Modeling the behavior of selenium in Pulverized-Coal Combustion systems

Combustion and Flame 157 (2010) 2095–2105

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Modeling the behavior of selenium in Pulverized-Coal Combustion systems Constance Senior *, Brydger Van Otten, Jost O.L. Wendt, Adel Sarofim Reaction Engineering International, 77 W. 200 South, Salt Lake City, UT 84101, USA

a r t i c l e

i n f o

Article history: Received 21 February 2010 Received in revised form 2 May 2010 Accepted 4 May 2010 Available online 22 May 2010 Keywords: Selenium Pulverized-Coal Combustion Model

a b s t r a c t The behavior of Se during coal combustion is different from other trace metals because of the high degree of vaporization and high vapor pressures of the oxide (SeO2) in coal flue gas. In a coal-fired boiler, these gaseous oxides are absorbed on the fly ash surface in the convective section by a chemical reaction. The composition of the fly ash (and of the parent coal) as well as the time–temperature history in the boiler therefore influences the formation of selenium compounds on the surface of the fly ash. A model was created for interactions between selenium and fly ash post-combustion. The reaction mechanism assumed that iron reacts with selenium at temperatures above 1200 °C and that calcium reacts with selenium at temperatures less than 800 °C. The model also included competing reactions of SO2 with calcium and iron in the ash. Predicted selenium distributions in fly ash (concentration versus particle size) were compared against measurements from pilot-scale experiments for combustion of six coals, four bituminous and two low-rank coals. The model predicted the selenium distribution in the fly ash from the pilot-scale experiments reasonably well for six coals of different compositions. Ó 2010 The Combustion Institute. Published by Elsevier Inc. All rights reserved.

1. Introduction Selenium is used in a variety of industrial applications, the largest being the manufacture of glass and ceramics. Selenium is used to tint glasses and glazes red, as well as to ‘‘de-colorize” iron-containing glasses. Selenium is used in photocopying and solar cells, because of its photovoltaic properties. Selenium is also used as a toner for photographic prints to extend the tonal range of black and white photographs. Biologically, selenium is an essential trace element, but is toxic in larger doses. The Tolerable Upper Intake Level of selenium for humans is 400 mg/day; exceeding this intake can lead to selenosis [1]. The bioavailability of elemental selenium and most metallic selenides is low. Oxyanions of selenium (selenates and selenites), in contrast, are very toxic [2]. Some soils are naturally high in selenium, which can result in high levels of selenium in water bodies and a resulting impact on flora and fauna. For example, high selenium levels have been found to cause certain congenital disorders in wetland birds [3]. Human activities can also contribute to selenium contamination of water bodies [4], including: irrigation runoff from high-selenium soils; discharges from coal-fired power plant sources such as ash ponds; discharges from coal mining or hard rock mining; and effluents from petroleum refineries. In December 2004, the US Environmental Protection Agency (EPA) proposed revised water quality criteria recommending safe * Corresponding author. Fax: +1 801 364 6977. E-mail address: [email protected] (C. Senior).

levels for short-term and long-term selenium exposure in both freshwater and saltwater environments. EPA released a draft ambient water quality criterion for selenium in October, 2008. The criteria recommendations were intended to protect aquatic life under the Clean Water Act [5]. Selenium is also listed as a hazardous air pollutant under the Clean Air Act. Currently there are no limits on air emissions of selenium from combustion sources. However, US EPA is in the process of setting maximum achievable control technology (MACT) standards for coal-fired power plants, which could include standards for air emissions of selenium. A review of full-scale data on selenium partitioning in coal-fired boilers showed that from 27% to greater than 90% of the selenium in bituminous-coal-fired boilers exited the electrostatic precipitator (ESP), presumably in the gaseous phase [6]. Fig. 1 shows the partitioning for these boilers with the coal sulfur content indicated. In pilot-scale coal combustion studies, Seames and co-workers observed evidence for reactions between selenium and both iron and calcium in the fly ash, with interference from SO2 [7,8]. Ghosh-Dastidar et al. [9] carried out TGA studies on the sorption of SeO2 gas on hydrated lime and other sorbents at temperatures from 400 °C to 1000 °C. They tested hydrated lime, kaolinite, bauxite and limestone for sorption capacity with SeO2(g) in N2 and in air. In air, hydrated lime captured more SeO2 than the other sorbents. The capacity of hydrated lime for SeO2 peaked at 600 °C. Desorption studies showed that the reaction product decomposed at temperatures greater than 825 °C. Using several different experiments, the authors concluded that these two reactions were responsible:

0010-2180/$ - see front matter Ó 2010 The Combustion Institute. Published by Elsevier Inc. All rights reserved. doi:10.1016/j.combustflame.2010.05.004

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Nomenclature cfs d f ko kr np r t x A C D Ea Fasa Fi Kn MW P R S T k

l

Fuchs–Sutugin correction factor (see Eq. (8)) ash particle diameter mass fraction pre-exponential factor for rate constant rate constant number of ash particles in a given size bin reaction rate time conversion fraction of the active surface area concentration (g/g) gas-phase diffusivity activation energy fraction of active surface area flux to a particle of size i Knudsen number, 2k/d molecular weight, g/mol vapor pressure universal gas constant surface area temperature mean free path in gas

h u

q w

Subscripts a ash property c coal property f final fg flue gas i particle size bin o initial s conditions at particle surface Se selenium SeO2 selenium dioxide 1 conditions in bulk gas Superscripts m exponent for surface area n exponent for pressure

CaðOHÞ2 ) CaO þ H2 OðgÞ;

ð1Þ

CaO þ SeO2 ðgÞ ) CaSeO3 :

ð2Þ

Li and co-workers carried out a series of studies on sorption of SeO2 and SO2 on CaO [10,11]. Li et al. obtained a reaction order for SeO2 + CaO of 1 at 600–700 °C, which increased to 1.52 at 800 °C. The reaction order for SO2–CaO was 1. The rate constant for Se was greater than the rate constant for sulfur at temperatures less than 720 °C, but dropped sharply above this temperature, probably because CaSeO3 decomposed at higher temperatures. Calcium hydroxide has been compared with sodium carbonate as a sorbent for selenium in the glass industry. Selenium is often used as an additive in the manufacture of flint glass to control color. Some of the selenium vaporizes in the glass furnace. Kircher [12] compared sodium and calcium sorbents by injecting the sorbents in the off-gases of a glass furnace for temperatures between 330 °C and 390 °C. At these temperatures, Ca(OH)2 was a good

140%

Fraction of Se

120%

Units with FGDs

viscosity in gas phase fraction of selenium, in coal vaporized mass fraction of submicron ash density accommodation coefficient, fraction of successful collisions with the particle surface

sorbent for selenium, but Na2CO3 was not as effective. No fundamental information on sodium sorbents has been found in the literature to date. Selenium is soluble in silicate glasses, and is used in the glass industry to ‘‘de-colorize” glass, by reacting with iron to form a Fe–Se compound. Fang and Lynch [13] measured the solubility of selenium in silicate slags in equilibrium with Cu–Se alloy. As the amount of Fe2+ increased in the slag (relative to Fe3+), the solubility of Se increased in the slag. They reported 0.2–3 wt.% Se in the slags at temperatures of 1185 °C and 1250 °C. Their results were consistent with a model that had been proposed, in which selenium was postulated to exist in both a neutral state and as an Fe–Se complex in the slag. In a coal-fired boiler, selenium might dissolve in molten silicate particles in the flame. The presence of iron in the silicates, particularly under reducing conditions, could promote the formation of Fe–Se compounds, which might serve to retain selenium in the

Stack Gas Scrubber Out ESP Ash Econ. Ash Bottom Ash

100% 80% 60% 40% 20% 0%

Fig. 1. Selenium partitioning in bituminous coal-fired plants with coal sulfur content in wt.% indicated on the graph.

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2. Model description The model is one-dimensional, and requires the initial condition of the system: fly ash size and composition, coal composition, time–temperature profile, vapor-phase selenium concentration, and initial particle selenium concentration. The initial selenium concentration on the ash particles is computed assuming zero concentration for all submicron particles, which are assumed to be formed from ash vaporization followed by homogeneous condensation and agglomeration. Selenium not vaporized is assumed to be uniformly distributed on the supermicron ash particles and is computed as

C Seo;i ¼

ð1  hÞ C o  ; ð1  uÞ C a

ð3Þ

where h is the fraction of selenium in coal vaporized, u is the mass fraction of submicron ash, Co is the concentration of selenium in coal (g/g), and Ca is the mass fraction of ash in the coal, taken from the ultimate analysis. C Seo;i is the initial concentration of selenium in g/g on size i. The model requires as input the amount of selenium vaporized in the flame. From full-scale data [6], it was noted that Se vaporization in the combustion zone ranged from 65% to 100% (Fig. 2). Given the observation of high-temperature reaction between Se and

100%

Se Vaporized in Combustion

ash particles. This might explain the correlation noted by Seames and Wendt [7] between iron content and selenium content in certain bituminous fly ash samples. Selenium retention by both fly ash and activated carbon has been studied in simulated coal combustion flue gas and gasification syngas. López-Antón et al. have studied capture of As and Se at 120 °C by carbon-containing fly ash as well as by activated carbons [14,15]. Jadhav et al. [16] studied removal of Se by activated carbon in the temperature range of 125–250 °C. Reactions between vapor-phase selenium and carbonaceous materials are outside the scope of the present work, which focuses on reactions between selenium and inorganic fly ash. The behavior of selenium in coal-fired power plants is relevant to air emissions of this hazardous air pollutant, as well as to emissions of selenium in wastewater discharges (ash ponds, scrubber water treatment). There is a need for a fundamentally based, predictive model for vapor–solid partitioning of selenium in coal-fired boilers in order to predict selenium concentrations in fly ash and vapor-phase selenium entering FGD scrubbers. Previously, a model was developed for the gas-to-solid conversion of arsenic in coal-fired boiler exhaust gases [17]. This was a one-dimensional model that included arsenic condensation and reaction with calcium on the fly ash as temperatures decrease in the post-combustion section of the boiler. The model took into account the available surface area of calcium in the fly ash as well as the competition between arsenic and sulfur for calcium surface sites. The model was intended to predict the concentration of arsenic in the vapor phase at the inlet to the air preheater. The model was validated against pilot-scale data for partitioning of arsenic in fly ash from combustion of six different coals. In the case of selenium, reaction with both iron and calcium on the fly ash surface (as well as competition with SO2) should be taken into account. The pilot-scale data of Seames and Wendt [7,8] were used to test this model. In this set of experiments, six coals (of different ranks and sulfur contents) were burned and the associations between selenium and other fly ash constituents were measured using chemical analysis of size-segregated fly ash. Furthermore, doping experiments were carried out in which the coal was spiked with additional sulfur, iron or calcium, in order to separate the effects of these potential reactants.

80% 60% 40% 20% 0%

Fig. 2. Se vaporization in the combustion zone calculated from full-scale data [6].

Fig. 3. Relationship between coal iron and Se vaporization, from full-scale data.

iron, there may be limited interaction between Se and iron-containing ash in the burning char particle, which might reduce the amount of selenium vapor that escapes during combustion and thus reduces the apparent vaporization of selenium. A relationship between the iron content of the coal and Se vaporization was derived from full-scale data and is shown in Fig. 3. This relationship will be used to estimate vaporization of Se. In the model, the non-vaporized portion of the Se will be distributed among the supermicron ash particles. 2.1. Gas–solid interactions The vapor-phase selenium is transported to the particle surface where it interacts by one of two mechanisms, heterogeneous condensation or surface reaction. These mechanisms are computed as fluxes, with units of moles SeO2 per second per particle. Only one mechanism is assumed to occur for a given set of conditions. The equations for each mechanism depend on the gas-phase transport regime: free molecular (submicron) or continuum (supermicron). Thus, four flux equations are used to describe the transport. The concentration of selenium on a given ash particle at a given time tf is given by the following equation:

C Sei ¼ C Seo;i þ

MW Se

 p  q  d3 i i 6

Z

t¼tf

F i ðtÞ  dt:

ð4Þ

t¼0

Here, qi and di are the ash particle density, and diameter, respectively, which are both assumed constant. The total particle mass is also assumed constant, since the total mass of selenium is much less than the mass of the ash particle. Selenium concentrations

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are measured in lg/g. Eq. (4) is integrated numerically for each particle size, allowing for changing gas and transport properties with time. At each time step, the moles of SeO2 removed from the flue gas is computed as

DnSeO2 ¼

X

F i  npi  Dt;

ð5Þ

i¼sizes

where Fi is the flux for the given size, referred to above, and defined below, and npi is the number of ash particle of size i for our given basis of 1 g of coal. npi is computed as

npi ¼

C a  ni pq

i

6

3

 di

ð6Þ

:

The following equations are used to describe the flux of SeO2 to the particle surfaces. 2.1.1. Heterogeneous condensation, continuum regime The equation used to describe the condensation flux in the continuum regime is given by:

Fi ¼

  2p  di  D 1 þ Kn ;  ðP1  Ps Þ  RT 1 þ 1:71  Kn þ 1:33  Kn2

ð7Þ

where Kn = 2k/di is the Knudsen number, D is the diffusivity of SeO2, P1 and Ps are the bulk and surface pressures of SeO2, respectively. The term in brackets in Eq. (7) is the Fuchs–Sutugin correction for the non-continuum behavior of small particles. For large particles the term becomes unity. The mean free path is computed as



k¼

l p  MW fg  q 2RT

2

Fi ¼

ð8Þ

The diffusivity of SeO2 is computed using the empirical equation of Fuller, Schettler, and Giddings [18], which utilizes molecular diffusion volumes. The diffusion volume of selenium was not available so the value was calculated by plotting the atomic radius versus the diffusion volume of eight elements and using this result to estimate the diffusion volume for selenium. The surface pressure, Ps is computed as the equilibrium vapor pressure of the gas in the following reaction: SeO2(s) = SeO2(g). Fig. 4 shows a plot of the vapor pressure versus temperature. The relatively high vapor pressure at low temperature implies condensation will only occur at very low temperatures, and relatively high concentrations of selenium in the vapor phase. 2.1.2. Heterogeneous condensation, free molecular regime The flux in the free molecular regime (submicron particles) is given by

Fig. 4. Vapor pressure of SeO2.

ð2  p  MW  R  TÞ1=2

;

ð9Þ

where w is the accommodation coefficient, or the fraction of successful collisions with the particle surface, taken as unity. 2.1.3. Surface reaction, continuum regime The flux equation for the surface reactions is given by the following equation:

2  p  di Fi ¼ RT

"

# P1 ; 2 1 þ Dcfs di kSe Fasa

ð10Þ

where kSe is the reaction rate constant in units of (m/s), Fasa is the fraction of the active surface area, where reactions are assumed to occur between the SeO2, which is an oxy-anion, and calcium cations at the particle surface. Fasa is assumed equal to the mass fraction of CaO in a given ash size. The same equation is used for the reaction between selenium and iron. Eq. (10) is written in terms of the bulk selenium partial pressure, rather than the surface pressure, so the diffusive resistance is added to the reactive resistance in the denominator of the term in brackets. The corresponding surface pressure for reaction-only is computed by

Ps ¼

2  D  P1 : 2  D þ d  kSe  Fasa

ð11Þ

2.1.4. Surface reaction, free molecular regime The reaction flux for the free molecular regime is given by

1=2 :

w  p  d  ðP1  Ps Þ

Fi ¼

pd2 RT

 kSe  P1  Fasa:

ð12Þ

For the free molecular regime, there is no diffusion resistance, and the surface pressure is equal to the bulk pressure. The reaction fluxes presented are not expected to occur in parallel, rather, only heterogeneous condensation, or surface reaction occurs, depending on the conditions. For temperatures above 800 °C, neither mechanism will occur for calcium. For iron, the surface reaction was assumed to take place at temperatures above 1200 °C. As discussed above, if the equilibrium vapor pressure is greater than the bulk partial pressure, the heterogeneous condensation flux is assumed zero; negative values are not allowed. For the surface reactions, if the surface pressure, Eq. (11) (or the bulk pressure for the free molecular regime) is greater than the equilibrium vapor pressure, the reaction fluxes are assumed zero, and only heterogeneous condensation occurs. 2.2. Surface reaction kinetics Ghosh-Dastidar et al. [9] carried out TGA studies on the sorption of SeO2 gas on hydrated lime and other sorbents at temperatures from 400 °C to 1000 °C. The capacity of hydrated lime for SeO2 peaked at 600 °C. Desorption studies showed that CaSeO3 decomposed at temperatures greater than 825 °C. Agnihotri et al. [19] continued the same type of TGA experiments as in Ref. [9], but they determined the activation energy and exponents from their TGA data. They also looked at simultaneous reactions of SO2 and SeO2 with hydrated lime. Even when the gas phase concentrations of SeO2 and SO2 were comparable, sulfation (at 600 °C) was eight times faster than selenition. This conclusion is somewhat contradicted by the experimental results of Li et al., discussed below, at least for the chemical kinetic-controlled regime of sorption. Li and co-workers carried out a series of studies on sorption of SeO2 and SO2 on CaO [10,11]. One study was designed to

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understand the impact of the product sulfate layer on sorption of SeO2 and SO2 [10]. In some cases, sorbent was pre-sulfated to build up a product layer, to determine if this would interfere with the reactions of CaO with SeO2 and SO2. The loss of sorption rate of SO2 at high CaO conversion was only significant when the concentration of SO2 was higher than 100 ppmv. This led the authors to conclude that sorption of species present in lower concentrations, like SeO2, would not be affected by the sulfate product layer. Experiments with 26 ppmv SeO2 showed no effect of the degree of CaO sulfation on sorption. Thus, the product layer did not seem to be what reduced SeO2 sorption in the presence of SO2. In a more detailed study of SO2 and SeO2 sorption [11], the authors identified two stages of the SO2–CaO reaction: (1) initial chemical kinetic control at low CaO conversion; (2) late-stage (high CaO conversion) control by diffusion through the product layer. The authors hypothesized that in stage (1) the number of active sites, which can react with SO2 or SeO2, was large enough so that there was no competition. But in stage (2), the number of active sites was limited and there was competition between SO2 and SeO2. Li et al. obtained a reaction order for SeO2 + CaO of 1 at 600– 700 °C, which increased to 1.52 at 800 °C. The reaction order for SO2–CaO was 1. The observed rate constant for Se was greater than the rate constant for sulfur at temperatures less than 720 °C, but dropped sharply above this temperature, probably because CaSeO3 decomposed at higher temperatures. Agnihotri et al. [19] gave the following equation for the SeO2– CaO reaction:

  Ea ; kf ¼ kfo  ðSo ð1  xÞÞm  ½SeO2 n exp RT

ð13Þ

where x is the conversion of CaO and So is the initial surface area of the particle. Agnihotri et al. [19] determined that Ea = 16.849 kJ/gmol, m = 0.25 and n = 0.66 from data taken in the range of 400–600 °C. For this modeling effort, it is assumed that m = 1 and n = 1 (in Eq. (13)) and that the activation energy is equal to 16.84 kJ/gmol. Li et al. [11] observed that above 650 °C, the reaction rate fell to zero by 800 °C. The fall in reaction rate with increasing temperature is indicative of the decomposition of the reaction product, CaSO3, as has been noted by Fan and co-workers. Above 800 °C, therefore, the reaction rate will be set to zero. The reaction rate for Se with iron will use the same form as Eq. (13). The sulfur present in flue gas is able to react with calcium or iron in ash to reduce the reactivity of the ash towards selenium. This important effect is included in the model. Li and co-workers noted the simultaneous removal of SeO2 and SO2 in their TGA experiments [10,11]. Yang and Shen [20] measured sulfation rates of crystalline calcium silicates at temperatures up to 900 °C and reported that the rates were comparable to those of the reaction between SO2 and limestone. Graham also observed sulfation of high-calcium fly ash in laboratory experiments with high-calcium fly ash [21]. Kang and co-workers [22] measured sulfation rates in a laboratory experiment using ash from an Eagle Butte subbituminous coal. They calculated a rate constant ks (in units of kg m2 s1 MPa1) that was first-order in PSO2 (in units of MPa) as

kS ¼ 6:23  eð9210=TÞ :

ð14Þ

The reaction of CaO with SO2 is reported to be a two-step process, first forming CaSO3 and then forming CaSO4. Simons and Boni [23] reported the rate of SO2 reaction with CaO (in units of kg s1) as ð6750=TÞ

r ¼ 7PSO2 Fasa  e

;



T < 827 C;

ð15Þ

r ¼ 6x104 PSO2 Fasa  eð17;000=TÞ ;

827  C < T < 1227  C:

ð16Þ

Fly ash has a mixture of CaO (free calcium) and calcium aluminosilicates. The submicron particles that contain calcium, which represent the largest fraction of the total ash surface area, are more likely to consist of CaO than calcium aluminosilicate, since calcium aluminosilicates are formed by the coalescence of supermicron mineral particles during combustion. For the model, Eqs. (15) and (16) will be used to calculate the reaction rate of SO2 with calcium surfaces. All the sulfur in coal is assumed to form SO2, and the SO2 flux is computed using the flux equations for surface reaction, as for selenium, listed in Eqs. (10) and (12). In these equations, the diffusivity and partial pressure of SO2 replace those of selenium oxide. The reaction of iron and SO2 will be modeled similarly. The same activation energy as for the calcium–SO2 reaction will be used, in the absence of experimental data. There are four reactions in the model that occur simultaneously. In calcium-containing ash particles, both SeO2 and SO2 react with the surface. The Ca–Se reaction is only active at temperatures less than 800 °C. The activation energy for the Ca–Se reaction is taken from Fan and co-workers. The pre-exponential factor for this reaction is taken from the previous arsenic model. The Ca–SO2 kinetic parameters were also taken from the previous arsenic model [17]. In iron-containing ash particles, both SeO2 and SO2 are assumed to react with the surface. The activation energy for the Fe–Se reaction was set to be high (50 kJ/mol K) to reflect the bias toward high-temperature reaction. The pre-exponential factor was fit to experimental data. The Fe–Se reaction was assumed to occur only at temperatures greater than 1200 °C, because high-temperature reaction with iron was observed in the laboratory data, as discussed below. The activation energy for the Fe–SO2 reaction was the same as that for the Ca–SO2 reaction, since there were no data on this reaction. Reaction rate parameters are summarized in Table 1. 2.3. Model inputs The inputs to the model can be split into three groups: (1) coal composition and time–temperature profile, (2) ash size distribution and properties, and (3) initial selenium concentration and speciation. In order to compute the selenium partitioning, the composition and time versus temperature profile of the flue gas are required. The flue gas composition is obtained by specifying the coal ultimate analysis (weight percent of C, H, O, N, S, Cl, Ash, Moisture), the combustion air composition, and the stoichiometric ratio, assuming complete combustion. The time–temperature profile is input directly. A basis of 1 g of as-received coal is taken as a basis for the calculation. For input group two, the post-combustion ash size distribution is specified as particle diameter versus mass fraction. In addition, ash density, calcium content and iron content for each size are specified. The group three inputs require the sele-

Table 1 Parameters used in selenium model. Parameter

Units

Comments

ko Ea

8.8 16,860

m/s J/mol

Ca–Se reaction Only <800 °C

ko SO2 Ea SO2 ko SO2 Ea SO2

0.000117 56119.5 1 141,338

moles/s/m2/Pa J/mol moles/s/m2/Pa J/mol

Ca–SO2 reaction T < 827 °C

koFe EaFe

6000 50,580

m/s J/mol

Fe–Se reaction Only >1200 °C

ko FeSO2 Ea FeSOj2

0.1 141,338

moles/s/m2/Pa J/mol

Fe–SO2 reaction

827 °C < T < 1227 °C

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nium concentration in the coal, and fraction of selenium volatilized during combustion. Because the selenium is present in only trace quantities in the coal, the volatilized selenium as SeO2 is added to the flue gas directly, without regard to the material balance defined by the major coal constituents. In coal-fired boilers, the SeO2 in the flue gas will be on the order of 1 ppmv or less. 3. Model validation 3.1. Validation data set Pilot-scale data were taken from experiments carried out at the University of Arizona [7,8]. The experimental facility was a 2.2 kg/h refractory-lined, coal-fired furnace. A quench probe was used to take size-segregated ash samples near the flame (upper) and at the end of combustion (lower). A series of experiments were carried out in which a coal was burned and size-resolved ash samples were taken. Each data set contained the following data:  Coal ultimate and proximate analysis, plus ash chemistry and trace element content.  Time–temperature history of the furnace for individual coals.  Low-pressure impactor data (0.03–15 lm) at each of the three sampling locations, including mass distribution, major elements and trace elements. Six coals were burned in this study, and their properties are summarized in Table 2. The four bituminous coals had a range of sulfur contents, from 0.8% (Kentucky) to 3.6% (Illinois 6). The content of Fe2O3 in the ash ranged from 5.1% (Kentucky) to 22% (Illinois 6). The Kentucky coal was also very low in calcium (1.8% CaO in ash). The low-rank coals, Wyodac subbituminous and North Dakota lignite, had high CaO in ash – 22.7% and 30.3%, respectively. The North Dakota lignite also had moderate iron in the ash, 11.8% as Fe2O3. The datasets are identified as follows: 1. 2. 3. 4.

Pittsburgh bituminous coal. Illinois 6 bituminous coal. Ohio 5, 6, 7 bituminous coal. Elkhorn/Hazard Kentucky bituminous coal.

5. Wyodac PRB coal. 6. North Dakota lignite. 7. Pittsburgh coal plus injection of Fe-nitrate solution (Fe-doping) in order to increase the Fe in the ash equivalent to doubling the moles of calcium in the ash. 8. Kentucky coal plus added SO2 (SO2-doping) in order to achieve the same SO2 concentration in the flue gas as in the Ohio coal. 9. Ohio coal co-injected with Ca(OH)2 (Ca-doping) in order to set the ratio of Ca/As the same as in the Wyodac coal. 10. Wyodac coal plus injection of Se solution into flame (Se-doping) in order to set the ratio of Se/Ca to be the same as in the Ohio coal, about an 8 times increase in Se. 11. Wyodac coal plus injection of Fe-nitrate solution (Fe-doping) in order to set the ratio of Fe/Se to be the same as in the Ohio coal, about a 2.5 times increase in Fe. Each coal that was burned in the pilot-scale furnace produced its own temperature profile. The bituminous coals produced higher temperatures than the low-rank coals, because of the higher water content of the low-rank coals. Before examining the modeling results, it is useful to review the experimental data, to observe trends and look for underlying mechanisms. The data will be displayed in terms of concentration of Se in lg/g (ppmw) as measured in the different stages of the low-pressure impactor. Each stage is associated with a mean aerodynamic diameter. Low-pressure impactor samples were taken at Port 4 (Upper Samples) and Port 14 (Lower Samples). The temperatures at these locations are given in Table 3. For three of the bituminous coals, the Port 14 temperature was approximately equal to 800 °C, while for the Kentucky coal, the sampling temperature was greater than 800 °C. Thus, we expect to see little reaction with calcium in the bituminous coal data, but ample time for reaction with calcium in the low-rank coal data. In the next series of figures, the experimental data will be examined. The experimental data (concentration of Se in fly ash) have been normalized to the total amount of Se in the coal (Table 2). First, Figs. 5 and 6 show the distribution of Se as a function of particle size for the bituminous coals and low-rank coals, respectively.

Table 2 Properties of coals in University of Arizona experiments [8]. Coal

Pittsburgh bituminous

Ultimate analysis (wt.%) C (%) 76.6 H (%) 4.8 N (%) 1.5 S (%) 1.6 Cl (%) 0.1 O (%) 6.9 Ash (%) 7.0 Moisture (%) 1.4 Ash composition (wt.% ash, S-free) SiO2 45.37 Al2O3 24.18 Fe2O3 20.28 TiO2 1.14 P2O5 0.60 CaO 4.77 MgO 1.03 Na2O 1.35 K2O 1.28 BaO SrO Selenium, lg/g coal

0.62

Illinois 6 bituminous

Kenutcky bituminous

Ohio bituminous

Wyodac subbituminous

North Dakota lignite

67.7 4.7 1.2 3.6 0.0 9.2 10.3 3.3

74.9 4.6 1.4 0.8 0.2 8.4 7.4 2.3

71.1 4.8 1.4 2.6 0.1 8.1 9.7 2.3

51.2 3.6 0.7 0.3 0.0 12.3 6.0 25.8

38.6 2.6 0.4 0.6 0.0 12.5 9.4 35.9

49.40 19.31 22.04 1.01 0.13 4.45 0.95 0.70 2.00

55.00 33.76 5.10 1.68 0.23 1.81 0.59 0.32 1.51

38.09 39.41 13.03 2.04 0.51 2.24 1.02 0.92 2.55 0.20

43.28 17.17 6.32 1.38 1.78 22.74 4.03 1.67 0.45 0.68 0.49

22.00 20.38 11.75 0.50 0.13 30.25 8.00 5.12 1.38 0.50

2.2

3.1

1.4

1.08

1.5

C. Senior et al. / Combustion and Flame 157 (2010) 2095–2105 Table 3 Sampling temperatures from University of Arizona experiments [8]. Coal

Port 4 temperature, °C

Port 14 temperature, °C

Pittsburgh Illinois Kentucky Ohio Wyodac North Dakota

1187 1187 1267 1107 1027 967

792 792 947 797 687 598

For the bituminous coals, the concentration of Se decreased as particle size increased. This pattern suggested that the higher surface area of the submicron particles attracted more Se than the supermicron particles. For the Pittsburgh and Illinois coals, there was evidence that at the high temperature (Port 4) there was vapor-phase SeO2 in the furnace, which nucleated homogeneously in the sampling system to form ultrafine (0.03–0.05 lm) particles. For submicron particles from 0.1 to 1 lm, there was little difference between the shape of the Se distributions between Port 4 and Port 14, which might be because of the reaction of Se with bituminous ash particles at temperatures greater than 1267 °C.

(a) Pittsburgh

(c) Kentucky

There were less data for the low-rank coals; a significant portion of the North Dakota selenium distribution was not available. Nonetheless, it is clear for the Wyodac coal that there was little Se in the submicron particles at Port 4 (1027 °C), but by Port 14 (687 °C), Se concentrations in the submicron particles increased. This is consistent with reaction between calcium and selenium, which takes place at temperatures below 800 °C. Doping experiments were performed in which (a) calcium was added to the coal (as Ca(OH)2); (b) iron was added to the flame by injecting an iron nitrate solution; (c) additional SO2 was added to the burner; or (d) additional Se was added to the flame as a salt solution. When iron was added to the Pittsburgh coal, the iron concentration increased in particles with diameters from 2 to 7 lm. When iron was added to the Wyodac coal, iron concentration increased in particles from 0.5 to 7 lm. When calcium was added to the Ohio coal, the concentration of calcium increased in particles from 0.5 to 7 lm. Fig. 7 gives the results from adding iron to the Pittsburgh coal compared with the baseline; these are results from Port 14. The addition of iron increased the concentration of Se in the ash, particularly at the smallest diameter for which there are data (only data for diameters greater than 0.5 lm were taken). This experimental

(b) Illinois

(d) Ohio

Fig. 5. Se concentrations in ash as a function of ash diameter: baseline experimental results for bituminous coals.

(a) Wyodac

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(b) North Dakota

Fig. 6. Se concentrations in ash as a function of ash diameter: baseline experimental results for low-rank coals.

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Fig. 7. Se concentrations in ash as a function of ash diameter at Port 14: baseline experimental results compared with Fe doping experiments for Pittsburgh coal.

result supports the observation from full-scale data that the apparent vaporization of Se decreases as the iron content of the coal increases. The apparent vaporization represents the fraction of selenium that leaves the flame in the vapor phase. If selenium can react with iron in silicate glasses (as hypothesized here), then this reaction would probably take place in the flame and would therefore result in an apparent decrease in selenium vaporization. Better full-scale data (such as size-resolved measurements of Se, Fe, and Ca in fly ash) should be obtained to confirm this hypothesis. The Ohio bituminous coal was low in calcium. Therefore, calcium was added to the Ohio coal. The concentration of Se in the Ohio ash did not increase consistently in the fly ash (0.5–9 lm) as shown in Fig. 8. Again, the Port 14 temperature might have been too high for significant reaction between Se and calcium. The Kentucky coal was low in sulfur (and iron). Therefore, SO2 was added to the Kentucky coal. As shown in Fig. 9, the concentration of Se in the supermicron ash did not change upon addition of SO2. The Wyodac coal was low in iron and selenium, but high in calcium. Therefore, iron or selenium was added to the Wyodac coal, as shown in Fig. 10. Adding Se increased the Se concentration in the ash, especially in the submicron range. Adding iron (as supermicron particles) increased the concentration of Se in the supermicron particles modestly. Several conclusions related to the mechanism for selenium–ash interactions can be drawn from the experimental data. For most of the bituminous coals, the Se distribution in the ash appears to track the iron distribution. In the bituminous coal experiments, Se was observed in the submicron particles at temperatures greater than 1000–1100 °C, suggesting reaction with iron in the submicron ash at high temperatures. Adding iron or calcium to the bituminous coals increased the concentration of Se in fly ash. Iron had the larger effect on Se concentration. The temperature of the lower furnace sampling location might have been too high to see significant Se–Ca reaction. There is evidence for reaction between Se and calcium in the Wyodac coal at temperatures less than 1000 °C, which is consistent

Fig. 8. Se concentrations in ash as a function of ash diameter at Port 14: baseline experimental results compared with doping experiments for Ohio coal.

Fig. 9. Selenium concentrations in ash as a function of ash diameter at Port 14: baseline experimental results compared with doping experiments for Kentucky coal.

Fig. 10. Selenium concentrations in ash as a function of ash diameter at Port 14: baseline experimental results compared with doping experiments for Wyodac coal.

with the laboratory observation that the reaction product between calcium and selenium is not stable above 800 °C. Adding Se to the Wyodac coal increased Se concentration in all particle sizes. Adding supermicron iron to the Wyodac combustion experiment resulted in a modest increase in Se in the supermicron ash particles. Perhaps this reflects mass transfer limitations, because the Se content of the Wyodac coal was low. 3.2. Modeling results Fig. 11 shows the predicted and measured selenium distributions at Port 14 for combustion experiments with the Pittsburgh coal. The model shows that adding iron to the Pittsburgh coal shifts selenium from submicron to supermicron particles. Fig. 12 shows the predicted and measured selenium distributions at Port 14 for combustion experiments with the Illinois 6 coal. The model predicts the shape of the experimental curve and the absolute concentrations of selenium in the ash. Fig. 13 shows the predicted and measured selenium distributions at Port 14 for combustion experiments with the Kentucky coal. As with the Pittsburgh coal, the predictions appear to overestimate the amount of vaporization of selenium, which results in the predicted supermicron concentrations being lower than the experimental values. It is somewhat hard to understand the increase in Se concentration with the addition of SO2, particularly without data on the submicron particles. Fig. 14 shows the predicted and measured selenium distributions at Port 14 for combustion experiments with the Ohio coal. Again, vaporization might have been overestimated for this coal. Since the Ohio coal was low in calcium, experiments were carried out with the addition of calcium. The model predicts an increase in selenium concentration in the supermicron particles, because the sampling temperature was slightly below 800 °C. This was also observed in the experimental data. Fig. 15 shows the predicted and measured selenium distributions at Port 14 for combustion experiments with the Wyodac

C. Senior et al. / Combustion and Flame 157 (2010) 2095–2105

(a) Baseline Pittsburgh

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(b) Fe addition

Fig. 11. Measured and predicted selenium concentrations in fly ash for Pittsburgh coal.

Fig. 12. Measured and predicted selenium concentrations in fly ash for Illinois 6 coal.

subbituminous coal. The model reproduces the trends of the Wyodac Se distribution from the experiments. When additional Se was added to the coal, the model correctly reproduced the peak apparent in the smallest particles (0.05 lm). Addition of iron to the Wyodac flame was predicted to shift Se from the submicron particles to the supermicron iron-containing particles. The model

(a) Baseline Kentucky

predicted concentrations of Se that matched the experimental data in the supermicron particles; however, there were no submicron data available for this experiment. Fig. 16 compares the predicted North Dakota lignite Se distribution with the experimental. Data are missing for several particle sizes. Thus, it is hard to draw conclusions, but the model appears to reproduce the general slope of the data. Table 4 summarizes the predicted levels of removal of SO2 and SeO2 under the University of Arizona experimental conditions (that is, between Port 4 and Port 14). Seames and co-workers did not report a mass balance for selenium, so it is not possible to estimate the amount of selenium removed by the fly ash in their experiments. There are no data on the removal of SO2 either, but the model predictions give some insight into how the model behaves. Adding Fe to the Pittsburgh coal decreased SO2 and SeO2 removal. SeO2 removal could have decreased because there was more high-temperature removal by Fe and this lowered the mass transfer to the Ca at lower temperatures. Adding SO2 to the Kentucky coal decreased the SO2 removal percentage, but the total amount of SO2 removed was higher by about 2.5 times; SeO2

(b) SO2 addition

Fig. 13. Measured and predicted selenium concentrations in fly ash for Kentucky coal.

(a) Baseline Ohio

(b) Ca addition

Fig. 14. Measured and predicted selenium concentrations in fly ash for Ohio coal.

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(a) Baseline Wyodac

(b) Se addition

(c) Fe addition Fig. 15. Measured and predicted selenium concentrations in fly ash for Wyodac coal.

4. Conclusions

Fig. 16. Measured and predicted selenium concentrations in fly ash for North Dakota lignite.

Table 4 Model predictions for removal of SeO2 and SO2 in University of Arizona experiments.

Pittsburgh Illinois Kentucky Ohio Wyodac North Dakota Pittsburgh-Fe addition Kentucky-SO2 addition Ohio-Ca addition Wyodac-Se addition Wyodac-Fe addition

SO2 removal (%)

SeO2 removal (%)

2.2 1.6 0.2 0.3 13.2 11.7 1.1 0.2 4.8 9.7 4.7

99.8 92.0 56.2 52.7 93.7 99.9 98.9 38.5 54.3 90.2 68.5

removal was reduced, because of competition from SO2. Adding Ca to the Ohio coal increased SO2 removal by more than a factor of 10 and increased SeO2 removal slightly. Adding Se to the Wyodac coal (8 times increase) decreased the percentage removal of SeO2 slightly (but the absolute amount of SeO2 removal increased), and SO2 removal decreased (competition with SeO2). Adding Fe to the Wyodac coal decreased SO2 removal and SeO2 removal. SeO2 removal could have decreased because there was more high-temperature removal by Fe and this lowered the mass transfer to the Ca at lower temperatures.

At flame temperatures, selenium is soluble in silicate glasses. Larger solubility is indicated for iron-containing silicates under reducing conditions. With sufficient calcium in fly ash, selenium might be expected to form a calcium selenite/selenate at temperatures below about 800 °C. Without sufficient calcium or iron (or in the presence of relatively high concentrations of SO2), some of the selenium might be expected to condense as SeO2 at temperatures below about 200 °C. Partial condensation would lead to selenium leaving the particulate control device as gaseous SeO2, which is observed in bituminous coal-fired plants. A model was created for interactions between Se and fly ash post-combustion. The reaction mechanism assumed the following:  Iron reacts with Se at high temperatures (above 1200 °C) – perhaps this reaction is with Fe–Si–Al glasses and only happens for sufficiently high temperatures to result in low viscosity of the ash.  Ca reacts with Se at temperatures less than 800 °C.  SO2 reacts with calcium and iron, but more strongly with calcium. The Se distributions in fly ash (concentration versus particle size) were reasonably well predicted in University of Arizona pilot-scale experiments for six coals. The model could prove useful in estimating the partitioning of selenium between gas and solid at the ESP inlet. The model requires the iron and calcium distribution in fly ash, which was measured in the pilot-scale experiments, but which is not commonly measured in power plants. These distributions could be estimated from the ash composition and coal rank. Acknowledgments The authors gratefully acknowledge financial support for this work from Southern Company, Birmingham, Alabama, USA.

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