The role of ammonia on mercury leaching from coal fly ash

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Chemosphere 69 (2007) 1586–1592 www.elsevier.com/locate/chemosphere

The role of ammonia on mercury leaching from coal fly ash Jianmin Wang a

a,*

, Tian Wang a, Harmanjit Mallhi a, Yu Liu a, Heng Ban b, Ken Ladwig

c

Department of Civil, Architectural and Environmental Engineering, University of Missouri-Rolla, Rolla, MO 65409, United States b Department of Mechanical and Aerospace Engineering, Utah State University, Logan, UT 84322, United States c Electric Power Research Institute (EPRI), 3420 Hillview Avenue, Palo Alto, CA 94304, United States Received 26 February 2007; received in revised form 21 May 2007; accepted 22 May 2007 Available online 2 July 2007

Abstract The Federal Clean Air Interstate Rule issued in March 2005 will result in many power plants employing ammonia-based technologies to control NOx emission. The Clean Air Mercury Rule, issued at the same time, will encourage many power plants to use various technologies to remove mercury from flue gas, generating fly ashes that contain elevated concentrations of mercury. Ammonia forms relatively strong complexes with mercury compared to most other cationic elements and, therefore, may change the leaching characteristics of mercury. Understanding the impact of ammonia on the leaching of mercury from fly ash is critical in predicting the potential environmental impact of future fly ash. Batch methods were used to investigate the ammonia impact on mercury leaching from fly ash under different pH conditions. The results indicated that mercury leaching without external ammonia addition is not significant. However, ammonia addition increased mercury leaching in the alkaline pH range, due to the formation of less adsorbable mercury–ammonia complexes. Washed ash released more mercury than the raw ash if the ammonia concentration is the same, mainly due to the dissolution of some ash components during washing which exposed more mercury on ash surface. Mercury adsorption data indicated that more than 90% of available mercury was adsorbed by fly ash even in the presence of 1000 mg l1 ammonia addition.  2007 Elsevier Ltd. All rights reserved. Keywords: Fly ash; Mercury; Leaching; Ammonia

1. Introduction The US EPA issued two air pollutant emission control regulations in March of 2005, the Clean Air Interstate Rule (CAIR) and the Clean Air Mercury Rule (CAMR) (EPA, 2005). CAIR permanently caps emissions of NOx and SOx from large stationary sources including coal-fired plants in the eastern United States. It is expected that more than 55% of the coal-fired power generating capacity will install ammonia-based technologies, either selective catalytic reduction (SCR) or selective non-catalytic reduction (SNCR), to control NOx emission (Chu et al., 2001). Ammonia is also commonly used as a flue gas conditioner to enhance the fly ash capture by electrostatic precipitators (ESP). These implementations will result in fly ashes that *

Corresponding author. Tel.: +1 573 341 7503; fax: +1 573 341 4729. E-mail address: [email protected] (J. Wang).

0045-6535/$ - see front matter  2007 Elsevier Ltd. All rights reserved. doi:10.1016/j.chemosphere.2007.05.053

contain ammonia. The ammonia-based NOx controls can also increase the oxidation of elemental mercury therefore enhance the mercury capture by fly ash (Miller et al., 2006). CAMR requires enhanced capture of mercury from flue gas. Various approaches to reduce mercury air emission are currently being tested, including injection of powdered activated carbon prior to primary particulate collection. This will increase the mercury content in fly ash. According to American Coal Ash Association (ACAA) statistics, US electric utilities generated 123 million short tons (1.12 · 1011 kg) of coal combustion products (CCPs) in 2005 (ACAA, 2006). Fly ash accounted for about 58% of the total CCPs, or about 71 million short tons (6.4 · 1010 kg). Approximately 41% of the fly ash was beneficially used in cement and concrete industries, in waste stabilization, as a structural fill, and in other applications, while the remaining 59% was primarily managed in landfills or ponds. Depending on the source of the coal, fly

J. Wang et al. / Chemosphere 69 (2007) 1586–1592

ash contains various levels of trace elements such as arsenic, barium, boron, cadmium, chromium, cobalt, copper, lead, mercury, and selenium (Kim and Cardone, 1997; Querol et al., 1999; Kim and Kazonich, 2001). Mercury is of special interest because of its toxicity and bioaccumulative nature (EPA, 2001), and the likelihood that concentrations in fly ash will increase due to new mercury control regulations. Recent studies indicated that mercury leaching from ammonia-free, traditional or carbon-injected fly ash is not a significant environmental concern (EPRI, 1999, 2001a, 2002; Gustin and Ladwig, 2004; EPA, 2006; Xin et al., 2006a). However, both theoretical and experimental studies indicated that ammonia can form complexes with many cationic metals including mercury (Stumm and Morgan, 1996; Teng et al., 2003a,b; Wang et al., 2003a,b; Wang et al., 2006). Therefore, the presence of ammonia may change the mercury leaching characteristics in the fly ash. It is important to understand the role of ammonia on mercury leaching in order to predict the potential environmental impact of future fly ash.

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in an oven at 103–105 C until completely dry (normally 24 h) prior to experiments. 2.2. Ash characterization

2. Materials and methods

Important characteristics of fly ash related to mercury leaching are loss-on-ignition (LOI), specific surface area, total mercury concentration, and total ammonia concentration. LOI was used as an indicator for unburned carbon, and was determined using a gravimetric method based on the weight loss at 550 C for samples dried at 105 C. The specific surface area (BET area) was determined using a Quantachrome Autosorb-1-C high performance surface area and pore size analyzer. For the determination of total mercury in fly ash, the samples were first digested according to EPA method 3052 procedure. The total dissolved mercury concentration was then determined using a Tekran Series 2600 Ultra-trace Mercury Analysis System based on EPA method 1631. The total soluble ammonia concentration in the fly ash leachate was determined using a Nitrogen–Ammonia Reagent Set, Test ’N Tube, Salicylate Method, 0.4–50 mg l1, with estimated detection limit of 0.4 mg l1 (Hach Comp., USA).

2.1. Ash samples and preparation

2.3. Batch equilibrium experiments

Three ash samples collected from electrostatic precipitators at two power plants burning eastern bituminous coals were selected for this research. Sample 33106-1008 (Ash #1008) was collected from a plant that was using SNCR technology to control NOx emission, but did not have any special mercury control. Sample 33104-85 (Ash #85) was collected from second plant burning a different eastern bituminous coal, and using brominated powered activated carbon injection technology to control mercury emission. The second plant did not have any NOx control. Sample 33104-86 (Ash #86) was collected from the same unit burning the same coal as that for Ash #85, but without activated carbon injection. Batch leaching experiments were conducted under different pH and ammonia additions using raw ash samples. In field-management scenarios, chemical weathering occurs when ash is sluiced to a pond (low S/L ratio) or precipitation infiltrates through landfilled ash (high S/L ratio). In a landfill environment, when fresh ash with ammonia is placed on top of partially weathered ash, ammonia can be quickly released to water, which may enhance mercury leaching from the weathered ash below. Batch experiments were conducted using washed ash to simulate the impact of ammonia on mercury leaching from weathered ashes. Deionized (DI) water was used to wash the ash sample. A batch method with a solid-to-liquid (S/L) ratio (by weight) of 1:5 was used for ash washing, and air was used to mix the ash–water mixture during the washing process. Each washing cycle lasted for approximately 24 h. Depending on the purpose of the experiment, different washing cycles were applied to the fly ash. Ash samples were heated

Batch methods were used to determine the impact of ammonia on mercury leaching or adsorption under different pH conditions. For the leaching experiment, raw ash was used without external mercury addition. For the adsorption experiment, either raw ash or washed ash was used, and external mercury was added to the system. Therefore, DI water or ammonium nitrate (NH4NO3) solution was used as leachants for leaching experiments. Solutions containing both ammonium nitrate and mercury were used as leachants for adsorption experiments. Ten grams of dry ash was mixed with 100 ml of leaching solution (S/L = 1:10 by weight) in a series of 125 ml HDPE bottles. One or 5 M stock HNO3 or NaOH solution was added to these bottles to adjust pH, except for the control unit to which no acid or base was added. The bottles were then sealed and shaken at 180 oscillation/min using a EBERBACH 6010 shaker for 24 h to achieve equilibrium (EPRI, 2005). After shaking, all bottles were allowed to settle for at least 12 h, and supernatants were collected for mercury analyses. The supernatant samples were not filtered to avoid the potential adsorption loss of mercury by the filter paper. The final pH values were determined using the remaining contents in the bottle, using an Orion pH electrode (model 9207BN) and an Orion pH meter (perpHecT LogR model 370). 2.4. Mercury analysis A Tekran Series 2600 Ultra-trace Mercury Analysis System (Tekran Inc., Toronto, Canada) was used to determine the mercury concentration in all liquid samples.

J. Wang et al. / Chemosphere 69 (2007) 1586–1592

Tekran Series 2600 is characterized by dual stage gold pre-concentration followed by Cold Vapor Atomic Fluorescence Spectrophotometer (CVAFS) as detector. Samples and standards were pre-digested with 5 ml/l BrCl stock solution for 12 h as required by EPA Method 1631. The BrCl stock solution was prepared with 11 g KBr + 15 g KBrO3 + 200 ml H2O + 800 ml hydrochloric acid solution. After oxidation, the sample was reduced with NH2OH Æ HCl to destroy free halogens. NH2OH Æ HCl was prepared by dissolving 75 g NH2OH Æ HCl in 250 ml DI water, then purified by adding 0.125 ml 3% SnCl2 reducing solution and purging using ultra-high purity argon for at least 2 h. The sample was then reduced with SnCl2 to convert Hg(II) to volatile Hg(0). The Hg(0) was separated from solution by purging with mercury-free argon onto a gold coated sand trap and then a gold trap. The trapped Hg was then thermally desorbed into the argon gas stream for CVAFS detection.

3. Results and discussion 3.1. Fly ash characteristics Table 1 lists the major characteristics of fly ash samples. Ash #85 and Ash #86 (generated from the same coal) contained significantly greater amount of unburned carbon than those ashes we normally see in the field. The total unburned carbon content for Ash #86 (without carbon injection) was 27%, greater than that for Ash #85 (with carbon injection) of 20%, suggesting that the combustion condition played a more important role than activated carbon injection in the final carbon content in fly ash. The specific surface area results indicated that, among the three samples studied, Ash #1008 had the lowest specific surface area of 6.5 m2 g1, while the Ash #85 had the greatest one of 30 m2 g1. The significantly greater specific surface area for Ash #85 than Ash #86 (21 m2 g1) was likely caused by the high specific surface area associated with the injected activated carbon. For the two ashes that do not contain injected activated carbon, the specific surface area was proportional to their LOI. Ash #1008 had the greatest total mercury concentration, while Ash #85 and Ash #86 had similar but lower concentrations. Interestingly, the injection of activated carbon did not appear to substantially affect the total mercury concentration in Ash #85 compared to Ash #86. This result is not

consistent with other studies, which has generally shown a significant increase in total mercury concentration in ash collected with activated carbon (EPRI, 1999, 2002; Gustin and Ladwig, 2004). The ammonia concentrations were determined from raw ash (i.e. unwashed ash) leachates obtained through batch experiments under different pH conditions. The pH did not impact the leachate ammonia concentration, indicating both NH3 and NHþ 4 are highly soluble. For an S/L ratio of 1:10, the average leachate ammonia concentration for Ash #1008 was 50 mg l1; the presence of ammonia was due to the SNCR process used at the plant. The ammonia concentrations in Ash #85 and Ash #86 leachates were negligible. 3.2. Impact of added ammonia on mercury leaching from raw ash Fig. 1 shows the batch leaching results for raw Ash #1008 as a function of pH under added ammonia concentrations ranging from 100 to 10 000 mg l1. These concentrations, which were 2–200 times greater than the raw ash leaching concentration for the same condition, were added to induce a response in the leaching behavior rather than simulate expected field conditions. The data results from high ammonia concentration can provide mechanistic information for better understanding of a leaching process. In addition, a column leaching study indicated that ammonia is rapidly released from fresh ash upon contact with water, and the dissolved ammonia concentration in leachates can be seven times more than that in the ash (in ppmw, 4000 0 mg/l NH3 100 mg/l NH3 1000 mg/l NH3 5000 mg/l NH3 10000 mg/l NH3

3500 3000

Hg (ng/l)

1588

2500 2000

Raw Ash #1008

1500 1000 500 0 2

4

6

8 pH

10

12

14

Fig. 1. Mercury release from raw Ash #1008 under different added ammonia concentrations. Experimental conditions: temperature = 20– 25 C; S/L = 1:10; equilibration time = 24 h.

Table 1 Sample characteristics Sample ID

Emission control

LOI (%)

BET area (m2 g1)

Total Hg (mg g1 ash)

Soluble NH3 at S/L = 1:10 (mg l1)

Ash #1008 Ash #85 Ash #86

SNCR (for NOx) ACI (for Hg) None

8.5 20 27

6.5 30 21

0.98 0.22 0.19

50 ND ND

SNCR: selective non-catalytic reduction. ACI: activated carbon injection. ND: non-detectable.

J. Wang et al. / Chemosphere 69 (2007) 1586–1592

parts per million by weight) under high S/L ratio landfill conditions (EPRI, 2001b). Therefore, ammonia concentrations of greater than 50 mg l1 in some field leachates are possible. A control experiment was also conducted using DI water as a leachant (no ammonia addition). For all added ammonia concentrations, mercury release was negligible when pH was less than 6 and greater 12, and reached a maximum value at pH approximately 9. Mercury concentrations also increased with increasing ammonia concentration in the pH range between 6 and 12. This behavior was hypothesized to be caused by the formation of some mercury–ammonia complexes which are less adsorbable by fly ash than mercury-hydroxide species. When DI water was used as a leaching solution, a very low concentration of mercury (less than 150 ng/l) was released in the pH range between 9 and 10. This low level leaching reflects what might be expected under field conditions when large amount of water is used to carry ammoniated fly ash to an ash pond (50 mg l1 at the S/L ratio of 1:10 for Ash #1008 in this study). The total ammonia in the system include free ammonia (NH3) and ammonium ion ðNHþ 4 Þ, and only the free ammonia is responsible for the formation of mercury– ammonia complexes. Based on the formation constants of mercury-hydroxide mercury–ammonia complexes, and the acidity constant of ammonium ion, the mercury speciation as a function of pH in the presence of 1000 mg l1 ammonia was calculated, shown in Fig. 2. Mercury–ammonia complexes are the dominant species in the pH range between 2 and 10.5. Because the mercury release also increased in the pH range corresponding to the formation of HgðNH3 Þ2þ and HgðNH3 Þ2þ species (Figs. 1 and 2), 3 4 these two mercury–ammonia complexes maybe less adsorb2þ able than Hg(NH3)2+ and HgðNH3 Þ2 species. As shown in Fig. 2, when pH is greater than 10, the fraction of Hg(OH)2 increases. Since Hg(OH)2 has high affinity for many adsorbents (MacNaughton and James, 1974; Newton et al., 1976; Kinniburgh and Jackson, 1978), the formation of

Hg(OH)2 reduced mercury leaching, as indicated in Fig. 1. When pH reached 12, almost all mercury was adsorbed by fly ash due to the formation of the Hg(OH)2 species in the system. The impact of added ammonia on the leaching of mercury from Ash #85 and Ash #86 was also investigated (data not shown). The leaching patterns for both ashes were similar to that for Ash #1008, except that less mercury was released. Therefore, formation of mercury– ammonia complexes enhanced mercury leaching. In addition, for these two ashes, no mercury was detected in leachates across the entire pH range when DI water was used as the leaching solution. These results were consistent with previous studies that have shown little mercury leaching from traditional ash and ash that contains activated carbon (EPRI, 2002; Gustin and Ladwig, 2004). As indicated earlier, both Ash #85 and Ash #86 contained a relatively large amount of unburned carbon (LOIs were, respectively, 20% and 27%). Previous research suggested that unburned carbon has stronger adsorption strength toward mercury than other ash components (Hwang et al., 2002; Xin et al., 2006b). Therefore, high contents of unburned carbon may significantly contribute to the relatively low mercury release from these two ashes. 3.3. Impact of added ammonia on mercury leaching from washed ash Washed Ash #1008 with 5-washing cycles was used to simulate the leaching behavior of weathered ash under different pH and ammonia concentration conditions. The 5washing cycles may represent long term water contact in a pond environment, or hundreds to thousands of years in a landfill environment, depending on infiltration rate. Fig. 3 shows the batch leaching results for added ammonia concentrations ranging from 100 to 10 000 mg l1, a much higher concentration range than what would normally be expected. Results were similar to those for the raw ash,

1. 2

Hg(OH)2

Hg2+

1. 0

Hg(NH 3)22+

0. 8 Fraction

1589

0. 6 0. 4

Hg(NH 3)42+ Hg(OH)3-

Hg(NH 3)32+

Hg(NH 3) 2+

0. 2 0. 0 0

2

4

6 pH

8

10

12

14

Fig. 2. Calculated mercury speciation in 1000 mg l1 ammonia solution based on the following constants: the overall formation constants (log b) for 2þ 2þ 2þ 2+ Hg(OH)+, Hg(OH)2, HgðOHÞ 3 , Hg(NH3) , HgðNH3 Þ2 , HgðNH3 Þ3 , and HgðNH3 Þ4 are, respectively, 10.6, 21.8, 20.9, 8.8, 17.4, 18.4, and 19.1. The acidity constant of ammonium ion, pKa = 9.3 (Stumm and Morgan, 1996).

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J. Wang et al. / Chemosphere 69 (2007) 1586–1592 Table 2 Concentrations of mercury and TDS in DI water washing decants for Ash #85

8000 100 mg/l NH3 500 mg/l NH3 1000 mg/l NH3 5000 mg/l NH3 10000 mg/l NH3

7000

Hg (ng/l)

6000 5000

Washed Ash #1008

4000

Washing cycle

1

2

3

4

5

Hg in leachate (ng l1) TDS (mg l1)

2.0 1710

6.2 380

1.6 150

6.8 90

9.5 100

3000 2000 1000 0 2

4

6

8

10

12

14

pH Fig. 3. Impact of ammonia on mercury leaching from washed Ash #1008 (with 5-washing cycles). Experimental conditions: temperature = 20– 25 C; S/L = 1:10; equilibration time = 24 h.

showing that ammonia enhanced mercury leaching in the pH range from 6 to 12. Compared with Fig. 1, about 2–3 times more mercury was leached from the washed ash than from the raw ash for the same ammonia and pH conditions. The data suggest that the dissolution of ash during the washing process increased the amount of mercury available for complexation with ammonia. Impacts of the number of ash washing cycles on mercury leaching were also investigated using Ash #85 (fly ash with activated carbon). Raw ash and DI water-washed ash samples with washing cycles of 1, 5, and 10 were used. All reactors were added with 1000 mg l1 ammonia. Results in Fig. 4 indicate that the release of mercury increased with the increase in the number of washing cycles, up to five cycles. Additional washing cycles did not result in increased mercury leaching. The concentrations of mercury and total dissolved solids (TDS) in decants for different washing circles were determined, shown in Table 2. Mercury concentrations in all decants were extremely low, less than 10 ng l1. These low concentrations reflected the strong retention of mercury by fly ash under ammonia-free conditions, and indicated that the washing process using DI water did not remove much of the total available mercury.

The total loss of mercury during the washing process was less than 0.3 ng g1. However, the TDS results indicate that some ash was dissolved during the washing process, especially in the first two washing cycles. The TDS in decants decreased significantly between the first and second washing cycle. The dissolution of the ash may have exposed mercury that originally resided in the subsurface layers of the ash, resulting in more mercury on the surface available for leaching. To determine the impact of washing on the mercury adsorption in the presence of ammonia, a stock mercury solution was added to yield 1 mg l1 of total mercury concentration in the leachant. This concentration was significantly greater than that in the background (available mercury in fly ash). In this case, the spiked mercury can be treated as the total available mercury in the reactor. By monitoring the adsorption behavior of the spiked mercury, one can estimate the mercury adsorption strength on fly ash. Fig. 5 shows the soluble mercury concentration as a function of pH for raw and washed Ash #85, when 1000 mg l1 of ammonia solution which also contained 1 mg l1 mercury was used as a leachant. The leaching results without mercury addition are also displayed. Results indicate that the soluble mercury concentration for the experiment without external mercury addition was negligible compared to that with 1 mg l1 mercury addition. Therefore, the total leachable mercury concentration in the system with 1 mg l1 Hg addition can be treated as 1 mg l1. Results also indicate that less than 4% of spiked

50

Raw Ash 1 Washing 5 Washing 10 Washing

400 300

Hg (μ g/l)

500

Hg (ng/l)

Raw ash without Hg spiking Raw ash with 1 mg/l Hg Washed ash with 1 mg/l Hg

45 40

Ash #85

200

35 30

Ash #85

25 20 15 10 5

100

0 0

0

2

4

6

8

10

12

14

2

4

6

8

10

12

14

pH

pH Fig. 4. Impact of washing cycles on mercury release from Ash #85. Experimental conditions: ammonia concentration = 1000 mg l1; temperature = 20–25 C; S/L = 1:10; equilibration time = 24 h.

Fig. 5. Soluble mercury concentration as a function of pH in the presence of 1000 mg l1 ammonia for raw and washed Ash #85, with and without mercury spiking. Experimental conditions: temperature = 20–25 C; S/L = 1:10; equilibration time = 24 h.

J. Wang et al. / Chemosphere 69 (2007) 1586–1592

mercury remained in the solution even in the pH range favorable for leaching. As shown in Fig. 5, the maximum soluble mercury in washed ash leachates was approximately 25% greater than that in raw ash leachates when pH is in the range between 8.5 and 10.5. Under other pH conditions, soluble concentration curves for both ashes overlap. Since both systems had the same available mercury concentration (i.e. 1 mg l1), the slight increase in the amount of mercury in solution for washed ash is indicative of reduced adsorption strength. Leaching results for raw and washed ash with no mercury spiking, shown in Fig. 4, demonstrate that washing with five cycles increased the leachate mercury concentration by 150%. However, leaching results under the same total available mercury condition, shown in Fig. 5, indicate that washing increased the soluble mercury concentration by only 25% at the maximum for washed ash, as a result of the reduction of the adsorption strength. Therefore, the 150% increase in mercury leaching from washed ash shown in Fig. 4 was mainly the result of the increase in total available mercury after washing. 3.4. Impact of ammonia on mercury adsorption onto raw ash

Adsorption Ratio (%)

Adsorption–desorption could be a major mechanism controlling the mercury leaching from fly ash. In order to determine the ammonia impact on mercury adsorption onto raw ash, the solution containing a high concentration of external mercury (1 mg l1) was used to conduct adsorption experiment. By monitoring the adsorption ratio of the added mercury (i.e. the ratio of the adsorbed mercury to the total available mercury in the system), the adsorption behavior of mercury by raw ash was observed. The adsorption ratio of spiked mercury as a function of pH for all three ashes in the presence of 1000 mg l1 ammonia was determined, Fig. 6. Most spiked mercury was adsorbed by the ashes in the entire experimental pH range. When pH was in the range between 9 and 10, the adsorption ratio decreased but was always greater than 90% for these three ashes. The maximum fraction of solu100 98 96 94 92 90 88 86 84 82 80

Ash #1008 Ash #85 Ash #86

0

2

4

6

8

10

12

14

pH Fig. 6. Mercury adsorption ratio as a function of pH in the presence of 1000 mg l1 ammonia.

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ble mercury for Ash #1008, Ash #85, and Ash #86 were 8.8%, 2.9%, and 3.3%, respectively. The greater soluble mercury ratio for Ash #1008 compared to the other two samples could be a result of the lower LOI in this ash because the unburned carbon has stronger adsorption strengths than other ash components (Hwang et al., 2002). Therefore, the leaching performance of the available mercury as a result of ammonia is highly dependent on the ash characteristics especially the unburned carbon content, and the field pH, which is normally in the range of 5.8– 12.09 (5th–95th percentile) (EPA, 2006). 4. Conclusions This research indicated that mercury is strongly adsorbed onto fly ash. High concentrations of ammonia enhanced the mercury leaching in the alkaline pH range, due to the formation of less adsorbable mercury–ammonia complexes. Weathered ash released more mercury than the fresh ash under the same pH and ammonia conditions, probably due to the dissolution of some soluble ash constituents which in turn increased the mercury availability. Results indicated that more than 90% of available mercury is retained even under the 1000 mg l1 ammonia and all pH conditions for all three ash samples tested. Actual mercury leaching under field conditions is a factor of many variables, including ash composition especially unburned carbon content, age, water contact, field pH, and ammonia concentration. Acknowledgements This work was partially supported by Electric Power Research Institute (EPRI), by the University of Missouri System Research Board, and by the Environmental Research Center (ERC) for Emerging Contaminants at the University of Missouri-Rolla (UMR). Conclusions and statements made in this paper are those of the authors, and in no way reflect the endorsement of the funding agencies. References ACAA, 2006. 2005 Coal Combustion Product (CCP) Production and Use Survey. American Coal Ash Association. . Chu, P., Goodman, N., Behrens, G., Roberson, R., 2001. Total and specified mercury emissions from US coal-fired power plants. In: Proceedings of the Fourth Electric Utilities Environmental Conference, Tucson, Arizona. EPA, 2001. 2000 PBT Program Accomplishments. EPA-742-R-01-003. EPA, 2005. and . EPA, 2006. Characterization of Mercury-Enriched Coal Combustion Residues from Electric Utilities Using Enhanced Sorbents for Mercury Control. EPA-600/R-06/008. EPRI, 1999. The Stability of Mercury Captured on Sorbent Surfaces. EPRI, Palo Alto, CA: Report TE-113926. EPRI, 2001a. Occurrence and Fate of Mercury in Coal Ash and Flue Gas Desulfurization Sludge. EPRI, Palo Alto, CA: Report 1005212.

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