Adsorption of Elemental Mercury on the Residual Carbon in Coal Fly

elemental mercury (Hg0) from flue gas streams. The purpose of this project was to determine whether the unburned carbon that remains in coal fly ash c...
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Ind. Eng. Chem. Res. 2000, 39, 1723-1730

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Adsorption of Elemental Mercury on the Residual Carbon in Coal Fly Ash Shannon D. Serre* and Geoffrey D. Silcox Department of Chemical and Fuels Engineering, University of Utah, 1495 East 100 Street, Salt Lake City, Utah 84112-1114

The injection of large quantities of pulverized activated carbon is one method used to remove elemental mercury (Hg0) from flue gas streams. The purpose of this project was to determine whether the unburned carbon that remains in coal fly ash could be used as an inexpensive and effective replacement for activated carbon. Bench-scale tests were conducted at conditions representative of those found in the flue gas trains of coal-fired power plants and municipal waste incinerators. The temperatures and concentrations ranged from 121 to 177 °C and from 0.019 to 11.7 mg of Hg/m3. Two types of data were obtained: equilibrium data suitable for obtaining adsorption isotherms and breakthrough data suitable for obtaining adsorption kinetics. Adsorbed-phase concentrations were as high as 600 ppm. Forward adsorption rate constants were ≈0.06-2.3 m3/g/s for particle sizes and carbon contents ranging from 59 to 206 µm and from 2% to 36%. Mathematical models were developed to simulate the capture of Hg0 in flue gas ducts and in baghouses. The results of the simulations indicate that a negligible amount of Hg0 can be adsorbed by a dilute suspension of fly ash. The best option for controlling Hg0 emissions using fly ash appears to be injection in pulses prior to a baghouse. Introduction Mercury is released into the environment from natural and anthropogenic sources. One of the largest anthropogenic sources is coal-fired utilities. Currently, there are no federal regulations governing the release of Hg0 from coal-fired utilities. Under the Clean Air Act Amendments of 1990, the U.S. Environmental Protection Agency (EPA) is required to determine whether to regulate the emission of Hg0 from coal-fired utilities. Mercury is present in trace amounts in coal with concentrations varying with coal rank. Measurements indicate Hg0 levels in U.S. coals that range from 0.02 to 0.25 ppmw with lignites having a higher concentration.1 Typical flue gas Hg0 concentrations range from 0.3 to 35 µg/Nm3;2 however, values as high as 70 µg/ Nm3 have been reported.3 Physical coal cleaning can remove a significant amount of Hg0 before combustion, but flue gas treatment may also be required. There are several methods that are used to control Hg0 emissions, one of which is the injection of activated carbon into a flue gas stream. The carbon is then removed in the particulate control device, either an electrostatic precipitator (ESP) or a baghouse. The main drawback to using activated carbon as a sorbent is its high cost. The goal of this research was to determine if the carbon that remains in pulverized coal fly ash could be used as an inexpensive adsorbent for elemental Hg0. The idea of using recycling fly ash for metals capture was first proposed by Owens et al.4 The fly ash would be injected into the flue gas train prior to the particulate control device, similar to the way in which activated carbon is used, thus eliminating large capital and sorbent costs. One of the fly ashes used in this study, * To whom correspondence should be addressed: U.S. EPA, 86 T.W. Alexander Dr., ORD/NRMRL/APPCD/APTB, MD-65, Research Triangle Park, NC, 27711.

Cherokee ash, has been shown to reduce Hg0 emissions by 98%-99% in the power plant in which it was generated.5 Hassett and Eylands6 has also shown that fly ash can be used to capture mercury. Experiments were conducted using several coal fly ashes to determine the feasibility of using fly ash as a sorbent for Hg0. Specifically, the roles of temperature, particle size, time, gaseous Hg concentration, and fly ash carbon content were studied. The experimental results were used to obtain kinetic rate constants for use in simulations of duct and baghouse capture. Experimental Section Adsorption experiments were carried out in an apparatus consisting of several separate components: (1) Hg0 vapor generator, (2) temperature control system, (3) reactor, and (4) Hg0 analyzer. A schematic of the experimental system is included as Figure 1. All of the experiments were performed using nitrogen as the carrier gas, except as noted. The Hg0 laden gas stream was generated by passing nitrogen through an impinger containing Hg0. This stream was then diluted with a second nitrogen stream to obtain the desired Hg0 concentration. The impinger was kept at a constant temperature using a water bath, and the temperature of the bath was varied to change the concentration of Hg0 in the gas stream. The Hg0-laden stream was transferred throughout the system using Teflon tubing which was heated to prevent condensation of the Hg0 and to preheat the gas to the reactor temperature. The reactor was heated and kept at the desired temperature using a gas chromatograph oven. Two reactor configurations were used. The first was the fixed-bed reactor shown in Figure 1. The gas was forced down through a bed of fly ash that was supported on a glass frit. The second was a fluidized-bed reactor. The concentration of Hg0 in the inlet and outlet gas flows was measured with a gold-film Hg0 analyzer,

10.1021/ie990680i CCC: $19.00 © 2000 American Chemical Society Published on Web 05/04/2000

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Figure 1. Schematic of experimental apparatus used for the adsorption of Hg0 onto fly ash, shown with the fixed-bed reactor. Table 1. Physical Characteristics of Fly Ashes source of ash (power plant)

carbon content (wt %)

initial Hg0 content (ppm)

surface area (m2/g of ash)

mean diameter (µm)

Nixon Cherokee Clark Huntington

2.0 8.7 32.7 35.9

0.33 0.25 1.1 0.51

6.8 34.1 65.1 63.8

59 77 206 93

Jerome Model 411. Solid samples were digested and analyzed for elemental Hg0 using cold vapor atomic absorption (CVAA). Four different fly ashes, from full-scale coal-fired power plants, were used in this study. The physical characteristics of the fly ashes are included in Table 1. The carbon contents ranged from 2% to 35.9%. Their BET surface areas ranged from 6.8 to 65.1 m2/g of ash, significantly lower than surface areas of commercial activated carbons. Results The key parameters affecting Hg0 adsorption include the carbon content, particle diameter, temperature, surface area of the ash, and the vapor-phase Hg0 concentration. The effects of these parameters are discussed in the sections that follow. Role of Fly Ash Carbon Content. The amount of carbon in the fly ash has a strong effect on the removal of Hg0 as shown in Figure 2. These results were obtained at a Hg0 concentration of 4 mg of Hg/m3 and a temperature of 121 °C. Although this concentration is an order of magnitude higher than that typically found in the flue gas of municipal waste incinerators, it was used to obtain relative adsorptive capacities. The low or 2% carbon ash (Nixon) adsorbed the least amount of Hg0, 30.7 ppm, followed by the 8.7% carbon ash (Cherokee), adsorbing an average of 108 ppm. The two high carbon ashes adsorbed the most Hg0: the 32.7% carbon ash (Clark) contained 340 ppm and the 35.9% carbon ash (Huntington) contained 807 ppm. One can conclude from Figure 2 that Hg0 sorption increases with increas-

Figure 2. Adsorbed Hg0 as a function of fly ash carbon content at a temperature of 121 °C and a gaseous Hg0 concentration of 4 mg of Hg0/m3.

ing carbon content; however, the level of adsorbed Hg0 was not directly proportional to the carbon content. For example, an increase by a factor of 18 was not observed when comparing the 2%- and the 35.9%-carbon ash. The sorption capacity may depend on more than just the carbon content of the ash because all of the carbon sites need not be accessible to the Hg0 molecules. The surface morphologies of the Clark (32.7% C, 65 m2/g) and Huntington (35.9% C, 64 m2/g) ashes were examined with a scanning electron microscope (SEM) to be better understood as to why the Huntington ash adsorbs twice as much Hg0 as the Clark ash.7 The carbon contents and N2-BET surface areas of both are similar. The SEM images revealed that the surface of the Clark ash is relatively smooth with limited surface porosity and that the surface of the Huntington ash is covered with relatively large pores. Its interior is more porous than that of the Clark ash. The higher porosity of the Huntington ash provides greater access to interior adsorption sites and helps explain its higher capacity for adsorption.

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Figure 3. Mercury adsorption as a function of fly ash surface area at a temperature of 121 °C and a Hg0 concentration of 4 mg of Hg0/m3.

Figure 4. Effect of temperature on adsorption for the 8.7% carbon ash (Cherokee) at a Hg0 concentration of 4.64 mg of Hg/m3 in the fixed-bed reactor.

Effect of Surface Area. A large surface area enhances the effectiveness of activated carbons for the removal of Hg0 from a gas stream. The surface area of the ashes used in this work ranged from 6.8 to 65 m2/g of ash. The amount of Hg0 adsorbed as a function of fly ash surface area is shown in Figure 3. These tests were run at a gaseous Hg0 concentration of 4 mg of Hg0/m3 and a temperature of 121 °C. The amount of Hg0 adsorbed generally increases with increasing surface area; however, the surface area of the Clark ash was higher than that of the Huntington ash, but the Clark ash adsorbed less than half as much Hg0. This could be the result of the larger particle size of the Clark ash (206 versus 93 µm), or it could be due to the higher porosity of the Huntington ash. Effect of Temperature. The effect of temperature on equilibrium adsorption was studied for a typical temperature range found in power plant flue gas trains, 120-180 °C. Equilibrium sorption data are presented in Figure 4 for the 8.7% carbon ash at 121, 149, and 177 °C. Figure 4 shows an obvious decrease in sorption capacity as the temperature is increased. At 121 °C the ash contained 73 ppm Hg0, compared to 25 ppm at 177 °C. The sorption capacity dropped 48% when the temperature was increased from 121 to 149 °C. A decrease of 34% was observed when the temperature was increased from 149 to 177 °C. The decrease in capacity

Figure 5. Effect of gaseous Hg0 concentration on the amount of Hg0 that can be adsorbed by the 8.7% carbon ash (Cherokee) at a temperature of 121 °C.

Figure 6. Ash-loading results for the Nixon and Clark fly ashes, showing the role of carbon content, temperature, and adsorbate concentration on adsorption.

with an increase in temperature is typical of a physisorption mechanism. Effect of Adsorbate Concentration. The effect of gaseous Hg0 concentration on adsorption for the 8.7% carbon ash (Cherokee) in the fixed-bed reactor is shown in Figure 5. The amount of Hg0 adsorbed increased nonlinearly with the vapor-phase concentration. At the lower Hg0 concentrations, 0.019-0.079 mg/m3, the amount that is retained in the ash is still significant and appeared to descend to a plateau. This is important because the concentrations found in coal flue gas are near the lower gas concentrations shown in Figure 5. Similar effects of vapor-phase concentration on the adsorbed-phase concentration were observed for the 2%- (Nixon) and 32.7%- (Clark) carbon ashes for concentrations of 0.048 and 0.48 mg of Hg0/m3. The additional effects of carbon content, surface area, and temperature are also shown in Figure 6 for these two fly ashes at mercury concentrations of 0.048 and 0.48 mg of Hg0/m3. As discussed above, the Hg0 capacity increases with an increase in carbon content, an increase in vapor-phase Hg0 concentration, and a decrease in temperature. Comparison of Fly Ash with Commercial Activated Carbon. The sorption characteristics of the 2%- (Nixon), 8.7%- (Cherokee), and 32.7%- (Clark)

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Figure 7. Cumulative adsorption comparison for fly ashes and Calgon HGR at a temperature of 177 °C and a gaseous Hg concentration of 0.48 mg of Hg0/m3.

carbon ashes are compared with Calgon HGR in Figure 7. The tests were performed in the fixed-bed reactor at a temperature of 177 °C and a Hg0 concentration of 0.48 mg of Hg0/m3. The Calgon HGR, with a surface area of around 1000 m2/g, adsorbed 2-times more Hg than the Clark ash after 20 min of exposure. Two different size fractions of Calgon HGR were used to determine whether particle size has any effect on adsorption. The asreceived Calgon HGR has a mean size of 1190 µm. This HGR was crushed and sieved. Two size fractions were kept, the material passing a No. 325 screen (44 µm) and passing a No. 40 screen (420 µm), but retained on a No. 60 screen (250 µm). On the basis of these two size fractions, there was not any effect of particle size. Effect of Sulfur Dioxide. The effect of sulfur dioxide on the adsorption characteristics of the 8.7%-carbon ash (Cherokee) at a temperature of 121 °C is shown in Figure 8. A sulfur dioxide calibration gas (SO2 in N2) was used to give a concentration of 400 ppm. The ash was exposed to the gas stream for more than 2 h. The presence of sulfur dioxide in the simulated gas stream resulted in a 40% reduction in the amount of Hg0 that was adsorbed. Schager et al.8 also noticed a decrease in the presence of sulfur dioxide. An increase was reported by Benson et al.9 A decrease could be a result of the sulfur dioxide competing for adsorption sites. Under reducing conditions the sulfur could react with the Hg0 to form mercuric sulfide, thus boosting the removal efficiency. Modeling and Analysis. Two mathematical models were developed to simulate the removal of elemental Hg0 in flue gas ducts and baghouses by coal fly ash. The first applies to the adsorption of Hg0 by a dilute suspension of fly ash particles. The second predicts removal in a baghouse in a growing cake or bed of fly ash. The kinetics of Hg0 removal were described using Langmuir kinetics10,11 for the adsorption of a single solute from a fluid phase onto a solid adsorbent. The rate parameters were obtained by analyzing the adsorption isotherms and breakthrough curves. The data are analyzed below, followed by a discussion of the development of the models and their application to industrial Hg0 control processes. Adsorption Isotherms. Equilibrium adsorption data were obtained at temperatures of 121 and 177 °C and

Figure 8. Effect of sulfur dioxide on Hg0 adsorption for the 8.7% carbon ash (Cherokee) at a temperature of 121 °C and a Hg0 concentration of 0.055 mg of Hg0/m3.

at gas-phase concentrations ranging from 0.06 to 6 mg of Hg0/m3. Known amounts of ash were placed in the fixed-bed reactor and allowed to achieve equilibrium with the gas. Equilibrium was assumed when the concentration of Hg0 in the feed stream was equal to that in the exit stream. An example of one of the isotherms obtained is shown in Figure 9 for Clark ash (32.7% carbon) at 177 °C. The solid line is a fit of the Langmuir isotherm to the data. The isotherms resemble Type I isotherms that can be described by the theory developed by Langmuir. The net rate of adsorption is equal to the difference between the rate of adsorption and the rate of desorption:

rads )

∂ω ) k1C(ωmax - ω) - k2ω ∂t

(1)

where the concentration of adsorbate on the solid is ω, ωmax is the maximum solids loading, and C is the concentration of Hg0 in the vapor phase. The adsorption rate constant is k1 and the desorption rate constant is k2. At equilibrium the rate of adsorption is equal to the rate of desorption and from eq 1

ω)

ωmaxKC 1 + KC

(2)

where K is the equilibrium constant which is the forward rate constant divided by the reverse rate constant. The equilibrium data were used to estimate the Langmuir parameters (ωmax, K) for the Nixon, Cherokee, and Clark ashes at temperatures of 121 and 177 °C. Nonlinear regression was used to fit the equilibrium data to eq 2 and the Langmuir parameters are reported in Table 2. The maximum adsorbed-phase concentration, ωmax, decreases with increasing temperature. The high-carbon Clark and the low-carbon Nixon ashes had the highest and lowest values of ωmax. Isotherm data for a commercially activated carbon, Darco G60, were obtained from Karatza et al.12 and are included in Table 2. The Darco G60 has a bulk density of 450 kg/m3 and a surface area of 230 m2/g. The values of ωmax for the Darco carbon are lower than those for the Clark ash. This could be because the pores are not

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Figure 9. Equilibrium adsorption data for Clark ash (32.7% carbon) at 177 °C. The solid line is a fit of the Langmuir isotherm to the data. Table 2. Langmuir Parameters for Selected Coal Fly Ashes and Darco G60 Activated Carbon

Table 3. Adsorption Rate Constant, k1

ash

temp. (°C)

ωmax (g/g)

K (m3/g)

Nixon Nixon Cherokee Cherokee Clark Clark Darco G60 Darco G60 Darco G60

121 177 121 177 121 177 90 120 150

6.95 ( (0.6) × 10-5 1.34 ( (0.11) × 10-5 1.31 ( (0.12) × 10-4 6.05 ( (0.56) × 10-5 7.51 ( (0.64) × 10-4 2.37 ( (0.21) × 10-4 6.92 × 10-4 1.91 × 10-4 1.07 × 10-4

666 ( 51 155 ( 12 1594 ( 124 472 ( 33 655 ( 50 395 ( 30 1200 668 420

as accessible. The surface area is also low for an activated carbon; for example, the Clark ash has a surface area of 200 m2/g of carbon. Analysis of Breakthrough Data. Breakthrough curves were analyzed to give the rate constant, k1, as follows. The governing partial differential equation for the adsorption of Hg in a packed bed is

Dax

∂2 C ∂C ∂C 1 -  ∂ω )0 - Vi ∂Z ∂t  ∂t ∂Z2

(3)

The axial dispersion term (Dax) can be assumed negligible if the ratio of the column diameter to the particle diameter is larger than 35 and if the ratio of the column length to the particle diameter is larger than 75.13 When the axial dispersion term is neglected, eq 3 reduces to

∂C 1 -  ∂t ∂C ) -Vi ∂t ∂Z  ∂ω

(4)

Equation 4 is coupled with the net rate of adsorption given by eq 1 and both are integrated with the initial and boundary conditions; at time equal to zero, the amount of Hg0 in the vapor phase and loaded on the solids is zero. When the Hg0-laden gas is suddenly passed through the bed, at times greater than zero, the gaseous Hg0 concentration at the inlet of the bed is equal to the baseline Hg0 concentration, C0.

ash

temp. (°C)

k1 (m3/g/s)

Nixon Nixon Cherokee Cherokee Clark Clark Darco G60 Darco G60 Darco G60

121 177 121 177 121 177 90 120 150

0.23 ( 0.04 2.3 ( 0.1 0.20 ( 0.04 0.40 ( 0.08 0.056 ( 0.012 0.16 ( 0.03 0.415 0.505 0.630

The Thomas solution was fit to the breakthrough data using the solver routine in Microsoft Excel 97. The only fitting parameter was the adsorption rate constant k1. Values for ωmax and K were obtained from the isotherm data. The breakthrough data were analyzed for a specific ash at a given temperature over a range of Hg0 concentrations. The values of k1 that were obtained at the different concentrations were then averaged. Values for k2 are easily determined because the equilibrium constant K is the adsorption rate constant, k1, divided by the desorption rate constant. The values of k1 that best fit the data are shown in Table 3. Values of k1 for the Darco G60 activated carbon 12 are also included in Table 3. An example of the fit obtained with the Thomas solution is shown in Figure 10 for the Cherokee (8.7% C) ash at 177 °C and 0.044 mg of Hg/m3. The Thomas solution fits the breakthrough data fairly well, although it approaches the asymptotic limit too rapidly. Simulation of Industrial Applications. To examine the practical application of fly ash injection for Hg0 control, mathematical models were developed for capture in the dispersed phase of a duct or ESP and for capture in growing beds of a baghouse. The models were used to determine the amount of sorbent required to reduce Hg0 emissions to a desired level. The kinetic and isotherm parameters obtained above were used. Dispersed-Phase Simulation. The steady-state equation for the dispersed-phase Hg capture model was obtained from a mass balance on vapor-phase Hg over

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Figure 10. Comparison of the Thomas solution to experimental breakthrough data for the Cherokee (8.7% C) ash at a temperature of 177 °C and a Hg0 concentration of 0.044 mg of Hg0/m3.

a section of the duct of width ∆Z.

QCHg|Z - QCHg|Z+∆Z - radsβAc∆Z ) 0

(5)

where Q is the volumetric flow rate of the gas, CHg is the concentration of Hg in the vapor phase, rads is defined in eq 1, Ac is the cross-sectional area of the duct, and Z is the length in the axial direction. The parameter β is defined as the fly ash or sorbent loading in grams of sorbent per unit volume of gas. Equation 5 assumes that the fly ash particles are moving with the same velocity as the gas. Dividing by ∆Z and letting ∆Z f 0, the following equation is obtained:

-Q

dCHg ) radsβAc dZ

(6)

Assuming that the adsorption process is not masstransfer limited, we can couple eq 1 with eq 6 to obtain removal throughout a duct. Equations 6 and 1 were integrated using the fourth-order Runge-Kutta method described in Press et al.14 Simulations were run for each of the three fly ashes and the Darco G60 activated carbon at a temperature of 121 °C and a gaseous Hg0 concentration of 1 mg of Hg0/m3. The results are shown in Figure 11. At sorbent injection rates of 10, 50, and 100 g/m3 of gas, little Hg is removed. An injection ratio of 100 g of sorbent/m3 of gas is equivalent to ≈100 ton (90.7 × 106 g) of sorbent/h based on a gas flow rate of 250 m3/s. Simulation of Capture in a Growing Bed. If a baghouse is used for particulate control, dispersed-phase capture in a duct is coupled with capture in a baghouse filter cake. The concentration in the dispersed-phase calculation is used as the initial concentration for the growing bed model, and the solids loading parameter (ω) is initialized to the value of ω at the exit of the dispersed-phase section. The Thomas solution could be used if the thickness of the fly ash cake were constant and if the bed were initially free of Hg0; however, the thickness of the baghouse cake increases over time until the cake is removed. The governing equations for the growing bed model are eqs 1 and 4 and they can be solved using the method of lines. The bed of thickness Z is divided by N nodes that are equally spaced with the first node (1) at the start of the bed and continuing to the end node (N), with node 1 being the node most recently added. Before eqs

Figure 11. Dispersed-phase simulation for the three fly ashes and Darco G60 activated carbon at a temperature of 121 °C and a gaseous Hg0 concentration of 1 mg of Hg0/m3. Injection rates of 10, 50, and 100 g/m3 of gas were used. This figure shows the percent of Hg0 removed based on a given injection rate. Table 4. Operating Parameters for a Pulverized Coal Boiler Utilizing an ESP net power distance from air heater to ESP or baghouse gas flow air heater duct dimensions ESP residence time temperature

780 MW 36 m 250 m3/s 6 m × 5.2 m 7-8 s 121-177 °C

1 and 4 can be solved, an approximation must be provided for ∂C/∂Z. For a problem involving the flow of flue gas through the bed, the central difference approximation is not appropriate. An upwind finitedifference approximation must be used to place emphasis on the upstream conditions. Substituting, we obtain a set of two ODEs

(

)

dCi Ci-1 - Ci 1 -  dω ) -Vi dt ∆Z  dt

(7)

dωi ) k1C(ωmax - ωi) - k2ωi dt

(8)

and

that can be integrated with LSODE.15 The entering vapor-phase Hg concentration for the fixed bed will be the exiting concentration from the dispersed-phase model. The concentration of Hg on the ash (ω) is similarly set to the value for ω at the end of the dispersed-phase section. Baghouse capture of Hg was simulated using the growing bed model at conditions that are characteristic of a full-scale power plant. These conditions are specified in Table 4 and were used to estimate residence time for the gas and fly ash in a flue gas duct as well as to give an estimate of the amount of ash required to achieve a given Hg removal rate. The conditions in Table 4 are for a power plant that operates an ESP and were obtained from a case study published by the Electric Power Research Institute.16 For simulation of a baghouse, an air-to-cloth ratio of 1 m/min and a pressure drop of 5 in. of H2O were used.17 The calculated Hg concentration leaving a bed as a function of time and the amount of Hg on the sorbent is shown in Figure 12 for the Clark ash (32.7% C). The time that the bed has been in service is shown on the x

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Figure 12. Simulation of baghouse capture showing the concentration of Hg exiting the system and the concentration of Hg0 on the ash at node N for the Clark (32.7% C) ash at 121 °C and a gaseous Hg0 concentration of 1 mg of Hg0/m3. The prediction is for 60 min at an injection rate of 20 g of sorbent/m3 of gas.

Figure 13. Simulation showing the concentration of Hg exiting the system as a function of time for the three fly ashes and the Darco G60 at 150 °C and a gaseous Hg0 concentration of 0.1 mg of Hg0/m3. The prediction is for baghouse capture for 60 min at an injection rate of 1000 g of sorbent/m3 of gas for the first minute.

axis. The conditions are 121 °C, 1 mg of Hg0/m3, and an injection rate of 20 g of ash/m3 of flue gas. Mercury is removed as the ash moves through the duct, but by the time the ash reaches the baghouse, only 0.4% of the Hg0 has been removed. The gaseous Hg0 concentration at the start of the baghouse cake is 0.996 mg of Hg/m3. The concentration in the vapor phase is high at the first node and decreases as the gas passes through the bed. The exiting vapor-phase concentration, after 60 min, reaches a lower limit of about 0.43 mg of Hg0/m3sa reduction of 57%. The adsorbed-phase concentration on the ash at the Nth node is also shown in Figure 12. It is initially zero and increases to about 49 ppm at node N after 60 min. This is far below the maximum solids loading value of 751 ppm. When ash is injected at a constant rate, the newly deposited ash does not have time to adsorb a significant amount of Hg0. More Hg0 could be removed, and the ash could be better utilized, if the ash were injected at a high rate for the first minute or so, followed by no injection for the remainder of the cycle. A simulation of pulsed ash injection, with conditions similar to those used in Figure 12, is shown in Figure 13. In Figure 13 the ash was injected at a rate of 1200 g/m3 of gas for the first minute, followed by no injection for the remaining 59 min. The adsorbed-phase concentration in Figure 13 is shown for node 1. Node 1 is exposed to the highest concentration of Hg0 in the vapor phase and after 60 min of exposure its concentration is 85 ppm. The vaporphase concentration at the cloth is consistently lower in Figure 13 compared to that shown in Figure 12. A pulsed injection system provides superior Hg0 removal. An automatic control system could also be used to inject short spurts of carbon to prevent the concentration of Hg0 from increasing above a desired level. A comparison of the three fly ashes with the Darco G60 for pulsed injection is included as Figure 13. The simulation is at a temperature of 150 °C, an inlet concentration of 0.1 mg of Hg0/m3, and a sorbent injection rate of 1000 g/m3 of gas for the first minute. The Darco G60 initially reduces the exiting Hg0 concentration more than the ashes, but as time increases, the Cherokee and Clark ashes give lower exit concentrations. This could be due to the high value of the

equilibrium constant, K, for the Cherokee and Clark ash. Summary and Conclusions The purpose of this research was to evaluate the feasibility of using the carbon that remains in pulverized-coal fly ash as a sorbent for Hg0. Experiments were conducted using four coal fly ashes, with carbon contents ranging from 2% to 35.9% carbon. The effects of temperature, particle size, time, gaseous Hg concentration, and fly ash carbon content were examined. The key conclusions follow. (1) Fly ash containing between 2% and 35% carbon can adsorb a significant amount of Hg0. Equilibrium capture approached 600 ppm for the Clark fly ash at a temperature of 121 °C and a gaseous Hg0 concentration of 6 mg of Hg0/m3. (2) The adsorbed-phase Hg0 concentration at equilibrium, per unit weight of ash, is proportional to the carbon content of the ash. As the amount of carbon increases, the adsorbed-phase Hg0 increases. The mineral matter in the ash did not adsorb detectable levels of Hg0. (3) The amount of Hg0 adsorbed generally increases with the surface area of the carbon in the ash. One exception is the Huntington ash, which has a lower surface area than the Clark ash, but adsorbed twice as much Hg0. This could be due to the larger particle size of the Clark ash and to morphological differences. Studies of ash morphology by SEM reveal that the Huntington ash was more porous. (4) The amount of Hg0 that is adsorbed decreases as the temperature is increased from 121 to 177 °C. This decrease is representative of a physisorption process. (5) The presence of sulfur dioxide resulted in a 40% reduction in the amount of Hg0 adsorbed. The sulfur dioxide may be competing for adsorption sites. (6) The adsorption characteristics of the Clark, Cherokee, and Nixon ashes were compared with those of a commercially activated carbon designed to adsorb Hg, Calgon HGR. The HGR adsorbed considerably more Hg0 than any of the ashes. At the end of a 100-min run the Calgon carbon adsorbed 40 ppm Hg and was continuing to adsorb Hg0 while the ashes had become saturated by 50 min and contained less than 15 ppm Hg0.

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(7) Equilibrium adsorption data were described by the Langmuir isotherm. The equilibrium constant (K) and maximum solids loading value (ωmax) decreased with increasing temperature. The equilibrium constant (K) ranged from 395 to 1594 and ωmax ranged from 13 up to 751. (8) A mathematical model was used to simulate the dispersed-phase capture of elemental Hg0 in ducts by coal fly ash injection. The predictions indicate that less than 5% of the vapor-phase Hg0 is adsorbed. (9) A second model simulated the capture of Hg0 in a baghouse filter cake. Contiuous injection of fly ash resulted in poor utilization (7%) of the available carbon. Pulsed injection gave higher utilization and permitted adsorption to occur for the majority of the cake’s life. For example, pulsed injection of 1000 g of sorbent/m3 of gas over a period of 1 min kept emissions below 50% of the inlet concentration of 1 mg of Hg0/m3 for the entire 60-min cycle and gave a carbon utilization of 12%. Acknowledgment Funding for this project was provided primarily by the National Science Foundation through the Advanced Combustion Engineering Research Center. Additional funding was provided by the U.S. Department of Energy through a grant by the Small Business Innovation Research Program as awarded to Reaction Engineering International. Literature Cited (1) Chu, P.; Porcella, D. B. Water, Air, Soil Pollut. 1995, 80, 135-144. (2) Meij, R. Water, Air, Soil Pollut. 1991, 56, 21-33. (3) Livengood, C. D.; Huang, H. S.; Mendelsohn, M. H.; Wu, J. M. Mercury Capture in Bench Scale Absorbers. Presented at the 10th Annual Coal Preparation, Utilization, and Environmental Control Contractors Conference, Pittsburgh, PA, July 1994.

(4) Owens, W. D.; Sarofim, A. F.; Pershing, D. W. The Use of Recycle for Enhanced Volatile Metal Capture. Presented at the Trace Elements Transformations in Coal-Fired Power Systems Workshop, Scottsdale, AZ, April 1993. (5) Grover, C. et al. Mercury Measurements Across Particulate Collectors of PSCO Coal-Fired Utility Boilers. Presented at the 1999 Mega-Symposium, Atlanta, GA, Aug 1999. (6) Hassett, D. J.; Eylands, K. E. Fuel 1999, 78, 243-248. (7) Serre, S. D. Doctoral Dissertation, University of Utah, Salt Lake City, UT, 1999. (8) Schager, P.; Hall, B.; Lindquist, O. Retention of Gaseous Mercury on Flyashes. Presented at the Second International Conference on Mercury as a Global Pollutant, Monterey, CA, June 1992. (9) Benson, S. A. et al. Center for Air Toxic MetalssFinal Report; Report 95-EERC-04-01; Energy & Environmental Research Center: Grand Forks, ND, June 1995. (10) Thomas, H. C. J. Am. Chem. Soc. 1944, 66, 1664-1666. (11) Thomas, H. C. Ann N. Y. Acad. Sci. 1948, 49, 161-182. (12) Karatza, D.; Lancia, A.; Musmarra, D.; Pepe, F. Adsorption of Metallic Mercury on Activated Carbon. Presented at the TwentySixth Symposium on Combustion, Pittsburgh, PA, July 1996. (13) Noll, K. E.; Gounaris, V.; Hou, W. Adsorption Technology for Air and Water Pollution Control; Lewis Publishers: Chelsea, MI, 1992. (14) Press, W. H.; Teukolsky, S. A.; Vetterling, W. T.; Flannery, W. P. Numerical Recipes in Fortran, The Art of Scientific Computing; Cambridge University Press: Victoria, Australia, 1992. (15) Hindmarsh, A. C. LSODE and LSODI, Two New Initial Value Ordinary Differential Equation Solvers. ACM-SIGNUM Newsletter 1980, 15. (16) Electric Power Research Institute. Engineering Assessment of an Advanced Pulverized-Coal Power Plant; Report EPRI CS2223; Electric Power Research Institute: Palo Alto, CA, Jan 1982. (17) Shannon, R. H. Handbook of Coal-Based Electric Power Generation; Noyes: Park Ridge, NJ, 1982.

Received for review September 14, 1999 Revised manuscript received March 3, 2000 Accepted March 9, 2000 IE990680I