A Laboratory Investigation on Combined In-Furnace Sorbent Injection

Tiffany L. B. Yelverton , David G. Nash , James E. Brown , Carl F. Singer , Jeffrey V. Ryan , Peter H. Kariher ,. AIMS Environmental Science 2016 3 (1...
0 downloads 0 Views 537KB Size
Download PDF
Environ. Sci. Technol. 2000, 34, 4855-4866

A Laboratory Investigation on Combined In-Furnace Sorbent Injection and Hot Flue-Gas Filtration to Simultaneously Capture SO2, NOx, HCl, and Particulate Emissions BROOKE SHEMWELL, AJAY ATAL, YIANNIS A. LEVENDIS,* AND GIRARD A. SIMONS Northeastern University, Boston, Massachusetts 02115, and Simons Research Associates, Lynnfield, Massachusetts 01940

The effectiveness of a novel integrated technique for removing pollutants generated during the combustion of fuels in furnaces is investigated in this experimental and theoretical study. Sorbents, containing calcium/magnesium and organic components, are wet- or dry-sprayed into the postcombustion zone of a furnace. The organic components of the sorbents pyrolyze and reduce NOx under oxygen-lean conditions. The calcium based residues calcine and react with sulfur- and/or chlorine-bearing gases (such as SO2, H2S, HCl, Cl2, etc.), forming stable salts of calcium (CaSO4, CaCl2, etc.). The partially reacted sorbent particles are then collected in the high-temperature ceramic honeycomb filter/reactor, where they continue to react for a prolonged period of time, until the filter is regenerated (cleaned). Using this technique, both the likelihood and the duration of contact between the solid sorbent particles and the gaseous pollutants increases, since reaction takes place both upstream of the filter and inside the filter itself. Hence, the sorbent utilization increases drastically. High filtration efficiency wall-flow honeycomb ceramic monoliths retain particulates, including partially burned carbon (such as soot, char, tar, etc.), ash, and spent sorbent. Complete combustion of carbon may occur inside the hightemperature filter. Periodic aerodynamic regeneration (backpulsing) of the filters removes the collected minerals (ash and spent sorbent). Using this technique, most of the major combustion-generated pollutants can be removed from the effluent with certain efficiencies. In this laboratory study, powders of calcium formate (CF), calcium propionate (CP), calcium(magnesium) acetate (CMA), calcium carbonate (CC), and calcium oxide (CX) sorbents were sprayed in an electrically heated, laminar flow drop-tube furnace where toxic gases of SO2, NO, and HCl were also introduced. Gas temperatures were in the range of 1023-1423 K (7501150 °C) and gas residence times ≈1 s. A honeycomb ceramic filter, placed at the exit of the furnace, was externally heated to 873-1173 K (600-900 °C). Concentration reductions of the pollutants were measured both with and without the presence of the filter. The “pore tree” mathematical model that describes heterogeneous reaction and transport in porous media was modified to (a) account for simultaneous reaction of calcium with SO2 and HCl, and (b) account for time-dependent reaction in 10.1021/es001072t CCC: $19.00 Published on Web 10/21/2000

 2000 American Chemical Society

an accumulating bed of sorbent in the filter. Experimental results showed that (i) the sorbent can react with both HCl and SO2 and result in high removal efficiencies for both pollutants, (ii) the filter greatly enhances the SO2 removal and captures particulates, and (iii) the sorbents that contain organics reduce NOx at oxygen-lean conditions. For instance, SO2 removal by CF or CMA, which was 40% without the filter, improved to 80% with the filter (kept at 1173 K), at a Ca/S of 1.5. CMA also reduced NOx by 70%. As NOx reduction mostly depends on fast gas-phase reactions, it was only mildly enhanced at the presence of the filter. Sorbent utilizations, however, were drastically increased in the longer residence time allowed by the presence of the filter. The model’s predictions of the experimental data for SO2/HCl concentration reduction and calcium utilizations with and without the filter were fairly successful in most cases, especially at high temperatures.

1. Introduction and Literature Review Phase II of the U.S. Clean Air Act Amendments of 1990 is in effect (as of year 2000), calling for further reduction of air pollutants, such as those causing acid rain. Most coal-fired electric utility plants and waste-to-energy plants use some kind of particulate emissions control, but, unfortunately, relatively few plants have installed devices to control emissions of sulfur oxides, nitrogen oxides, and other pollutants. Since the control of these pollutants typically happens on an “as needed” basis, it involves separate processes. There is rarely, in commercial practice, an integrated approach to the air pollution of commercial and utility furnaces. Removal of SO2. The SO2 target for the Clean Air Act Amendments is a 40% reduction from the 1980 emissions levels by the year 2000. Relatively few of the powerplants impacted by Phase I of the aforementioned Clean Air Act have installed hardware to meet SO2 compliance. Instead, they have resorted to purchasing SO2 emission credits on the open market, or they practice fuel switching to low-sulfur coal or to the cleaner natural gas (1). Many power companies may continue the same practice in Phase II, depending on the price of these commodities. Among those powerplants that have opted for emission control, the most widely used flue gas desulfurization (FGD) strategies involve low-temperature “wet-scrubbing” with aqueous slurries of nonregenerated calcium compounds, such as slaked lime [Ca(OH)2] or limestone [CaCO3]. Wet-scrubbing is effective (in the order of 90% SO2 reduction), but it is costly and it has large space requirements. Injection of limestone (or sodium bicarbonate) in the high-temperature boiler gases (“dry-scrubbing”) has received much attention (mostly at the R&D level) because of the favorable kinetics of the sulfation reactions at elevated temperatures and the low associated capital investment (2). The SO2 removal efficiency with dry-scrubbing ranges from 20 to 70% (3), in comparison to the wet-scrubbers that have efficiencies in the aforementioned 90% range. Sorbent utilizations in dry-scrubbing have been generally reported to be low because of short residence times and gaseous pore-diffusion limitations to the heterogeneous reaction CaO + SO2 + 1/2O2 a CaSO4. Particle size, porosity, surface area, and reaction time are the most important parameters * Corresponding author phone: (617)373-3806; fax: (617)373-2921; e-mail: [email protected]. VOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4855

for the heterogeneous sulfation and sulfidation reactions, as illustrated by many experimental studies (4, 5) and by modeling work (6, 7). A technique that prolongs the residence time of the sorbent in the furnace is presented herein. NOx Control. Phase II of EPA’s Acid Rain Program requires that beginning with the year 2000, U.S. utilities reduce NOx emissions from powerplants by an additional 15% (900,000 tons/year) beyond those of Phase I. NOx emissions may, again, be lowered by fuel-switching to nitrogen-free natural gas. Otherwise, NOx control techniques rely on combustion modifications and postcombustion treatments. Combustion modifications include low-NOx burners and staged combustion, but such modifications alone (which provide NOx reductions in the order of 40-70%) may not be sufficient to meet EPA’s Phase II emission reductions. To further reduce NOx, it appears necessary to use additional postcombustion treatment, such as selective noncatalytic reduction (SNCR) and selective catalytic reduction (SCR), wherein nitrogencontaining compounds such as ammonia (8) and urea (9) are injected downstream of the boiler. Moreover, another promising technology involves injection of secondary fuels in a fuel-rich zone downstream of the primary combustion zone (reburning) (10, 11). NOx reduction efficiencies as high as 80-90% have been reported for SNCR and SCR, whereas reburning techniques reported efficiencies of 40-70%, depending on the fuel and the primary burner configuration (12). Combined SO2 and NOx Removal. A variety of combined SO2 and NOx control methods are in development, but few, if any, have been completely adapted at a commercial scale (13). Such methods include addition of sorbents, such as limestone, dolomite, lime, sodium bicarbonate, potassium sulfate, etc., to fuel in a burner, in conjunction with lowNOx burners (14-16) or SNCR (17, 18) for NOx reduction. Other technologies employed gas/solid catalytic systems and electron-beam irradiation using electron charging, both using, for example, NH3 as a sorbent. Some of these technologies have been demonstrated throughout the 1990s by U.S.-DOE Clean Coal Program (19). These demonstrations include the SNOX system developed by Haldor Topsoe and ABB and the SNRB system developed by Babcock and Wilcox, both systems belonging to the gas/solid catalytic systems category. Soud (20) compiled a succinct review of these and other systems. Occasionally, removal of SO2 and NOx may generate new environmental problems. For example, using urea in SNCR to reduce NOx can create emissions of ammonia and SO3 (if SO2 is present) and significantly increase N2O emissions (12). Spent catalyst is also considered a toxic pollutant (13). Processes also exist for the simultaneous removal of SO2 and NOx, using one sorbent for the removal of both pollutants. One process is NOXSO (21) that uses a regenerative, dry-injection sorbent (made by spraying Na2CO3 on the surface of alumina spheres) that has been reported to remove 90% of SO2 and 70 to 90% of NOx. Greene et al. (4) evaluated simultaneous SO2/NOx reduction by injecting secondary fuel and SO2 sorbents together, at temperatures as high as 1400 °C, but concluded that optimum overall reductions were obtained when the sorbents were injected with over-fire air at 1150 °C after the hotter secondary fuel “reburning” zone, to minimize sintering. Removal of Particulates. The removal of particulates, such as fly ash and soot, is most commonly achieved by electrostatic precipitators (ESP) with particle removal efficiencies of 95-99%, down to 1 µm particle size; and less commonly achieved by fabric bag filters with particle collection efficiencies of 99% for particles in the neighborhood of 1 µm; by venturi scrubbers; etc. (see Table 7.2 in ref. 22). However, the efficiency of ESPs decreases when power plants switch to cleaner fuels and the flue gas contains little sulfur. Removal of soot agglomerates also takes place by the most efficient 4856

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

of such devices. However, to avoid the formation of dioxins and other air-toxics, particularly in waste-to-energy plants, it may be necessary to remove the unburned carbon and the fly ash at a medium to high temperature (>500 °C) zone of the furnace, instead of the lower-temperature flue gas duct, where the aforementioned devices operate. Work on employing high-temperature filters for removal of particulate emissions from combustors and gasifiers has been recently conducted and reported in the literature, see for instance the work of Miller et al. (23) and references therein as well as that of Andries et al. (24). Removal of HCl and Control of Dioxin Formation. The use of Ca-based sorbents for HCl removal (CaO + 2HCl a CaCl2 + H2O) is discussed in the literature (25-31), for gas temperatures mostly under 600 °C, and by Courtemanche and Levendis (32) and Shemwell et al. (33) at 600-1000 °C. The reactivity of CaO with HCl demonstrated a steady increase with temperature, at reaction rates much faster than those with SO2 (by a factor of 500 (33)). According to Stieglitz et al. (34), formation of polychlorinated dibenzo-dioxins and dibenzo-furans (PCDD/PCDF) occurs in the fly ash as the furnace effluent cools to moderate temperatures (≈300 °C). This is the “de-novo” mechanism, which accounts for most of the PCDD/PCDF formation, with particulate carbon and flue gases containing HCl, O2, and H2O as the main ingredients. Another mechanism, of lesser importance, postulates the formation of PCDD/PCDF from precursors, such as polychlorinated biphenyls, chlorinated benzenes, and pentachlorophenols (35) at moderately high temperatures, 300-800 °C. Previous Work at this Laboratory. Work over the past several years at Northeastern University has demonstrated the effectiveness of the chemical calcium magnesium acetate as well as other organic carboxylic salts of calcium (calcium formate, calcium propionate, calcium benzoate) as dual SO2 and NOx control agents (36-39). These highly water-soluble salts (≈30 wt %) can be wet- or dry-sprayed in the furnace as nonregenerative dry-scrubbing agents. When CMA calcines, the acetate evaporates and leaves porous, thin-walled cenospheres of Ca and Mg oxides and (depending on the temperature) CaCO3. The organic acetate further decomposes to acetone and, eventually, to hydrocarbon radicals that, in an oxygen-lean atmosphere, can reduce NOx to N2 (37, 40). The porous (in the order of 70%) thin-wall cenospheres are ideal sorbents for the heterogeneous reaction of CaO with SO2. Injection of CMA resulted in both SO2 and NOx concentration reductions in the flue gas as high as 95% under optimum conditions. However, the current price of CMA is an order of magnitude higher than that of limestone or lime. Hence, maximizing the utilization efficiency of the carboxylic salts of calcium sorbents (degree of conversion to calcium sulfate or chloride) is imperative. [Commercially available CMA is expensive ($500-800/ton) due to the relatively high cost of producing acetic acid from natural gas and due to the limited markets. This cost may not be offset by the minimal capital investment and the combining of two separate processes for SO2 and NOx removal into one. Therefore, it appears necessary to develop processes for producing acetic acid from renewable organic or “biomass” substrates including woody biomass and wastewater treatment sludge (41-43). If such production becomes commercially viable, the cost of CMA could drop drastically. An economic study (44) estimated that the price of CMA may change from a currently optimized cost of ≈$350/ton to an actual revenue, if appropriate “credits” for accepting and using municipal waste sludge are taken into account.] Moreover, while sorbents derived from CMA or from other carboxylic salts of calcium have a rather ideal structure for the heterogeneous sulfation reactions (porous and thin-walled cenospheres created by the evolution of the pyrolysates), other less

expensive sorbents may perform adequately if their residence time in the furnace is increased appropriately. This is a primary goal of the method described herein. In this work, a spectrum of sorbents, including the costly carboxylic acid salts of calcium as well as the inexpensive limestone and lime, are injected upstream of a ceramic filter. The effectiveness of different sorbents in reducing SO2 and NOx is compared under various conditions. The high residence times of the sorbent in the filter are expected to enhance the concentration reduction of acid gaseous emissions (SO2, HCl). The degree of such enhancement and the cost of the sorbent will finally dictate whether a carboxylic acid or merely a mix of limestone and organics, such as pulverized coal (45) or biomass, will be used to meet the requirements of a certain application. Furthermore, the filter is expected to retain unburned solid carbon (soot) and fly ash. Recent tests in this laboratory have shown silicon carbide (SiC) honeycomb filters (manufactured by Ibiden) to retain PM2.5 particulates at an efficiency 97-99%, in the exhaust of a diesel engine (46). Thus, these filters may be installed at the exit of a furnace, at an appropriate location, where the temperature is sufficiently high for the sulfation and chlorination reactions to occur and result in a high sorbent utilization. Injection of the sorbent blend will take place upstream of the filter at a higher temperature zone to expedite the reactions involved in the reduction of NOx. This reduction zone will be followed by a brief oxidation zone prior to the filter. Hence, the prevailing conditions in the filter will be oxidizing and any collected carbon will be burned therein. Removal of ash, soot, and HCl from the effluent of the hot filter will minimize formation of dioxins in incinerators and other furnaces. In the work presented in this manuscript experimental and theoretical studies were undertaken to (a) explore the combined capture of SO2 and HCl pollutants by calciumbased sorbents, (b) assess the enhancement of the SO2 capture afforded by the additional residence time of the sorbent in the high-temperature filter, and (c) explore the effects of the filter on the NOx reduction.

2. Experimental Apparatus and Procedure An experimental facility was designed and assembled for this work. It consists of two electrically heated, drop-tube furnaces (manufactured by ATS) placed in series (4.8 kW and 1.5 kW max). A simplified schematic showing key features of the vertical furnace assembly is depicted in Figure 1. To introduce sorbents into the furnace, a bed of particles was placed in a vibrated glass vial (test tube), which was advanced by a constant velocity syringe pump (Harvard Apparatus). Sorbent particles at the top of the bed were entrained in a regulated stream of a known nitrogen and oxygen mixture and entered the furnace through a water-cooled stainless steel injector. The injection rate of the powders was in the range of 0.03-0.3 g/min, depending on the sorbent. A background mixture of SO2 and NO gases (diluted in nitrogen) flowed into the furnace through an annulus concentric to the injector, upstream of its tip. This gas stream was preheated in a flow straightner. The combined input flowrate of the particle carrying gas and the background gas was 4.0 lpm. Reactions between the sorbent and the background contaminant gas occurred under laminar-flow conditions in the 25 cm hot zone of the first furnace at residence times ≈1 s. Furnace wall temperatures (Tw) were continuously monitored by type-S thermocouples embedded in the wall. Gas temperatures (Tg) inside the furnace were measured at various axial and radial positions by an aspirated shielded thermocouple (suction pyrometer) (39). The gas temperature profile along the centerline of the upper furnace was found to be fairly isothermal in the radiation zone (see ref 47, Figure 2, and ref 39); Tw - Tg ≈ 50 K. The gas temperature then cooled to ≈250 °C at the flip-flop valve between the two furnaces,

FIGURE 1. Schematic of the electrically heated primary laboratory drop-tube furnace, followed by the secondary furnace, containing the ceramic filter, and then by the expansion chamber. and, subsequently, it was raised again in the second furnace which contained the ceramic filter monolith. The gas temperature in the second furnace, inside the filter channels, was found to be 100 °C lower than the recorded furnace wall temperatures. Three sets of furnace/filter gas temperatures were mostly explored in these experiments: 1423/1173, 1223/ 1173, and 1023/973 K (1150/900, 950/900, and 750/700 °C). During the experiments the pressure drop across the filter monolith at a flowrate of 4 lpm was less than 1 mbar. Reactions continued to occur inside the ceramic honeycomb filter, which was placed in a stainless steel rectangular tube at the centerline of the second furnace, see Figures 1 and 2. The filter was a square cross-section ceramic wall flow honeycomb monolith, 3.5 × 3.5 cm, 8 cm long, see Figure 2. Both cordierite (2MgO‚2Al2O3‚5SiO2) and silicon carbide (SiC) filter monoliths were used, supplied by Corning/ Cermamem and Ibiden, respectively. The monolith was fixed into place in the tube using narrow strips of expandable ceramic matting (Unifrax) and high-temperature epoxy (Zircar), around the perimeter of its two tips. Most of the length of the monolith was not insulated to allow heat transfer from the furnace. The filter monoliths quantitatively capture the spent sorbent particles that are supermicron in size. Moreover, the particulate collection efficiency of these filters in capturing submicron particles has been previously measured to be excellent, 97-99%, in experiments involving diesel engine soot (48, 49). The filter was regenerated (cleaned) at the end of the injection period of each experiment. [In previous experiments with the filter, upon termination of the sorbent injection, the SO2 gas flow was maintained in order to assess the time required to return to the SO2 baseline in the furnace. Those tests showed that the sorbent which was captured in the filter continued to be sulfated for a long period of time (50).] The two furnaces were coupled with an externally operated stainless steel flip-flop valve, see Figures 1 and 2. During normal operation, this valve is kept open, and sorbent particles flow downward from the top furnace to the filter in the bottom furnace. During regeneration, the flip-flop valve closes to seal the exit of the furnace, see Figure 2, and compressed air bursts flow in the direction opposite the normal flow direction. These bursts of air send the particles VOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4857

FIGURE 2. Schematic of the aerodynamic regereration (cleaning) arrangement for the ceramic filter. A depiction of the flow through a honeycomb filter element is also included.

TABLE 1. Experimental and Theoretical Results in Simultaneous SO2-HCl-NOx Removal by CMA Sorbent temp (K)

res time (s)

Ca/Cl

1023 1223 1223

1 1 1

2.3 3.2 3.0

% HCl removed actual predicted 70 85 79

74-92 69-89 51-70

Ca/S 2.7 1.8 2.4

collected in the filter to an expansion chamber. The particles are collected by a glass fiber filter at the bottom of that chamber. For the experiments that were performed without the presence of the ceramic filter monolith the whole second furnace/valve assembly was removed. The spent sorbents were collected on a glass fiber filter placed at the end of the first furnace cooling zone, at ≈100 °C. The gaseous effluent of the furnaces passed through an ice-bath condenser and a Permapure dryer to remove moisture. Thereafter, it was channeled to continuous flow analyzers to measure emissions of SO2 (Rosemount ultraviolet), NOx (Beckman chemiluminescent NO/NOx), CO and CO2 (Horiba infrared), and O2 (Beckman paramagnetic). A limited number of simultaneous SO2-HCl-NOx removal tests were performed without the filter, injecting sorbents in a background gas containing 900 ppm SO2, 400 ppm NOx, and HCl in the vicinity of 600 ppm. The effluent was monitored by a real-time Gas Filter Correlation HCl Analyzer (Thermo Environmental Instruments, Inc.), model 15C with a nondispersive IR spectrometer. The output of all analyzers was recorded using an Omega analog-to-digital converter in a microcomputer. The signals from the analyzers were recorded for the duration of each experiment and were later converted from volts to partial pressures. In each experiment the sorbent injection lasted for 10-20 min. In these experiments, the following calcium-based sorbents were used in pulverized form: calcium formate (CF) [Ca(COOH)2], obtained from Fluka in spheroid particles 3540 µm; calcium propionate (CP) [Ca(CH2CH2COOH)2] obtained from Aldrich consisted of highly irregular particles with a wide size distribution, in the range of 10-200 µm; calcium magnesium acetate (CMA) [CaMg2(CH2COOH)6] obtained from Cryotech was ground and sieved in 60-90 µm size particles; calcium carbonate [CaCO3] from Aldrich in rectangular crystals under 10 µm; and calcium oxide (CaO) 4858

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

% SO2 removed actual predicted 72 27 34

1 5-7 6-9

% Ca utilization actual predicted 38 26 25

12-15 14-18 14-19

% NOx removed 30 54 67

from Mallinckrodt in particles under 5 m in size. The matrix of the experimental conditions of the tests presented herein are shown in Tables 1 and 2, along with the results. The listed physical properties were measured by Micromeritics Labs, as described by Steciak et al. (51). Sorbent particle feed rates were chosen targeting stoichiometric Ca/S molar ratios (Ca/S ) 1) and superstoichiometric ratios, where calcium oxide is in excess, in the following sulfation reaction scheme:

CaO + SO2 + 1/2O2 h CaSO4 The following equations also apply to the chlorination reaction CaO + 2HCl h CaCl2 + H2O. The organic salts of calcium pyrolyze, i.e., release their organic content in the temperature range of 573 to 733 K (300 to 460 °C), depending on the sorbent, and generate calcium carbonate (CaCO3). At temperatures above 973 K (700 °C), CO2 begins to evolve, and CaO forms (as well as MgO in the case of CMA). Under appropriately fuel-rich conditions the released organics may reduce NOx to molecular nitrogen. For instance, the decomposition of the organic acetate in CMA produces acetone and/or allene (52, 40). Reactions of nitric oxide (NO) and nitrogen dioxide (NO2) with acetone and/or allene may occur as follows:

8NO + C3H6O h 4N2 + 3CO2 + 3H2O 4NO2 + C3H6O h 2N2 + 3CO2 + 3H2O 8NO + C3H4 h 4N2 + 3CO2 + 2H2O 4NO2 + C3H4 h 2N2 + 3CO2 + 2H2O SO2 removal was calculated as a time integral of the amount of SO2 fed into the furnace minus the concentration

TABLE 2. Partial Listing of Experimental and Theoretical Results for SO2 Removal by Calcium-Based Sorbents with and without the Filtera model input parameters

temp (K)

Ca/ S

sorbent shell surf res size Mg/ thickness/ area time (µm) porosity Ca radius (m2/g) (s)

1023 1023 1023 1023 1223 1223 1223 1423 1423 1423 1423 1023/973 1023/973 1023/973 1023/973 1023/973 1223/1173 1223/1173 1223/1173 1423/1173 1423/1173 1423/1173 1423/1173 1423/1173 1423/973 1423/973 1423/973 1423/973 1423/873 1423/873 1023 1023 1223 1423 1023/973 1023/973 1223/1173 1223/1173 1423/973 1423/1173 1423/1173 1423/1173 1023 1023 1223 1423 1023/973 1023/973 1023/973 1223/1173 1223/1173 1223/1173 1223/1173 1423/1173 1423/1173 1423/1173 1023/973 1023/973 1023/973 1223/1173 1223/1173 1423/1173

4.3 4.2 2.1 4.4 4.3 1.6 2.2 2.9 2.9 2.0 2.1 2.7 3.5 1.3 1.8 2.3 3.3 2.6 2.7 3.7 3.3 5.1 2.9 5.5 1.4 2.5 3.7 4.3 2.1 4.0 1.0 2.7 0.8 1.2 2.7 1.3 2.9 1.5 3.2 1.5 3.4 3.5 2.4 3.2 2.8 4.0 2.8 3.4 3.6 3.6 3.9 2.4 3.6 5.4 2.1 2.3 4.7 1.9 1.8 1.3 3.2 3.6

35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 35-40 45-63 75-90 45-63 45-63 75-90 75-125 75-100 75-100 75-100 75-100 75-100 75-100 75-100 75-100 75-100 75-100 75-100 75-100 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 2-5 1-7 1-7 1-7 1-7 1-7 1-7

sorbent filter? O (act) CF

CMA

CP

CC

CX

a

N N N N N N N N N N N Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y N N N N Y Y Y Y Y Y Y Y N N N N Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y Y

0.02 0.05 0.03 0.07 0.05 0.02 0.05 0.04 0.06 0.03 0.06 0.03 0.11 0.02 0.03 0.03 0.06 0.05 0.02 0.09 0.09 0.06 0.04 0.05 0.28 0.43 0.7 0.81 1.11 2.58 0.16 0.44 0.24 0.32 0.35 0.22 0.41 0.27 0.98 0.25 0.36 0.53

0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.72 0.66 0.66 0.66 0.66 0.66 0.66 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.79 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15 0.15

0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.375 0.3 0.3 0.3 0.3 0.3 0.3 0.5 0.5 0.5 0.5 0.5 0.5 0.5 0.5 0.5 0.5 0.5 0.5 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1 1

14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 14.2 27 27 27 27 27 27 22.8 22.8 22.8 22.8 22.8 22.8 22.8 22.8 22.8 22.8 22.8 22.8 2 2 2 2 2 2 2 2 2 2 2 2 2 2 2.5 2.5 2.5 2.5 2.5 2.5

1 1 1 1 1 1 1 1 1 1 1 1320 720 1080 1080 900 600 600 600 480 480 360 480 720 900 900 720 780 1200 780 1 1 1 1 900 900 900 720 900 900 720 900 1 1 1 1 900 900 720 900 900 900 900 900 900 900 1080 1200 900 900 840 900

pgm run time (s) 1 1 1 1 1 1 1 1 1 1 1 720 720 130 1080 900 240 240-360 160 480 192 144 288 288 900 900 720 780 1200 780 1 1 1 1 720-900 1080 360-540 288-432 900 1080 288 360-1080 1 1 1 1 900 900 144 56-133 180 180 180 180-540 180 180 180 960-1200 720-900 180 168 180

% SO2 removal % Ca utilization prepreactual dicted actual dicted 35 33 19 17 28 16 30 49 48 33 34 88 87 41 65 50 99 93 74 92 99 98 90 95 58 86 70 88 75 92 6 17 15 17 77 23 68 55 51 52 56 42 3 2 10 33 29 47 41 55 51 22 42 82 41 33 79 43 29 51 95 98

2 2 1 3 18-19 8-9 12-13 34-37 75-80 29-31 30-32 99 83 42 96 66 100 100 100 100 100 100 100 100 72 96 100 100 59 80 0.4 1 2 6-8 99 59-62 99 99 73 95-98 100 99 1 1 3 16-17 39-40 44-48 59-71 50-71 73-82 55-68 76-87 91-99 55-68 67-78 35-57 40-45 26-28 55-68 54-80 66-88

8 8 9 4 6 10 14 17 17 17 16 32 25 32 36 22 30 35 27 25 30 19 31 17 41 34 19 20 36 23 6 7 18 15 29 18 23 36 16 33 16 12 1 1 3 8 13 14 11 15 13 10 12 15 19 14 30 40 29 70 53 49

% NOx reductn

1 1 1 1 4-5 5-6 5-6 12-13 26 14-16 14-15 16 19 18 29 33 26 34 33 21 21 17 33 24 53 38 27 23 28 20 0.3 0.3-0.4 2-3 5-7 17 46-48 32 37 31 64-65 25 24 1 1 1 4 9 8-9 14-16 9-11 12-13 16-20 13-15 11 16-19 13-16 18-29 10-11 10-11 16-19 19-27 15-20

14 0.8 0 3 5 0.8 6 3 12 5 4 3 34 2 7 0 6 0.3 2 6 13 3 2 0 24 26 66 56 55 80 2 9 9 9 14 10 10 82 11 36

Model input parameters are included. Experimentally determined NOx reduction values are also listed.

centration values in time intervals of ≈1 s.

of SO2 detected by the analyzer.

SO2 removal )

∫

n

t)1

[

[SO2]baseline -

(

)]

[SO2]t + [SO2]t-1 2

dt (1)

The SO2 analyzer through a microcomputer recorded con-

Calcium utilization, i.e., the fraction of the solid sorbent that reacted with SO2 to form (presumably) calcium sulfate, was calculated for each sorbent. Three different values of calcium utilization were calculated: the actual sorbent utilization, the maximum theoretical sorbent utilization, and the relative sorbent utilization. The actual sorbent utilization was calculated as VOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4859

Cautil. actual )

Ca S

[SO2]removal effic. Ca / actual S stoichiometric

[( )

( )

]

(2)

where (Ca/S)stoichiometric ) 1. For the HCl capture experiments (Ca/Cl)stoichiometric ) 1/2. The maximum theoretical calcium utilization depends on the actual Ca/S ratio. For Ca/S of 1 or less, maximum theoretical utilization is 100%. Maximum theoretical utilization for Ca/S greater than 1 depends on the ratio and is calculated by

Cautil. max

Ca ( [ ] S) ) Ca [( S ) ] stoichiometric

(3)

actual

The relative sorbent utilization is defined as the actual sorbent utilization divided by the maximum theoretical sorbent utilization which, interestingly, turns out to be equal to the sulfur removal efficiency.

Cautil. rel )

Cautil. actual ) Cautil. max

[

[]

[SO2]removal effic. ‚ Ca S actual Ca S stoichiometric

( ) ( )

(CaS) (CaS)

actual

stoichiometric

]

)

[SO2]removal effic. (4)

3. The Filter Bed Sulfation Kinetic Model The ceramic honeycomb filter collects individual sorbent particles and subjects them to the contaminant stream for an extended period of time. The filter model is based on a single particle sorbent model (53) that has evolved from the pore structure theory of Simons and Finson (54) and Simons (55). The single particle model and its recent modifications are described prior to its extension to the filter bed model. 3.1. Single Particle Model. Following the pore structure theory of Simons and Finson (54) and Simons (55), consider a spherical porous particle of radius a, containing pores of length lp and radius rp. The pore dimensions range from a microscale of the order of Angstroms to a macroscale, which is a significant fraction of the particle radius. The radius of the largest pore is denoted by rmax and is given by

rmax )

2aθ1/3 3Ko

(5)

where θ is the total porosity of the particle and Ko is a constant of integration, approximately equal to five, which relates the pore length to its radius:

lp )

K o rp θ

1/3

(6)

The radius of the smallest pore is denoted by rmin and is given by

rmin )

2θ βFsSp

(7)

where Fs is the density of either the solid CaCO3 or the solid CaO matrix, sp is the specific internal pore surface area (in m2/g), and 4860

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

β ) ln

( ) rmax rmin

(8)

The particle contains a continuous distribution of pore sizes from rmin to rmax. The number of pores within an arbitrary plane of cross-sectional area A and with radius between rp and rp + drp is denoted by gj(rp)Adrp. The pore distribution function gj(rp) is given by

gj(rp) )

θ 2πβr3p

(9)

where gj(rp) indicates an average over all inclination angles between the axis of the pore and the normal to the plane. Due to the random orientation of the pores, the intersection of a circular cylinder with a plane is an ellipse of average area 2πrp2. Hence, the porosity is the 2πrp2 moment of g˜(rp) and the internal surface area is the 4πrp moment of gj(rp). The expression for gj(rp) was derived (54) from statistical arguments and has been validated through extensive comparison of the predicted volume and surface area distributions with mercury intrusion data (56). This has been accomplished for coal, char derived from that coal (57), sorbents, catalysts, and even kidney stones from both men (oxalate) and women (phosphate). The pore volume distribution corresponding to the 1/rp3 distribution function is similar to that utilized in the random pore model (58, 59). However, the pore tree model and the random pore model differ dramatically in their choice of the pore aspect ratio (length to diameter) and its implications with respect to pore branching. The random pore model allows a single pore to connect two larger pores. This picture lends itself to the idealization of instantaneous mixing between the pores and requires that the pore aspect ratio be of the order of 100. The pore tree theory uses data for rmax to imply (via Ko) that all pores possess an aspect ratio of the order of ten. Hence, small pores may connect to larger pores only on one end and all pores must branch from successively larger pores such as a tree or river system. Each pore that reaches the exterior surface of the particle is depicted as the trunk of a tree. The size distribution of tree trunks on the exterior surface of the particle is denoted by gj(rt)4πa2drt where gj(rt) is functionally identical to gj(rp). Each trunk of radius rt is associated with a specific tree-like structure with continuous branching to ever decreasing pore radii. The radius and number of pores is a unique function of the distance x into the tree. The coordinate x is skewed in that it follows a tortuous path through the branches of the tree. Let n(x) represent the number of pores of radius rp at location x in a tree of trunk radius rt. An analysis (55) of this pore tree has demonstrated that

n(x) ) rt2/rp2(x)

(10)

and the coordinate x is related to rp by

drp/dx ) - rp/lt

(11)

The pore structure model described above has been coupled to a transport equation (55) that is capable of describing the simultaneous action of diffusive and kinetic processes within the pore tree. The net sorption rate of a pore tree is determined via integration of the transport equation from rt to rmin. The net sorption rate per particle is generated through integration of the rate per pore tree over all trees identified with the exterior surface of the particle. The features of the pore tree model are best verified through its transport properties. Simons and Garman (53) have utilized the pore tree structure to describe SO2 retention by precal-

cined grains of CaO. Their model includes an intrinsic rate constant for a first-order sulfation reaction, pore transport limitations, and the role of the CaSO4 product deposits in filling and plugging the pores. Simons, Garman, and Boni (6) derived and verified the intrinsic rate constant from a database that encompassed 3 orders of magnitude in sorbent particle size and SO2 partial pressure and precalcined CaO from three sources (carbonate, hydrate, and pressure hydrate). The agreement of the model with this extensive database indicates that the enhanced reactivity of the pressure-hydrated sorbents depends predominately on the superior intraparticle transport and not on the chemical composition of the source of the CaO. It is this characteristic that suggests that the intrinsic rates may be equally valid for CaO derived from organic calcium salts and that an effective way to enhance calcium utilization is to create better intraparticle transport through spray drying of water soluble organic calcium solutions to control the sorbent pore structure (60, 61). 3.2. Modifications to the Single Particle Model. The single particle model described above has been extended in previous work (51) to describe the retention of SO2 by porous cenospheres of arbitrary shell thickness and pore structure. The cenosphere model has also been extended to describe the simultaneous retention of SO2 and HCl by a porous calcium-based sorbent (33). The model uses the known rates of reaction between the CaO and SO2 and assumes a firstorder reaction between HCl and CaO. The SO2 and HCl react in independent, parallel competition for the CaO but combine to form a CaCl2/CaSO4 deposit layer in the filter. It has been demonstrated (62) that CaO ions diffuse through the CaSO4 deposit layer to react with gas-phase SO2. The model assumes that the CaO ions diffuse through the CaCl2/CaSO4 deposit layer fast enough to react with the SO2 and HCl on the gaseous side of the deposit layer at a rate limited by the gas phase diffusion and the heterogeneous kinetics (first-order reaction). Once the outer edge of the particle is completely plugged with solid deposits, the only way the reaction can proceed is via the solid state diffusion of CaO ions through an external deposit layer of ever increasing thickness. This solid state diffusion is terminated when the thickness of the external deposit layer exceeds an empirically inferred critical value of approximately 100 nm. The reaction rate parameters activation energy and preexponential factors for the sulfation reaction were set to be 17,000 K and 600 (g gas/cm2/s/atm gas), and the corresponding values for the calcination reaction were set at 17,000 K and 300,000 (g gas/cm2/s/atm gas), respectively. Thus, the kinetics of the calcination reaction were found in previous work in this laboratory (33), to be much faster (≈500 times) than those of the sulfation reaction. 3.3. Entrained Flow Simulation. The single particle sorption model has been used to simulate the sorption process in our flow tube reactor. The calcium utilization (u) of the particle, the growth of the product layer (δ) within the particle, and the background SO2 mass fraction (c) are simultaneously calculated throughout the entrained flow sorption process. The single particle model denotes M/M as the grams of SO2 removed per second per gram of calcined sorbent. It follows from Simons and Garman (53) that

du M ˙ Mcs ) dt M Mj

(12)

dδ ) f(u, δ, t) dt

(13)

dc Ca du ) -c0 dt S dt

(14)

[ ]

and

[ ]

where Mcs and Mj are the molar weights of the calcined stone and the reactant gas, respectively, c0 is the baseline SO2 mass fraction, and Ca/S is the calcium to sulfur molar ratio. It is important to note that eqs 12 and 13 are both valid following the particle, but eq 14 was derived following the gas. They are identical in an entrained flow system but not in the bed simulation. For later reference, the ratio of the solid to gas mass flow within the entrained stream is

Msolid Ca Mcs ) c0 Mgas S Mj

[ ]

(15)

3.4. Filter Bed Simulation. If an entrained particle latent stream is impacting a porous filter, a cake of particles will build up on the filter. If one stands on the interface between the filter cake and the entrained stream, one will observe the filter moving away at constant velocity. Each layer of particles between the interface and the filter is uniquely identified by its exposure time starting with t ) 0 at the interface and increasing to the total filter exposure time (t) at the filter. The decrease in the SO2 mass fraction at each point in the filter cake may be expressed as

dc -M ˙ /M ) dMsolid M ˙ gas

(16)

where Msolid is increasing in time due to the upstream particle flux

Msolid ) Msolidt

(17)

Combining eqs 15-17, the time dependence of the SO2 mass fraction is given by

dc -M ˙ Ca Mcs ) -c0 dt M S Mj

[ ][ ]

(18)

As eqs 12 and 13 were written following the particle, they are valid for the filter cake as well, and eqs 12, 13, and 18 represent the description of the filter bed. However, combining eqs 12 and 18, we recover eq 14. Hence, the models are mathematically the same, provided that time (t) is the exposure time of the particle in both systems. The filter bed model has been constructed to follow the simultaneous decay of both SO2 and HCl, if present, in the filter bed.

4. Results and Discussion 4.1. Simultaneous SO2-HCl Removal Experiments. The effectiveness of the sorbents to remove SO2 emissions has been assessed in our laboratory over the years (36, 39, 51), and their effectiveness to remove HCl has been documented recently (32, 33). The effectiveness of the sorbents to achieve a compounded removal of SO2 and HCl was demonstrated with a limited number of tests without the filter, employing CMA as the sorbent. These tests are summarized in Table 1 and illustrated in Figure 3, where the accompanying NOx removal efficiency (30-67%) is also included. It appears that the sorbent is, indeed, capable of partially removing all three pollutants with degrees of efficiency depending on the conditions. As expected, based on the aforementioned kinetics, HCl is effectively captured (70-85%), while SO2 is partially captured (27-72%). The total calcium utilization values are not very high (25-38%), since the sorbent was supplied in significant excess, as shown in Table 1. To calculate the utilization values, the calcium chlorination was assumed to occur and then the sulfation reactions. A calciumto-chlorine ratio was calculated based on the furnace input, upon partial conversion of calcium with chlorine a new calcium-to-sulfur ratio was calculated for the sulfur reaction. VOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4861

FIGURE 3. Experimental results from tests on the simultaneous SO2-HCl-NOx reduction by CMA sorbent, using the data listed in Table 1. Partial pressures were recorded at the exit of the furnace. Asterisks denote the start of sorbent injection. Theoretical predictions were successful for the chlorination reactions but underpredicted the sulfation reactions, especially at the lower temperatures. As a result, the model also underpredicted the calcium utilization at the lower temperature. 4.2. SO2 Removal at the Presence or Absence of the Filter. As mentioned before, most of these experiments were conducted at gas temperature combinations of 1423/1173, 1223/1173, and 1023/973 K in the upper and lower furnaces, respectively. The residence time of the sorbents in the upper furnace was under 1 s, while in the second furnace the sorbent particles accumulated in the filter. Therein, the initial particles resided for the full period of each test (10-20 min), while the 4862

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

last sorbent particles resided only momentarily, before they were removed by regenerating (back-pulsing) the filter. The particles were fluidized and introduced to the furnace where the gas, in most experiments, contained 1200 ppm of SO2, 600 ppm of NO, and 3% O2 in a balance of nitrogen. Partial results are listed in Table 2 and are also shown in Figure 4. The parameter SO2 removal efficiency also represents the relative calcium utilization. This parameter is appropriate for presenting the results, since it illustrates the performance of each sorbent relative to its maximum potential. On the other hand, the parameter actual calcium utilization is also important as it reflects the actual amount of the sorbent that is utilized in the process; the rest is wasted. Three trends are

FIGURE 4. Experimental results from tests on SO2 removal efficiency and sorbent utilization of calcium formate (CF), calcium(magnesium)acetate (CMA), calcium propionate (CP), calcium carbonate (CC), and calcium oxide (CX) sorbents, using the data listed in Table 2. evident in the experimental results. (1) The SO2 removal efficiency always increased at the presence of the filter. Often enhancements by factors of 2 to 3 were achieved or even higher in the case of the nonporous sorbents. The sorbent utilization increased accordingly. The main reason for this enhancement is that the drastically longer residence time of the sorbent in the furnace improves both the gas-phase pore diffusion of SO2 and the solid-phase diffusion of the CaO ions, as hypothesized earlier. The lowest SO2 removal efficiencies with the filter were higher than the best efficiencies without the filter under comparable other

conditions. For instance, SO2 capture efficiencies ranged 4199% (17-68%) for CF and 37-80% (3-33%) for CC with or (without) the filter. Maximum actual utilization for CF increased from 17% without the filter to 36% with the filter, as shown in Table 2. Similarly for CP maximum utilizations increased from 18 to 37%. For CMA the maximum calcium utilization was measured to be 60%. (2) At any fixed temperature, the SO2 removal efficiencies increased with the Ca/S ratio, i.e., with the amount of sorbent injected in the furnace. To the contrary, the sorbent utilization typically decreased with the Ca/S ratio, in the superstoiVOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4863

unfavorable both from chemical equilibrium considerations and physical structure considerations. At elevated temperatures, sintering of the sorbents may commence. In Figure 5, experimental profiles for both SO2 and NOx reduction are shown for one of the most successful experiments of this work. They are included to illustrate the potential of this technique, as injection of CMA at a Ca/S ratio of 2 and a 1223/1173 gas temperature combination resulted in a SO2 capture of 80% and NOx reduction of 80%.

FIGURE 5. Time-resolved profiles for SO2 and NOx reduction by calcium(magnesium)acetate (CMA) at 1223/1173 K furnace/filter gas temperature, Ca/S ) 2. Partial pressures were recorded at the exit of the series of the two furnaces. Asterisks denote the start of sorbent injection. chiometric regime. This is probably due to the depletion of SO2 locally, at high Ca/S ratios where dense clusters of particles are present. This observation suggests that there should be a Ca/S ratio where a compromise may be achieved between the environmental need to maximize the SO2 removal efficiency and the financial need to maximize the degree to which the sorbent is utilized. (3) Both, the SO2 removal efficiency and the sorbent utilization, increase with increasing gas temperature in the range of 973-1473 K. This increase is explained based on the temperature-enhanced kinetics of the sulfation reactions in this temperature window. Thus, to obtain optimum results the sorbent injection point as well as the filter should be kept at locations of high temperature. Excessive temperatures, above the aforementioned range, are undesirable as they are

In summary, the particulate collection efficiency for these filters is quantitative for PM10 particles and very high, 9799%, for PM2.5 particles. The HCl capture by the sorbents is excellent, and the presence of the filter may only provide a limited enhancement. The SO2 capture efficiency by the sorbent, however, is greatly enhanced by the filter, by factors of 2, 3, and even higher. Increasing the Ca/S ratio improves the sulfur reduction efficiency at the same filter temperature but decreases the sorbent utilization. Higher gas temperatures at the location of the sorbent injection point and the filter (in the region tested: 1023-1423 K and 973-1173 K, respectively) improved the SO2 reduction efficiency and utilization. The NOx reduction efficiency by those sorbents that have a hydrocarbon content was high at oxygen-lean, high-temperature conditions and was mildly enhanced by the filter. Overall, reduction efficiencies in the order of 90% could be achieved for all pollutants by injection of the sorbent upstream of the filter. However, sorbent utilization efficiencies were mostly under 40%. Thus, there is still room for improvement, and future work will undertake an optimization study. Generally, the process is not yet optimized. It is expected that better dispersion of the sorbents in the radial direction of the drop-tube furnace and mixing with the sorbent will improve sorbent utilizations. This will be addressed in future work. Moreover, sorbent utilizations are expected to improve with longer residence times in the filter, as the importance of the initial filter-coating period will be minimized. It is important to realize that while fairly high filter temperatures are desirable for this process, locating the filter in the high temperature region in a furnace may not be practical, as it may interfere with the radiative heat transfer therein. Moreover, at high temperatures a larger volumetric flowrate of (more expanded) gas will pass through the filter, and, thus, a larger filter will be needed to maintain an acceptable pressure drop. Hence a compromise needs to be achieved and will be addressed in future work. Furthermore, future work will test the effectiveness of less costly blends of

FIGURE 6. Experimental results on the NOx reduction efficiency of calcium formate (CF), calcium propionate (CP), and calcium(magnesium)acetate (CMA), using the data listed in Table 2. 4864

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

limestone-coal-fines as sorbents for powerplants in conjunction with filters. Finally, an economic analysis will be undertaken to estimate the size and the cost of the filter, for various types/sizes of powerplants. 4.3. Comparison of Kinetic Model Predictions with Experimental Results. As mentioned above a single particle sulfation model was modified to simulate a filter bed, representing a cake of sorbent particles building up on the filter walls. This modified model was used to predict SO2 reduction and calcium utilization (actual) for sorbent/gas interaction over the extended time period allowed by the filter. An activation energy of 17,000 K and a preexponential factor of 600 g gas/cm2/s/atm gas were used for the sulfation reaction. The sorbent properties, such as sorbent size and porosity, shell thickness to radius, and surface area, that were input to the kinetic model are included in Table 2. These properties were measured in previous work (39, 51), but they are subject to particle-to-particle variability. The model’s calculations on the theoretically predicted SO2 removal efficiency and the calcium utilization are also listed in Table 2 and are compared with the experimentally derived values. Overall, the agreement between theory and experiment is satisfactory under the majority of conditions examined. A notable exception is that the model underpredicts the experiments at the lowest temperatures in the absence of the filter. Most sorbents under such conditions should have converted to Ca(CO)3 but may have not fully converted to CaO. The model utilizes kinetic expressions for the sulfation of CaO, not Ca(CO)3, thus, the discrepancy. Indeed, the agreement is very good for the case of CaO sorbent. Furthermore, in most cases, the agreement can be further improved by adjusting the input physical parameters. For instance, initial particle size input was measured from microscopic observations on a limited number of particles used in these experiments. However, there is considerable particle-to-particle size variation and agglomeration or fragmentation may also occur during fluidization, further altering the particle size distributions. The overall agreement between the theory and experiment is deemed sufficient to allow the model to be used as a guiding tool. 4.4. Reduction of NOx. NOx is reduced only by sorbents that contain organics, i.e., the carboxylic salts of calcium, such as CF, CP, CMA. As most of the NOx reduction activity of these sorbents appears to be nonselective, sufficient amounts of organic components are needed to consume the oxygen in the furnace. When the overall conditions in the furnace are fuel-rich, i.e., the equivalence ratio is higher than unity, substantial reduction of NOx is expected. [Please note that the bulk equivalence ratio is defined as φ ) (mfuel/ mair)actual/(mfuel/mair)stoichiometric. The bulk equivalence ratio is equal to the inverse of the excess air ratio, λ.] At low oxygen partial pressures, NOx in the flue gas is reduced by hydrocarbon radicals from the decomposition of the carboxylic acids. NOx reduction by CF was not technically significant (generally 3-8%), while those by CP and CMA were good to excellent. The results are shown in Figure 6. NOx removal varied from 11 to 82% for CP and CMA, depending on the conditions. NOx reduction reactions are homogeneous and, therefore, are fast. The 1 s residence time in the first furnace appears to be nearly-adequate for NOx reduction since the presence of the filter in the secondary furnace only mildly enhances the reduction of NOx. The primary furnace temperature, i.e., the gas temperature at the injection point, has a marked effect on the NOx reduction. High temperatures are desirable.

Acknowledgments This research was funded by NSF Grant # BES-9505703, Dr. Edward Bryan, program director. The authors acknowledge help from Mr. Ali Ergut. Ibiden and Corning Corporations

kindly provided the Silicon Carbide and cordierite filter elements, respectively. The authors also want to thank Mr. Tom Cosgrove and Mr. John Wilbur for the use of the Thermo Environmental Instruments’ HCl analyzer 15C and for their kind technical assistance during the project period.

Literature Cited (1) Martinelli, R.; Doyle, J. B.; Redinger, K. E. SOx-NOx-Rox-Box Technology Review & Global Commercial Opportunities; Presented at the Fourth Annual Clean Coal Technology Conference, Denver, CO, Sept 5-8, 1995. (2) Cole, J. A.; Kramlich, J. C.; Seeker, W. R.; Silcox, G. D.; Newton, G. H.; Harrison, D. J.; Pershing, D. W. Proceedings: 1986 Joint Symposium on Dry SO2 and Simultaneous SO2 /NOx Control Technologies; Electrical Power Research Institute: 1986. (3) Electric Power Research Institute: 1995. (4) Greene, S. B.; Chen, S. L.; Clark, W. D.; Heap, M. P.; Pershing, D. W.; Seeker, W. R. EPA/600/7-85/012; 1985. (5) Bruce, K. R.; Gullett, B. K.; Beach, L. O. AIChE J. 1989, 35(1), 37. (6) Simons, G. A.; Garman, A. R.; Boni, A. A. AIChE J. 1987, 211. (7) Simons, G. A. AIChE J. 1988, 34, 167-170. (8) Lyon, R. K. Intl. J. Chem. Kinet. 1976, 8, 315-318. (9) Arand, J. K.; Muzio, L. J.; Texieria, D. P. U.S. Patent Number 4,325,924, 1982. (10) Myerson, A. L. U.S. Patent Number 3,867,507, 1975. (11) Wendt, J. O. L.; Sternling, C. V.; Matovich, M. A. Reduction of Sulfur Trioxide and Nitrogen Oxides by Secondary Fuel Injection; Fourteenth Symposium (International) on Combustion, The Combustion Institute: Pittsburgh, PA, 1973; pp 897-904. (12) Bowman, C. T. Control of Combustion-Generated Nitrogen Oxide Emissions: Technology Driven by Regulation; Twenty-Fourth Symposium (International) on Combustion, The Combustion Institute: Pittsburgh, PA, , 1992; pp 859-878. (13) Explaining Clean Coal Methods to Avoid Coal Use Constraints; COAL-TRANS International: November-December 1999. (14) Steward, F. R.; Couturier, M.; Morris, K. J. Institute Energy 1992, 65, 177. (15) Wang, M. R.; Chang, K. C. Energy 1991, 16(5) 849. (16) Helfritch, D.; Bortz, S.; Beittel, R.; Pergman, P.; Toole-O’Neil, B. Environ. Prog. 1992, 11(1), 7. (17) Gullett, B. K.; Bruce, K. R.; Hansen. W. F.; Hoffman, J. E. Environ. Prog. 1992, 11(2), 155. (18) Darmstaedler, E. Power Eng. 1990, 12, 24. (19) DoE report DOE/FE-0351 on Clean Coal Technology; U.S.-DOE: 1996. (20) Soud, H. N. Developments in Flue Gas Desulfurization, DRAFT; IEA Coal Research, Gemini House: London, U.K., 2000. (21) Yeh, J. T.; Ma, W. T.; Pennline, H. W.; Haslbeck, J. L.; Joubert, J. I.; Gromicko, F. N. Chem. Eng. Comm. 1992, 114, 65. (22) Flagan, R. C.; Seinfeld, J. H. Fundamentals of Air Pollution Engineering; Prentice Hall: Englewood Cliffs, NJ, 1988. (23) Miller, B. G.; Wincek, R. T.; Glick, D. C.; Scaroni, A. W.; Makris, P.; Krecker, J.; Jung, G.; Stubblefield, D. J. Proceedings of the 23th International Technical Conference on Coal Utilization and Fuel Systems; Clearwater, FL, 1998. (24) Andries, J.; Unal, O.; Hoppesteyn, P. D. J.; Hein, K. R. G. HighTemperature Dust Filtration Downstream of a Coal/Biomass Fuelled Gasifier Using a Ceramic Channel Flow Filter; 23rd International Conference on Coal Utilization and Fuel Systems, Clearwater, FL, 1998. (25) Mayer-Schwinning, G.; Laibold, E. Proceedings of the 1989 International Conference on Municipal Waste Combustion; EPA600/R-92-052c; 1989; Vol. 3. (26) Ellison, W. Utilization of Flue-Gas Cleaning Technologies for Waste Incinerators in Europe; International Conference on Municipal Waste Combustion, 1989; Vol. 3, p 7c-43. (27) Walters, J. K.; Daoudi, M. Management of Hazardous and Toxic Wastes in the Process Industries; Elsevier: 1987. (28) Gullett, B. K.; Jozewicz, W.; Stefanski, L. A. Ind. Eng. Chem. Res. 1992, 31, 2437. (29) Weinell, C. E.; Jensen, P. I.; Dam-Johansen, K.; Livbjerg, H. Ind. Eng. Chem. Res.1992, 31, 164. (30) Mura, G.; Lallai, A. Chem. Eng. Sci. 1992, 47, 2407. (31) Jozewicz, W.; Gullett, B. K. Ind. Eng. Chem. Res. 1992, 31 607. (32) Courtemanche, B.; Levendis, Y. A. Environ. Sci. Technol. 1998, 134, 407-432. (33) Shemwell, B.; Levendis, Y. A.; Simons, G. Chemosphere 2000, in press. (34) Stieglitz, L.; Zwick, G.; Beck, J.; Roth, W.; Vogg, H. Chemosphere 1989, 18, 1219. (35) Wiliams, P. T. J. Institute Energy 1992, 65, 46. VOL. 34, NO. 22, 2000 / ENVIRONMENTAL SCIENCE & TECHNOLOGY

9

4865

(36) Levendis, Y. A.; Zhu, W. G.; Wise, D. L and Simons, G. A. AIChE J. 1993, 39, 761. (37) Steciak, J. Zhu, W. G.; Levendis, Y. A.; Wise, D. L. Combustion Sci. Technol. 1994, 102, 193. (38) Steciak, J.; Levendis, Y. A.; Wise, D. L.; Simons, G. A. Dual SO2 -NOx Reduction by Fine Mists of CMA; Nineteenth International Conference on Coal Utilization and Fuel Systems, Clearwater, FL, 1994; p 553. (39) Steciak, J. Levendis, Y. A.; Wise, D. L. AIChE J. 1995, 41-3, 712. (40) Shuckerow, J. I.; Steciak, J. A.; Wise, D. L.; Levendis, Y. A.; Simons, G. A.; Gresser, J. D.; Gutoff, E. B.; Livengood, C. D. Resources, Conservation Recycling 1996, 16, 15. (41) Wise, D. L.; Levendis, Y. A.; Metghalchi, M. M. CMA: An Emerging Bulk Chemical for Environmental Purposes; Elsevier Scientific Publishers: The Netherlands, 1991. (42) Wise, D. L.; Augenstein, D. Solar Energy 1988, 41, 453. (43) Trantolo, D. I.; Gresser, J. D.; Augenstein, D. C.; Wise, D. Resources, Conservation Recycling 1990, 4, 215. (44) Palasantzas, I. A.; Wise, D. L. Resources Conservation Recycling 1994, 11, 225-243. (45) Steciak, J. Ph.D. Thesis, 1995. (46) Larsen, C.; Levendis, Y. A. Filtration Assessment and Thermal Effects on Aerodynamic Regeneration in Silicon Carbide and Cordierite Particulate Filters; SAE publication 1999-01-0114; 1999; Vol. SAE SP-1414. (47) Atal, A.; Steciak, J.; Levendis, Y. A. Fuel 1995, 74, 495. (48) Oey, F.; Mehta, S.; Levendis, Y. A. Diesel Vehicle Application of an Aerodynamically Regenerated Trap and EGR System; SAE publication 950370; 1995; Vol. SAE SP-1073, pp 103-116. (49) Larsen, C.; Levendis, Y. A. On the Effectiveness and Economy of Operation of ART-EGR Systems that Reduce Diesel Emissions; SAE publication 980537; 1998; Vol. SAE SP-1313, pp 97-115. (50) Levendis, Y. A.; Atal, A. Hot Flue Gas Filtration: A New Development; Proceedings of the 24th International Conference on Coal Utilization and Fuel Systems; Clearwater, FL, March

4866

9

ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 34, NO. 22, 2000

1999. (51) Steciak, J. Levendis, Y. A.; Wise, D. L.; Simons, G. A. J. Environ. Eng. 1995, 121, 595. (52) Su, F. Process for Manufacturing Crystalline Calcium Magnesium Acetate; International Patent, WO 90/15112, PCT/US90/02766; 1990. (53) Simons, G. A.; Garman, A. R. AIChE J. 1986, 32, 1491. (54) Simons, G. A. and Finson, M. L. The Structure of Coal Char: Part I Pore Branching. Combustion Sci. Technol. 1979, 19, 5 and 6, 217-226. (55) Simons, G. A. The Pore Tree Structure of Porous Char; Proceedings of the 19th Symposium (International) on Combustion, 1982; p 1067. (56) Stacy, W. O.; Walker, P. L., Jr. Structure and Properties of Various Coal Chars; Coal Research Section, College of Earth and Mineral Sciences, The Pennsylvania State University; University Park, PA, 1972. (57) Kothandaraman, G.; Simons, G. A. Evolution of the POre Structure in PSOC 140 Lignite During Pyrolysis; Proceedings of the 20th Symposium (International) on Combustion, 1984; pp 15231529. (58) Gavalas, G. R. AIChE J. 1980, 26, 577. (59) Gavalas, G. R. Combustion Sci. Technol. 1981, 24, 197. (60) Simons, G. A. Resource, Conservation, Recycling 1991. (61) Simons, G. A.; Parker, T. E.; Moore, T. E.; Senior, C. L.; Levendis, Y. A. Combined NOx/SO2 Control Using a Single Liquid Injection System; PSI-2186/TR-1169; Prepared for the Department of Energy, Contract No. DE-FG02-91 ER8118; 1992. (62) Hsia, C.; St. Pierre, G. R.; Raghunathan, K.; Fan, L.-S. AIChE J. 1993, 39, 4, 698-700.

Received for review March 8, 2000. Revised manuscript received August 10, 2000. Accepted August 25, 2000. ES001072T