the subbituminous ash samples, the iron was predominantly calcium ferrite, with some hematite (iron oxide), presumably coming from pyrite in the coal. 3.2. Fixed-bed results In the first series of fixed-bed experiments, ash samples were exposed to elemental mercury in simulated flue gas at temperatures of 121 and 177 jC (250 and 350 jF). Both oxidation and adsorption of elemental mercury were calculated from the CMM data as a function of time. In the second series, ash samples were exposed to simulated flue gas containing HgCl2. Tables 6 and 7 apply to Series I and II, respectively. Each table includes the concentration of mercury on the spent ash (if it was submitted for analysis), the amount of mercury generated during the test, the average mass balance based on data from the CMMs used during the test, the percentage of the elemental mercury that was oxidized across the fixed bed, and the mercury collection efficiency of the ash at 100% breakthrough. The mass balance is the ratio of the mercury measured (on the ash and in the flue gas) to the mercury generated (based on the average mercury permeation rate determined with wet chemistry measurements). The collection efficiency at complete breakthrough is based on the mercury on the spent ash or the integrated area under the Table 5 Mo¨ssbauer results (wt.% Fe in each phase) Sample
Hematite
Magnetite
Fe+ 3 oxide
Fe+ 3 glass
Ohio (pilot) Wyodak (pilot) North Dakota lignite (pilot) Pittsburgh (pilot) Comanche (full scale) Valmont (full scale) Valmy (full scale) Ash GA (full scale) Ash AA (full scale) Ash DMA (full scale) Ash DA (full scale) Ash MA (full scale)
23 17 19 10
51.5
2
24
69
20
5
28
47
48.5 56.5 25 60.5 43
25.5 9 39 18 30
Fe+ 2 glass
Ca ferrite 83 82 93
15.5 8 10.5 7
Mag. Ca ferrite
19 13 10 26.5 36 11 19.5
87
7
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Table 6 Results from tests with Hg0 injection (Series I) Fly ash
Temperature Hg0 on initial Hg0 on Hg0 gen. Average mass Percent Hg0 Collection (jC) ash (Ag/g) spent ash (Ag) balance (%) oxidized efficiency (Ag/g) @ 100% breakthrough
Ohio Ohio Wyodak Wyodak N D lignite N D lignite Coal Creek Pittsburgh Pittsburgh Valmy Valmy Valmy Valmy Valmont Valmont Comanche Comanche Comanche AA AA AA DA DMA GA MA MA MA Black Thunder Black Thunder Codero Rojo Codero Rojo Absaloka Absaloka
121 177 121 177 121 177 177 121 177 121 121 121 177 121 177 121 177 177 121 121 177 121 121 121 121 177 121 121 177 121 177 121 177
a b c
0.276 0.276 1.29 1.29 0.536 0.536 0.0067 0.140 0.140 0.557 0.557 0.557 0.557 NM NM 1.05 1.05 1.05 0.282 0.282 0.282 0.114 0.051 0.002 0.060 0.060 0.060 2.13 2.13 0.183 0.183 0.223 0.223
NM NM 2.51 NM NM NM NM 1.21 NM NM 1.40 1.67 NM NM 0.09 NM 0.40 NM 0.93 1.28 NM NM NM NM 0.87 0.20 0.74 NM NM NM NM 2.33 0.51
NM NM 43.8 NM NM NM NM 30.6 NM NM 28.8 24.1 NM NM 24.3 NM 18.2 NM 26.2 33.7 NM NM NM NM 21.5 22.9 23.3 NM NM NM NM 27.3 16.9
NMa NM 102 NM NM NM NM 91 NM NM 82 113 NM NM 104 NM 118 NM 75 109 NM NM NM NM 111 104 108 NM NM NM NM 88 94
0 20 40 60 0 0 0 45 25 50 50 50 40 70 30 25 15 15 85 85 65 55 20 15 80 80 60 40 30 20 20 70 70
NM NM 3b 42c NM NM NM 5b NM NM NM 14b NM NM 3b NM 0b NM 6b 10b NM 64c NM NM 20b NM 8b NM NM NM NM 17a 3a
NM = Not measured. Mercury captured based on mercury on spent ash. Mercury captured based on integration of breakthrough curve.
curve (if ash analysis is not available) and the amount of mercury generated (permeation rate) in the time it took to reach 100% breakthrough. It appears the collection efficiencies based on the integrated breakthrough curve are higher than those based on ash analysis. The difference may be caused by offgassing of collected mercury during the remainder of the test, which is typical of similar tests with carbon-based sorbents [4]. However, the results do give a basis for comparison of ashes and their ability to capture mercury. For many samples, the collection efficiency was zero, and the ash showed no ability to capture mercury.
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Table 7 Results from tests with HgCl2 injection (Series II) Fly ash
Temperature (jC)
Hg on initial ash (Ag/g)
Hg on spent ash (Ag/g)
HgCl2 gen. (Ag)
Semtech mass balance (%)
Tekran mass balance (%)
Efficiency @ 100% breakthrough
Ohio Ohio Wyodak Wyodak Pittsburgh Pittsburgh Valmy Valmy AA DMA MA MA Absaloka Absaloka
121 177 121 177 121 177 121 177 177 121 121 177 121 177
0.276 0.276 1.29 1.29 0.14 0.140 0.557 0.557 0.282 0.051 0.060 0.060 0.223 0.223
0.50 0.88 2.90 2.12 0.63 0.26 2.52 0.60 NM 0.38 0.57 NM 0.93 0.37
24.2 42 23.5 45 20.1 27 22.4 24 31 26.7 28.7 NM 16.5 14.4
69 NMb 78 NM 74 69 86 82 NM 82 76 NM 86 71
70 90 84 85 78 66 87 76 NM 80 77 NM 80 72
6a 7a 7a 10a 10a NM 30a 2a 43c 8a 17a 44c 29a 8a
a b c
Mercury captured based on mercury on spent ash. NM = Not measured. Mercury captured based on integration of breakthrough curve.
Fly ash can catalyze the oxidation of elemental mercury and it can adsorb elemental mercury. Both processes were observed in the Series I data although not always simultaneously. Fig. 4 shows that not all ash samples tested oxidized or captured elemental mercury in the simulated flue gas mixtures. All samples that oxidized mercury also capture elemental mercury and higher levels of mercury oxidation were associated with higher levels of mercury capture (Fig. 5). Others [5,6] have hypothesized that elemental mercury is captured on activated carbon via chemisorption, not physisoption. We suggest that this is true of fly ash as well although there may be more than one mechanism by which capture
Fig. 4. Equilibrium capture of elemental mercury and steady-state mercury oxidation at 121 jC.
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Fig. 5. Equilibrium capture of elemental mercury as a function of steady-state oxidation.
takes place in the case of fly ash. A more fundamental look at the effect of flue gas composition and surface properties is needed to elucidate the mechanisms. The capacity of the ash for elemental mercury was similar to that for HgCl2 although there are only a few ash samples on which capacity measurements were made for both species of mercury. In Fig. 6, there is some correlation between the equilibrium capacity of the ash for elemental mercury and surface area, but there is considerable scatter in the data. This may reflect the influence of flue gas composition as well as surface properties on adsorption of elemental mercury. In contrast, there is a good correlation between the capacity of the ash samples for HgCl2 and the surface area (r2 = 0.8 for both 121 and 177 jC) as shown in Fig. 7. The capacity showed an inverse dependence on temperature, as had been noted in the literature for adsorption of HgCl2 on activated carbon [2]. The correlation between HgCl2 capacity and loss on ignition (LOI) in the ash was not as strong, suggesting that it is not the carbon content alone that influences capture of HgCl2, but also properties of the carbon such as the surface area.
Fig. 6. Equilibrium capacity (in Ag/g) for elemental mercury as a function of surface area at 121 jC.
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Fig. 7. Equilibrium capacity (Ag/g) for HgCl2 as a function of ash surface area.
As discussed in the introduction, elemental mercury appears to interact with the surface of fly ash (or activated carbon) at specific sites. Both carbon and iron oxides have previously been identified as components in coal fly ash that might interact with elemental mercury. Fig. 8 shows the effect of surface area on oxidation of elemental mercury. Temperature had little discernable effect on oxidation although there was considerable scatter in the data. The scatter in the data also suggested that surface area alone may not account for the oxidation of elemental mercury. For the bituminous coals, which have a significant amount of iron oxide in the ash, oxidation may be catalyzed by iron. Fig. 9 shows the relationship between magnetite content of the ash and mercury oxidation at 121 and 177 jC. Ash with a high fraction of
Fig. 8. Effect of surface area on elemental mercury oxidation.
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Fig. 9. Effect of magnetite content of bituminous ash samples on oxidation of elemental mercury at 121 and 177 jC.
the iron as magnetite oxidizes more mercury. The Wyodak ash is a notable outlier on this figure; the oxidation of elemental mercury by this ash may have been due to carbon, not to any inorganic constituent. It should be noted that previous bench-scale investigations of mercury – ash interactions [8,22] identified, using X-ray diffraction (XRD) analyses, maghemite as the active phase for mercury oxidation and capture in coal fly ash, whereas in the current work magnetite was identified, using Mo¨ssbauer analyses, as a potentially active phase for mercury oxidation. It becomes increasingly difficult to distinguish maghemite from magnetite using XRD, when both phases are substituted by other elements (e.g., Al, Mg, etc.), as is undoubtedly the case in coal-derived fly ash. Both magnetite and maghemite have essentially the same structure (spinel type) and give similar XRD patterns. The substitution of other elements in the spinel structre further blurs any XRD distinctions that you might be able to discern between the two phases. Such substitution is also recognized in the Mo¨ssbauer spectra as the spectral features deviate from that of pure magnetite or maghemite. Hence, calling this spinel phase either maghemite or magnetite is likely incorrect as both names are too precise, referring only to the iron in the phase. Certainly, there exist other elements in the spinel phase because phases intermediate between magnetite and maghemite can also exist; perhaps neither name should be used, but rather the phase should be described as a complex iron-rich spinel. However, if one relies only on XRD data, such fine distinctions are not possible and calling the phase maghemite is overly specific for the reasons already discussed. The Mo¨ssbauer data for the full-scale samples AA, DA, DMA, GA, and MA, for example, show that the complex spinel phase is more closely related to magnetite than maghemite. For these samples, there are at least two magnetite absorptions: one with an Isomer shift of f 0.3 mm/s and magnetic hyperfine splitting (H0) of f 490 kG and the other with an Isomer shift of f 0.6 mm/s and H0 of f 450 kG. This compares well with pure magnetite, which also has two absorptions: one with an IS of 0.30 and H0 of 489 kG and the other with and IS of 0.66 and H0 of 460 kG mm/s. (The corresponding values for maghemite are 0.3 mm/s, 498 kG and 0.36 mm/s, 475 kG.) For the other fly ash samples,
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the distinction is somewhat less clear as the spinel second absorption has IS values less than 0.60 mm/s.
4. Summary Ash samples from a variety of sources and coal types were tested in a fixed-bed reactor using elemental mercury or HgCl2 in simulated flue gas mixtures. The fly ash samples were characterized for surface area, loss on ignition, and forms of iron in the ash. While many of the ash samples oxidized elemental mercury, not all of the samples that oxidized mercury also captured elemental mercury. However, no capture of elemental mercury was observed without accompanying oxidation. In general, oxidation of elemental mercury increased with increasing amount of magnetite in the ash. However, one high-carbon subbituminous ash with no magnetite showed considerable mercury oxidation that may have been due to the carbon. This work and previous work suggest that an iron oxide with a spinel-type structure is active in fly ash with respect to mercury oxidation. Surface area as well as the nature of the surface appeared to be important for oxidation and adsorption of elemental mercury. The capacity of the ash samples for HgCl2 was similar to that for elemental mercury. There was a good correlation between the capacity for HgCl2 and the surface area; the capacity decreased with increasing temperature. The correlation between HgCl2 and LOI was not as strong, suggesting that it is not the carbon content alone but also properties of the ash, such as surface area, that influence capture of HgCl2.
Acknowledgements This work was supported by the U.S. Department of Energy, Contract DE-AC2295PC9501, the Electric Power Research Institute (EPRI), and the North Dakota Lignte Council. We are grateful to Dr. Frank Huggins at the University of Kentucky for Mo¨ssbauer analysis of selected samples and to Dr. John Veranth of the University of Utah for ash microscopy.
References [1] R.F. Afonso, C.L. Senior, Assessment of Mercury Emissions from Full-Scale Power Plants, Presented at the EPRI-EPA-DOE-AWMA Mega Symposium and Mercury Conference, Chicago, IL, Aug. 21 – 23, 2001. [2] D. Karatza, A. Lancia, D. Musmarra, F. Pepe, Adsorption of metallic mercury on activated carbon, Proceedings of the 26th Symposium (International) on Combustion, The Combustion Institute, Pittsburgh, PA, 1996, pp. 2439 – 2445. [3] T.R. Carey, O.W. Hargrove Jr., C.F. Richardson, R. Chang, F.B. Meserole, Factors affecting mercury control in utility flue gas using activated carbon, J. Air Waste Manage. Assoc. 48 (1998) 1166 – 1174. [4] G.E. Dunham, S.J. Miller, Effects of oxygen, moisture, and acid gases on mercury capture with carbon sorbents, Presented at the Air and Waste Management Association’s Mercury in the Environment Specialty Conference, Minneapolis, MN, Sep. 15 – 17, 1999.
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[5] M.J. Holmes, J.H. Pavlish, S.J. Miller, G.E. Dunham, Sorbent development for control of mercury emissions from utility power plants, Presented at the Air and Waste Management Association’s 95th Annual Conference, Baltimore, MD, June 24 – 28, 2002. [6] Y.H. Li, C.W. Lee, B.K. Gullett, Importance of activated carbon’s oxygen surface functional groups on elemental mercury adsorption, Fuel 82 (2003) 451 – 457. [7] F.E. Huggins, N. Yap, G.P. Huffman, C.L. Senior, Identification of mercury species in unburned carbon from pulverized coal combustion, Presented at the Air and Waste Management Association’s 92nd Annual Conference and Exhibition, St. Louis, MO, June 20 – 24, 1999. [8] S.J. Miller, E.S. Olson, G.E. Dunham, R.K. Sharma, Preparation methods and test protocol for mercury sorbents, Presented at the Air and Waste Management Association’s 91st Annual Meeting, San Diego, CA, June 14 – 18, 1998. [9] M.S. DeVito, W.A. Rosenhoover, Flue gas mercury and speciation studies at coal-fired utilities equipped with wet scrubbers, Presented at the 15th International Pittsburgh Coal Conference, Pittsburgh, PA, Sept. 15 – 17, 1998. [10] J. Butz, J. Smith, C. Grover, S. Haythornthwaite, M. Fox, T. Hunt, R. Chang, T.D. Brown, Coal fly ash as a sorbent for mercury, Proceedings of the Mercury in the Environment Specialty Conference, Air and Waste Management Association, Minneapolis, MN, Sept. 15 – 17, 1999. [11] F.E. Huggins, N. Yap, G.P. Huffman, Investigation of mercury adsorption on Cherokee fly-ash using XAFS spectroscopy, Presented at the Air and Waste Management Association’s 93rd Annual Conference and Exhibition, Salt Lake City, UT, June 19 – 22, 2000. [12] Z. Li, J.-Y. Hwang, Mercury distribution in fly ash components, Presented at the Air and Waste Management Association’s 90th Annual Conference and Exhibition, Toronto, Ontario, Canada, June 8 – 13, 1997. [13] C.L. Senior, S.A. Johnson, Impact of carbon-in-ash on mercury removal across particulate control devices in coal-fired power plants, Presented at the Air and Waste Management Association’s 95th Annual Conference and Exhibition, Baltimore, MD, June 23 – 27, 2002. [14] S. Sjostrom, C.J. Bustard, M. Durham, R. Chang, Mercury removal trends in full-scale ESPs and fabric filters, Presented at the EPRI-EPA-DOE-AWMA Mega Symposium and Mercury Conference, Chicago, IL, Aug. 21 – 23, 2001. [15] G.A. Norton, H. Yang, R.C. Brown, D.L. Laudal, G.E. Dunham, J. Erjavec, Heterogeneous oxidation of mercury in simulated post combustion conditions, Fuel 82 (2003) 107 – 116. [16] C.W. Lee, J.D. Kilgroe, S.B. Ghorishi, Speciation of mercury in the presence of coal and waste combustion fly ashes, Presented at the Air and Waste Management Association’s 93rd Annual Conference and Exhibition, Salt Lake City, UT, June 19 – 22, 2000. [17] Y. Zhuang, P. Biswas, M.E. Quintan, T.G. Lee, E. Arar, Kinetic study of adsorption and transformation of mercury on fly ash particles in an entrained flow reactor, Presented at the Air and Waste Management Association’s 93rd Annual Conference and Exhibition, Salt Lake City, UT, June 19 – 22, 2000. [18] W.S. Seames, J.O.L. Wendt, Partitioning of As, Se, and Cd during the combustion of Pittsburgh and Illinois no. 6 coals in a self-sustained combustor, Fuel Process. Technol. 63 (2 – 3) (2000) 179 – 196. [19] K.C. Galbreath, D.L. Toman, C.J. Zygarlicke, J.H. Pavlish, Trace element partitioning and transformations during pulverized coal combustion of bituminous and subbituminous U.S. coals in a 7-kW combustion system, Energy Fuels 14 (2000) 1265 – 1279. [20] F.E. Huggins, G.P. Huffman, in: C. Karr Jr. (Ed.), Analytical Methods for Coal and Coal Products, vol. III, Academic Press, Boston, MA, 1979, pp. 371 – 423. [21] S.J. Miller, G.E. Dunham, E.S. Olson, T.D. Brown, Mercury sorbent development for coal-fired boilers, Proceedings of the Conference on Air Quality-Mercury, Trace Elements and Particulate Matter, McLean, VA, Dec. 1 – 4, 1998. [22] C.J. Zygarlicke, Y. Zhuang, K.C. Galbreath, J.S. Thompson, M.J. Holmes, J.E. Tibbetts, R.L. Schulz, G.E. Dunham, Experimental investigations of mercury transformations in pilot-scale combustion systems and a bench-scale entrained-flow reactor, Proceedings of the International Conference on Air Quality-Mercury, Trace Elements and Particulate Matter, Arlington, VA, Dec. 10 – 12, 2002.
Fixed-bed studies of the interactions between mercury and coal combustion fly ash Grant E. Dunham a, Raymond A. DeWall a, Constance L. Senior b,* a b
Energy and Environmental Research Center, 15 North 23rd Street, Grand Forks, ND 58203, USA Reaction Engineering International, 77 West 200 South, Suite 210, Salt Lake City, UT 84101, USA
Abstract Sixteen different fly ash samples, generated from both pilot-scale and full-scale combustion systems, were exposed to a simulated flue gas containing either elemental mercury or HgCl2 in a bench-scale reactor system at the Energy and Environmental Research Center to evaluate the interactions and determine the effects of temperature, mercury species, and ash type on adsorption of mercury and oxidation of elemental mercury. The fly ash samples were characterized for surface area, loss on ignition, and forms of iron in the ash. While many of the ash samples oxidized elemental mercury, not all of the samples that oxidized mercury also captured elemental mercury. However, no capture of elemental mercury was observed without accompanying oxidation. Generally, oxidation of elemental mercury increased with increasing amount of magnetite in the ash. However, one high-carbon subbituminous ash with no magnetite showed considerable mercury oxidation that may have been due to unburned carbon. Surface area as well as the nature of the surface appeared to be important for oxidation and adsorption of elemental mercury. The capacity of the ash samples for HgCl2 was similar to that for elemental mercury. There was a good correlation between the capacity for HgCl2 and the surface area; capacity decreased with increasing temperature. D 2003 Elsevier Science B.V. All rights reserved. Keywords: Mercury control; Coal-fired power plant; Fly ash
1. Introduction On December 14, 2001, the U.S. Environmental Protection Agency (EPA) announced that it would regulate mercury emissions from coal-fired utility boilers with regulations being proposed in 2003 and implementation expected by 2007. Before electric utilities can implement emission minimization strategies for mercury, they must have an accurate * Corresponding author. 0378-3820/03/$ - see front matter D 2003 Elsevier Science B.V. All rights reserved. doi:10.1016/S0378-3820(03)00070-5
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means of predicting emissions in all effluent streams for the range of fuels and operating conditions commonly utilized. Results from sampling mercury from full-scale coal-fired boilers as part of EPA’s Information Collection Request (ICR) [1] indicate a wide range of mercury concentrations and splits between elemental, oxidized, and particulate-bound mercury. The existing air pollution control devices (APCDs) on a coal-fired boiler can remove a significant amount of mercury. The split between gaseous elemental mercury and gaseous oxidized mercury affects how much mercury is removed in wet scrubbers as well as how much mercury is adsorbed by fly ash. The partitioning between gaseous and particulate-bound mercury at the back end of a boiler determines how much mercury will be removed by the particulate control device (PCD). Ash plays a role in both the adsorption of mercury and the oxidation of elemental mercury in flue gas. Some insight into the mechanism or mechanisms can be gained by looking at the body of research concerning mercury and activated carbon sorbents. Both elemental mercury and HgCl2 are adsorbed by activated carbon at temperatures characteristic of the PCD. Activated carbon capacities at saturation are much less than monolayer coverage [2], which suggests that only certain sites on the surface are active for mercury adsorption. The equilibrium capacity is a function of the concentration of mercury in the gas [2,3] and has an inverse dependence on temperature [2]. The adsorption of HgCl2 on activated carbon is not greatly affected by flue gas composition [3,4]. The adsorption of Hg0, however, is highly dependent on gas composition and involves interaction with the acid gases (NOx, SOx, HCl). Based on fixed-bed studies, researchers have concluded that the adsorption of elemental mercury on activated carbon occurs by chemisorption rather than physisorption [5,6]. The result of adsorption of Hg0 on carbon is an oxidized compound (and not elemental mercury) on the surface of the carbon, based on X-ray absorption fine structure spectroscopy (XAFS) [7] and desorption studies [2,8]. Unburned carbon has been suspected of adsorbing mercury for both eastern and western bituminous and subbituminous coals. Often, as a consequence of low-NOx burners or lowNOx combustion systems, pulverized coal boilers can produce high levels of unburned carbon when burning bituminous coals [9] or, less commonly, subbituminous coals [10]. Mercury has been found to concentrate in the carbon-rich fraction of fly ash [11,12]. Recently, data on mercury in ash and carbon content were collected from sampling campaigns on full-scale coal-fired power plants [13]. Data from nine different plants were examined. Surprisingly good correlations of Hg removal with C/Hg ratio were observed for a range of coal and boiler types. There was a clear effect of temperature in comparing data from hot-side electrostatic precipitators (ESPs) with that from cold-side ESPs. Thus, for bituminous coals, this suggested that mercury can be adsorbed by the unburned carbon in the ash. Unfortunately, the largest database of mercury speciation and removal in existing coalfired power plant APCDs, the EPA’s Information Collection Request (ICR) database, does not include any information on the carbon content of the ash. A recent attempt to estimate the carbon content from a subset of ICR data [14] concluded that increased carbon in the fly ash was positively correlated with increased mercury removal across cold-side ESPs, for both bituminous and subbituminous coal. However, it is not possible to generalize to all coal types and APCDs and conclude that high carbon in ash will always give high levels of particulate-bound mercury.
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Norton et al. [15] carried out fixed-bed tests on two fly ash samples taken from ESP hoppers in full-scale boilers. The ash samples were from a Powder River Basin (PRB) coal and a Pittsburgh seam bituminous coal. In spite of large differences in mineralogy, the two samples did not show large differences in their ability to oxidize elemental mercury. Acid gases (HCl, NO2, and SO2) promoted oxidation of elemental mercury. The presence of NO appeared to suppress oxidation. In the Pittsburgh seam bituminous coal, the magnetic portion (i.e., iron-rich) of the ash did not appear to be more catalytically active than the nonmagnetic phases; however, differences in surface area between the magnetic and nonmagnetic portions made it difficult for the authors to conclude this with certainty. Lee et al. [16] studied simulated ash mixtures and actual ash samples in another fixedbed experiment using simulated flue gas. They identified NOx and HCl as gases that were important for the oxidation of elemental mercury. Both copper oxide (CuO) and iron oxide (Fe2O3) were identified as active compounds for oxidation of elemental mercury in the simulated ash mixtures. In the presence of NOx, even alumina and silica became active in promoting mercury oxidation. Of the fly ash samples tested, a Pittsburgh seam bituminous ash had significant catalytic activity with respect to mercury oxidation, especially in the presence of NOx. The other fly ash samples were also more catalytically active in the presence of NOx, as compared to HCl. In an entrained flow reactor, using exhaust gas from coal combustion, Zhuang et al. [17] also observed that elemental mercury could be oxidized and adsorbed on the surface of fly ash particles. Experiments carried out with pure oxide particles in the entrained flow reactor showed that iron oxide was effective at oxidizing elemental mercury. Fixed-bed studies can provide insight into the mechanisms for mercury adsorption on fly ash. The specific goal of this work was to perform bench-scale tests to evaluate the interactions of mercury with ash and ash components at temperatures characteristic of cold-side particulate collection devices.
2. Experimental 2.1. Sample characterization Ash was collected from two pilot-scale rigs, the University of Arizona selfsustained combustor [18] and the University of North Dakota EERC particulate test combustor (PTC) [19]. The coals burned included eastern bituminous (Ohio blend, Pittsburgh), western subbituminous (Wyodak, Absaloka, Cordero Rojo), and North Dakota lignite. At the University of Arizona, a downflow combustor with a firing rate of 17 kW was used to generate ash samples. Combustion was self-sustained in the furnace, which was 6 m long and had an internal diameter of 0.15 m. Ash samples were collected at a temperature of 200 jC using a special collection system consisting of a heated filter. The filter was installed in a horizontal section between the exit of the hot zone and the baghouse. Because the filter was upstream of an ash trap, the largest ash particles might not have been collected.
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The Pittsburgh (Blacksville) bituminous coal was burned in the PTC, and the ash was collected at 180 jC in a fabric filter pulse-jet baghouse equipped with Ryton bags. Sampling with EPA Method 29 was conducted during this test. The total mercury concentration at the inlet of the baghouse was 6.9 with 0.7 Ag/m3 equivalent mercury on the fly ash. Based on the Method 29 results, 85% of the mercury in the flue gas was oxidized. However, this method has since been shown to have high bias for oxidized mercury. The Absaloka and Cordero Rojo subbituminous coals were also burned in the PTC, but the ash was collected in the advanced hybrid particulate collector (AHPC), which is a combined ESP and baghouse equipped with GORETEXR membrane filter bags. No gas-phase mercury sampling was associated with these ash samples. In addition to the pilot-scale ash samples, a number of ash samples were taken from the hoppers of particulate collection devices at full-scale power plants. Five ash samples (AA, MA, DMA, GA, DA) were taken from the ESP hoppers of full-scale power plants burning eastern bituminous coals [9]. Additional information on the source of these ashes is provided in Table 1. A full-scale ash sample obtained from the ESP hopper at Coal Creek station located in North Dakota was evaluated. This was the same parent coal as from the North Dakota lignite sample burned at the University of Arizona. The Coal Creek station had two tangentially pc-fired boilers with an ESP upstream of a wet scrubber. The Valmy station ash was from a station burning a western bituminous coal in two units with low-NOx burners. The ash was collected from Unit 2, which was equipped with a spray dryer followed by a reverse-gas baghouse. Based on the concentration of mercury in the coal, the estimated mercury flue gas concentration was 4.2 Ag/m3. The mercury concentration on the base ash was 0.557 Ag/g, which translates to roughly 3.8 Ag/m3 of mercury in the flue gas, assuming a particulate loading of 2.5 gr/scf (6.86 g/m3). Based on these concentrations, the ash captured roughly 91% of the mercury in the flue gas. Sampling at the station indicated nearly all of the mercury in the flue gas was captured by the fly ash. A sample of ash from the Comanche station (formerly Public Service of Colorado) was collected by ADA Technologies on January 23, 1998 when the plant was burning a PRB coal from the Belle Ayr Mine. The station had two combustion units. Unit 1 was a drybottom, tangential-fired boiler with no NOx controls. Unit 2 was a dry-bottom, wall-fired boiler using overfire air for NOx control. Both units were equipped with a fabric filter dust collector. It was not specified from which unit the sample was collected. According to information in the literature [21], the mercury removal across the reverse-gas baghouse was 34 –78% at 138 jC (280jF) and less than 30% at 165– 175 jC (330 –350jF). The Table 1 Characteristics of Eastern Bituminous coal/ash samples from full-scale power plants Parameter
GA
DMA
DA
AA
MA
Hg in coal (Ag/g) Cl in coal (Ag/g) Coal ash content (%) Hg removal by ash (%) Boiler type
0.18 1300 13.6 6.8 Pulverized Coal
0.09 1600 10.9 7.8 Pulverized Coal
0.12 1100 13.9 24.0 Pulverized Coal
0.09 1100 18.5 10.1 Cyclone
0.09 2300 10.1 42.4 Stoker
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concentration of mercury on the fly ash indicated a significant amount of mercury capture by the fly ash during combustion. This is consistent with the reported data that showed less than 2 Ag/m3 vapor-phase mercury in the flue gas. The mercury concentration in the coal indicated approximately 5.4 Ag/m3 total mercury in the flue gas. A western bituminous coal was burned at the Valmont plant in a dry-bottom, tangentialfired boiler equipped with low-NOx burners. Particulate control was achieved with a fabric filter baghouse. Ash samples were collected from this plant as part of ICR sampling activity. Very little mercury was present in the coal. The Black Thunder subbituminous coal was burned in the Big Stone plant in Milbank, SD. The unit was a cyclone-fired boiler with overfire air for NOx control. The ash was collected in a demonstration AHPC. No mercury sampling was associated with this sample. Loss on ignition (LOI) and surface area were measured for the ash samples; surface area was computed from the nitrogen Brunauer – Emmett – Teller (B.E.T.) isotherm equation. As a part of this program, three pilot-scale ashes (Ohio, Wyodak and North Dakota lignite) were also characterized by the University of Utah. The investigation of particle morphology was carried out with a scanning electron microscope. In addition to measuring the LOI and surface area, 57Fe Mo¨ssbauer spectroscopy was carried out by Dr. Frank Huggins at the University of Kentucky for selected samples in order to quantify the state of iron in the ash. This was done because previous work [16] has suggested that iron oxides in fly ash may be responsible for some portion of catalytic oxidation of mercury. Mo¨ssbauer spectroscopy is a powerful, element-specific technique for iron in complex materials. In the area of coal characterization, it has been used for identification of iron minerals, quantitative estimates of pyritic sulfur, oxidation and weathering studies, and conversion of iron minerals during coal combustion, liquefaction, and gasification. Details of Mo¨ssbauer spectroscopy as applied to coal and coal ash have been discussed elsewhere [20]. 2.2. Fixed-bed apparatus and procedures A bench-scale system has been developed at the EERC for evaluating mercury sorbents and fly ashes in a fixed-bed configuration. A detailed description of this system has been previously reported [21]. The system, diagrammed in Fig. 1, was composed of three subsystems: the mercury/gas delivery system, the fixed-bed system, and the mercury measurement system. Mass flow controllers were used to maintain constant gas flow for most of the flue gas constituents. The mercury and mercuric chloride permeation tubes were maintained at a constant temperature and continuously purged with nitrogen to ensure constant outlet mercury concentrations. The output from each of the permeation tubes was periodically verified by taking EPA Method 101A samples at the inlet of the fixed-bed system. A simulated flue gas was used for all tests. The flue gas compositions for some of the tests were based on the proximate/ultimate analysis data for the parent coal. The flue gas blends are presented in Table 2. The flue gas mixture used for the other ash samples was the baseline mix used for evaluating sorbents for mercury control (in previous work at the EERC) and is also presented in Table 2.
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Fig. 1. Fixed-bed reactor system.
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Table 2 Flue gas mixtures for bench-scale tests Flue gas component
Ohio blend
North Dakota lignite
Wyodak
Pittsburgh, AA, DA, DMA, GA, MA
Valmy, Comanche, Dale Valmont, Black Thunder, Cordero Rojo, Absaloka
O2 (%) CO2 (%) H2O (%) SO2 (ppmv) NO (ppmv) NO2 (ppmv) HCl (ppmv)
3.3 14.0 8.0 1900 300 15 60
3.0 14.0 14.7 840 300 15 5
3.0 14.0 12.0 300 300 15 5
4.0 14.0 6.0 1700 550 27 50
6.0 12.0 8.0 1600 300 20 50
The ash samples were evaluated in a fixed bed consisting of a Teflon-coated 2.5-in.diameter filter holder. A quartz filter nominally loaded with 1.5 g of fly ash made up the actual fixed bed. The fixed-bed assembly was maintained at the desired temperature inside an oven that could be controlled to F 1 K. A slipstream of the flue gas could be sent to the continuous mercury monitors (CMMs) either from the inlet or the outlet of the fixed bed. The mercury measurement system consisted of a conditioning/conversion system and several CMMs. The conditioning/conversion system conditioned the sample gas by removing moisture and acid gases before sending it to the CMMs. In order to determine the amounts of elemental and oxidized forms of mercury, the conditioning/conversion system either reduced all forms of mercury to elemental mercury for a measurement of total mercury concentration or removed the oxidized forms of mercury from the flue gas stream for a measurement of elemental mercury concentration. The difference between the total mercury and Hg0 was assumed to be some form of Hg2 +. Three different CMMs were used to measure outlet mercury concentrations from the bench-scale system: a Semtech 2000, a PS Analytical Sir Galahad, and a Tekran Model 2537A. Fig. 2 is a plot of the outlet mercury concentration as a function of time from one run where all three instruments were used to measure the mercury concentration. The agreement among the instruments was very good. The results from the tests are presented as mercury concentration (normalized to a percent of the inlet mercury concentration) as a function of time. In addition to having good agreement between instruments, the results from the system have been shown to be very repeatable [21]. The test protocol for each run was essentially the same. At the start of each test, the concentration of total mercury at the outlet of the fixed bed was measured until 100% breakthrough was reached (outlet mercury concentration equals the inlet mercury concentration). At this point, the conditioning/conversion system was switched to measure the elemental mercury concentration at the outlet of the fixed bed (noted on the plots as Hg0). If no oxidation was taking place, the concentration would remain at 100% of the inlet elemental mercury concentration; however, if any of the elemental mercury was being oxidized across the filter, it would result in a drop in the outlet concentration of elemental mercury. After establishing the outlet elemental mercury concentration, the system was switched to measure the total inlet mercury concentration. When measuring the elemental mercury concentration at the outlet of the fixed bed while injecting HgCl2, the expected
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Fig. 2. Example of mercury measurements from three continuous mercury analyzers.
concentration was zero. If it was higher than zero, this indicated that some of the HgCl2 had been reduced to elemental mercury.
3. Results 3.1. Sample characterization Table 3 shows the characteristics of the pilot-scale ash. The bituminous ash samples had approximately 5 wt.% LOI and surface areas of 1.5 and 2.0 m2/g. The Wyodak ash had a very high LOI and surface area not typical of low-rank coals burned at full scale. The other two subbituminous ashes had more typical properties. SEM analyses of three of the pilotscale ash samples showed that the Ohio and North Dakota ash samples contained large, irregular-shaped particles, and no ultrafine aggregates, whereas the Wyodak sample was found to contain a lot of ultrafine aggregates, as shown in Fig. 3. The surface area information and the SEM pictures suggested that the Wyodak sample contained some soot and that the other two pilot-scale samples from the University of Arizona combustor did Table 3 Characteristics of pilot-scale ash Coal
Coal rank
Test rig
LOI, dry (wt.%)
SA (m2/g)
Hg, as received (Ag/g)
Ohio blend ND lignite Wyodak Pittsburgh Absaloka Cordero Rojo
Bituminous Lignite Subbituminous Bituminous Subbituminous Subbituminous
Arizona Arizona Arizona EERC PTC EERC PTC EERC PTC
6.79 2.86 4.45 4.3 1.34 0.21
1.4 0.68 9.33 1.95 3.04 0.74
0.276 0.536 1.29 0.14 0.223 0.183
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Fig. 3. SEM photo of Wyodak ash from University of Arizona combustor (24,900 times magnification).
not. The presence of soot would account for the extraordinarily high surface area of the Wyodak ash and was probably the result of the specific pilot-scale combustion system in which the ash was generated. While such high carbon (and surface area) is not typical of ash from full-scale systems burning subbituminous coal, this unusual ash makes for an interesting study on the potential for soot to affect mercury in flue gas. Table 4 provides a summary of the properties of the full-scale ash samples. The LOI of the bituminous ashes spanned a wide range from 1 to 45 wt.%, which was probably indicative of different firing systems and coal rank. Some of these ash samples also had very high surface areas. The LOI of the western ashes (bituminous and subbituminous) Table 4 Characteristics of full-scale ash Fly ash
Coal rank
LOI, dry (%)
Surface area (m2/g)
Mercury, Ag of Hg/g of ash
Comanche Valmy Coal Creek Valmont Black Thunder AA MA DA GA DMA
Subbituminous Western bituminous Lignite Western bituminous Subbituminous Eastern bituminous Eastern bituminous Eastern bituminous Eastern bituminous Eastern bituminous
0.90 4.50 0.01 0.88 1.38 37.12 44.4 6.22 0.67 1.5
1.35 5.98 NMa 1.49 3.0 4.36 3.92 6.17 0.17 0.96
1.05 0.557 0.0067 NM 2.13 0.282 0.060 0.114 0.002 0.051
a
NM = Not measured.
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was less than 5 wt.%. The surface areas of these ashes were between 1 and 6 m2/g, with the exception of the ash from the Coal Creek plant. In general, the full-scale samples had a wide range of carbon contents (measured as LOI); some of the ash from bituminous coals had very high carbon. The surface area of the full-scale ash samples did not appear to follow the carbon content. For example, ash samples AA and MA had very high LOI values, but the surface area of these samples was not commensurately high. The Mo¨ssbauer data (Table 5) indicate that the iron in all the samples was present in either glass or iron oxide phases. Two different subsets of samples were identified depending on whether or not the calcium ferrite phases were significant. In the bituminous ash samples, iron was found in oxide phases or in glassy phases, whereas in the subbituminous ash samples, the iron was predominantly calcium ferrite, with some hematite (iron oxide), presumably coming from pyrite in the coal. 3.2. Fixed-bed results In the first series of fixed-bed experiments, ash samples were exposed to elemental mercury in simulated flue gas at temperatures of 121 and 177 jC (250 and 350 jF). Both oxidation and adsorption of elemental mercury were calculated from the CMM data as a function of time. In the second series, ash samples were exposed to simulated flue gas containing HgCl2. Tables 6 and 7 apply to Series I and II, respectively. Each table includes the concentration of mercury on the spent ash (if it was submitted for analysis), the amount of mercury generated during the test, the average mass balance based on data from the CMMs used during the test, the percentage of the elemental mercury that was oxidized across the fixed bed, and the mercury collection efficiency of the ash at 100% breakthrough. The mass balance is the ratio of the mercury measured (on the ash and in the flue gas) to the mercury generated (based on the average mercury permeation rate determined with wet chemistry measurements). The collection efficiency at complete breakthrough is based on the mercury on the spent ash or the integrated area under the Table 5 Mo¨ssbauer results (wt.% Fe in each phase) Sample
Hematite
Magnetite
Fe+ 3 oxide
Fe+ 3 glass
Ohio (pilot) Wyodak (pilot) North Dakota lignite (pilot) Pittsburgh (pilot) Comanche (full scale) Valmont (full scale) Valmy (full scale) Ash GA (full scale) Ash AA (full scale) Ash DMA (full scale) Ash DA (full scale) Ash MA (full scale)
23 17 19 10
51.5
2
24
69
20
5
28
47
48.5 56.5 25 60.5 43
25.5 9 39 18 30
Fe+ 2 glass
Ca ferrite 83 82 93
15.5 8 10.5 7
Mag. Ca ferrite
19 13 10 26.5 36 11 19.5
87
7
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Table 6 Results from tests with Hg0 injection (Series I) Fly ash
Temperature Hg0 on initial Hg0 on Hg0 gen. Average mass Percent Hg0 Collection (jC) ash (Ag/g) spent ash (Ag) balance (%) oxidized efficiency (Ag/g) @ 100% breakthrough
Ohio Ohio Wyodak Wyodak N D lignite N D lignite Coal Creek Pittsburgh Pittsburgh Valmy Valmy Valmy Valmy Valmont Valmont Comanche Comanche Comanche AA AA AA DA DMA GA MA MA MA Black Thunder Black Thunder Codero Rojo Codero Rojo Absaloka Absaloka
121 177 121 177 121 177 177 121 177 121 121 121 177 121 177 121 177 177 121 121 177 121 121 121 121 177 121 121 177 121 177 121 177
a b c
0.276 0.276 1.29 1.29 0.536 0.536 0.0067 0.140 0.140 0.557 0.557 0.557 0.557 NM NM 1.05 1.05 1.05 0.282 0.282 0.282 0.114 0.051 0.002 0.060 0.060 0.060 2.13 2.13 0.183 0.183 0.223 0.223
NM NM 2.51 NM NM NM NM 1.21 NM NM 1.40 1.67 NM NM 0.09 NM 0.40 NM 0.93 1.28 NM NM NM NM 0.87 0.20 0.74 NM NM NM NM 2.33 0.51
NM NM 43.8 NM NM NM NM 30.6 NM NM 28.8 24.1 NM NM 24.3 NM 18.2 NM 26.2 33.7 NM NM NM NM 21.5 22.9 23.3 NM NM NM NM 27.3 16.9
NMa NM 102 NM NM NM NM 91 NM NM 82 113 NM NM 104 NM 118 NM 75 109 NM NM NM NM 111 104 108 NM NM NM NM 88 94
0 20 40 60 0 0 0 45 25 50 50 50 40 70 30 25 15 15 85 85 65 55 20 15 80 80 60 40 30 20 20 70 70
NM NM 3b 42c NM NM NM 5b NM NM NM 14b NM NM 3b NM 0b NM 6b 10b NM 64c NM NM 20b NM 8b NM NM NM NM 17a 3a
NM = Not measured. Mercury captured based on mercury on spent ash. Mercury captured based on integration of breakthrough curve.
curve (if ash analysis is not available) and the amount of mercury generated (permeation rate) in the time it took to reach 100% breakthrough. It appears the collection efficiencies based on the integrated breakthrough curve are higher than those based on ash analysis. The difference may be caused by offgassing of collected mercury during the remainder of the test, which is typical of similar tests with carbon-based sorbents [4]. However, the results do give a basis for comparison of ashes and their ability to capture mercury. For many samples, the collection efficiency was zero, and the ash showed no ability to capture mercury.
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Table 7 Results from tests with HgCl2 injection (Series II) Fly ash
Temperature (jC)
Hg on initial ash (Ag/g)
Hg on spent ash (Ag/g)
HgCl2 gen. (Ag)
Semtech mass balance (%)
Tekran mass balance (%)
Efficiency @ 100% breakthrough
Ohio Ohio Wyodak Wyodak Pittsburgh Pittsburgh Valmy Valmy AA DMA MA MA Absaloka Absaloka
121 177 121 177 121 177 121 177 177 121 121 177 121 177
0.276 0.276 1.29 1.29 0.14 0.140 0.557 0.557 0.282 0.051 0.060 0.060 0.223 0.223
0.50 0.88 2.90 2.12 0.63 0.26 2.52 0.60 NM 0.38 0.57 NM 0.93 0.37
24.2 42 23.5 45 20.1 27 22.4 24 31 26.7 28.7 NM 16.5 14.4
69 NMb 78 NM 74 69 86 82 NM 82 76 NM 86 71
70 90 84 85 78 66 87 76 NM 80 77 NM 80 72
6a 7a 7a 10a 10a NM 30a 2a 43c 8a 17a 44c 29a 8a
a b c
Mercury captured based on mercury on spent ash. NM = Not measured. Mercury captured based on integration of breakthrough curve.
Fly ash can catalyze the oxidation of elemental mercury and it can adsorb elemental mercury. Both processes were observed in the Series I data although not always simultaneously. Fig. 4 shows that not all ash samples tested oxidized or captured elemental mercury in the simulated flue gas mixtures. All samples that oxidized mercury also capture elemental mercury and higher levels of mercury oxidation were associated with higher levels of mercury capture (Fig. 5). Others [5,6] have hypothesized that elemental mercury is captured on activated carbon via chemisorption, not physisoption. We suggest that this is true of fly ash as well although there may be more than one mechanism by which capture
Fig. 4. Equilibrium capture of elemental mercury and steady-state mercury oxidation at 121 jC.
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Fig. 5. Equilibrium capture of elemental mercury as a function of steady-state oxidation.
takes place in the case of fly ash. A more fundamental look at the effect of flue gas composition and surface properties is needed to elucidate the mechanisms. The capacity of the ash for elemental mercury was similar to that for HgCl2 although there are only a few ash samples on which capacity measurements were made for both species of mercury. In Fig. 6, there is some correlation between the equilibrium capacity of the ash for elemental mercury and surface area, but there is considerable scatter in the data. This may reflect the influence of flue gas composition as well as surface properties on adsorption of elemental mercury. In contrast, there is a good correlation between the capacity of the ash samples for HgCl2 and the surface area (r2 = 0.8 for both 121 and 177 jC) as shown in Fig. 7. The capacity showed an inverse dependence on temperature, as had been noted in the literature for adsorption of HgCl2 on activated carbon [2]. The correlation between HgCl2 capacity and loss on ignition (LOI) in the ash was not as strong, suggesting that it is not the carbon content alone that influences capture of HgCl2, but also properties of the carbon such as the surface area.
Fig. 6. Equilibrium capacity (in Ag/g) for elemental mercury as a function of surface area at 121 jC.
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Fig. 7. Equilibrium capacity (Ag/g) for HgCl2 as a function of ash surface area.
As discussed in the introduction, elemental mercury appears to interact with the surface of fly ash (or activated carbon) at specific sites. Both carbon and iron oxides have previously been identified as components in coal fly ash that might interact with elemental mercury. Fig. 8 shows the effect of surface area on oxidation of elemental mercury. Temperature had little discernable effect on oxidation although there was considerable scatter in the data. The scatter in the data also suggested that surface area alone may not account for the oxidation of elemental mercury. For the bituminous coals, which have a significant amount of iron oxide in the ash, oxidation may be catalyzed by iron. Fig. 9 shows the relationship between magnetite content of the ash and mercury oxidation at 121 and 177 jC. Ash with a high fraction of
Fig. 8. Effect of surface area on elemental mercury oxidation.
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Fig. 9. Effect of magnetite content of bituminous ash samples on oxidation of elemental mercury at 121 and 177 jC.
the iron as magnetite oxidizes more mercury. The Wyodak ash is a notable outlier on this figure; the oxidation of elemental mercury by this ash may have been due to carbon, not to any inorganic constituent. It should be noted that previous bench-scale investigations of mercury – ash interactions [8,22] identified, using X-ray diffraction (XRD) analyses, maghemite as the active phase for mercury oxidation and capture in coal fly ash, whereas in the current work magnetite was identified, using Mo¨ssbauer analyses, as a potentially active phase for mercury oxidation. It becomes increasingly difficult to distinguish maghemite from magnetite using XRD, when both phases are substituted by other elements (e.g., Al, Mg, etc.), as is undoubtedly the case in coal-derived fly ash. Both magnetite and maghemite have essentially the same structure (spinel type) and give similar XRD patterns. The substitution of other elements in the spinel structre further blurs any XRD distinctions that you might be able to discern between the two phases. Such substitution is also recognized in the Mo¨ssbauer spectra as the spectral features deviate from that of pure magnetite or maghemite. Hence, calling this spinel phase either maghemite or magnetite is likely incorrect as both names are too precise, referring only to the iron in the phase. Certainly, there exist other elements in the spinel phase because phases intermediate between magnetite and maghemite can also exist; perhaps neither name should be used, but rather the phase should be described as a complex iron-rich spinel. However, if one relies only on XRD data, such fine distinctions are not possible and calling the phase maghemite is overly specific for the reasons already discussed. The Mo¨ssbauer data for the full-scale samples AA, DA, DMA, GA, and MA, for example, show that the complex spinel phase is more closely related to magnetite than maghemite. For these samples, there are at least two magnetite absorptions: one with an Isomer shift of f 0.3 mm/s and magnetic hyperfine splitting (H0) of f 490 kG and the other with an Isomer shift of f 0.6 mm/s and H0 of f 450 kG. This compares well with pure magnetite, which also has two absorptions: one with an IS of 0.30 and H0 of 489 kG and the other with and IS of 0.66 and H0 of 460 kG mm/s. (The corresponding values for maghemite are 0.3 mm/s, 498 kG and 0.36 mm/s, 475 kG.) For the other fly ash samples,
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the distinction is somewhat less clear as the spinel second absorption has IS values less than 0.60 mm/s.
4. Summary Ash samples from a variety of sources and coal types were tested in a fixed-bed reactor using elemental mercury or HgCl2 in simulated flue gas mixtures. The fly ash samples were characterized for surface area, loss on ignition, and forms of iron in the ash. While many of the ash samples oxidized elemental mercury, not all of the samples that oxidized mercury also captured elemental mercury. However, no capture of elemental mercury was observed without accompanying oxidation. In general, oxidation of elemental mercury increased with increasing amount of magnetite in the ash. However, one high-carbon subbituminous ash with no magnetite showed considerable mercury oxidation that may have been due to the carbon. This work and previous work suggest that an iron oxide with a spinel-type structure is active in fly ash with respect to mercury oxidation. Surface area as well as the nature of the surface appeared to be important for oxidation and adsorption of elemental mercury. The capacity of the ash samples for HgCl2 was similar to that for elemental mercury. There was a good correlation between the capacity for HgCl2 and the surface area; the capacity decreased with increasing temperature. The correlation between HgCl2 and LOI was not as strong, suggesting that it is not the carbon content alone but also properties of the ash, such as surface area, that influence capture of HgCl2.
Acknowledgements This work was supported by the U.S. Department of Energy, Contract DE-AC2295PC9501, the Electric Power Research Institute (EPRI), and the North Dakota Lignte Council. We are grateful to Dr. Frank Huggins at the University of Kentucky for Mo¨ssbauer analysis of selected samples and to Dr. John Veranth of the University of Utah for ash microscopy.
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