Volume II Chapter 3.0 Pages 9 of 10 page next page 10

3.6.17. Coal Combustion 3-151

3.6.18. Combustion of Polychlorinated Biphenyls (PCBs) 3-152

3.6.19. Pyrolysis of Brominated Flame Retardants 3-153

3.6.20. Carbon Reactivation Furnaces 3-154

3.6.21. Cement Kilns 3-156

3.6.22. Additional Combustion and High Temperature Sources 3-165

3.7. RESERVOIR SOURCES 3-166

3.6.17. Coal Combustion

Fiedler and Hutzinger (1992) estimate that 1.1 g of dioxin TEQ may be released to the atmosphere in Germany annually from residential combustion of coal. In the United Kingdom, combustion of coal by residential, industrial, and utility sources is estimated to account for 38 percent (1,489 g TEQ/yr) of all dioxin TEQ releases to the atmosphere (ECETOC, 1992). The Clean Air Act requires an assessment of the emissions of toxic air contaminants (including CDDs and CDFs) discharged from the stacks of coal-fired power plants. The EPA is collaborating with the U.S. Department of Energy to conduct this study. Stack testing at seven plants is currently underway and the results will be incorporated to the extent possible in the final version of this report.

In the United States, the consumption of coal accounts for approximately 25 percent of the energy consumed from all sources (U.S. DOC, 1992). In 1991, 806 million metric tons of coal were consumed in the United States (EIA, 1993). Of this total, 87 percent (or 701 million metric tons ) was consumed by electric utilities, 12.3 percent (or 99 million metric tons) was consumed by the industrial sector, and 0.7 percent (or 5.5 million metric tons) was consumed by residential and commercial sources (EIA, 1993). These production estimates are assigned a "high" confidence rating since they are based on detailed studies specific to the U.S.

Derivation of emission factors is difficult due to the extremely limited data on emission of dioxin-like compounds from coal-fired utility boilers (NATO,1988). Most investigations of emissions from facilities in the United States have not reported the detection of dioxin congeners at the exit to the stack (NATO, 1988). Therefore in the development of emission factors representative of coal-fired utility boilers (coal-fired power plants), the reported limit of detection of the analytical method was applied as an upper-bound to the plausibility of emissions of dioxin from the source.

If it is assumed that burning one kg of coal in a modern power plant produces an estimated 6.2 dry standard cubic meters (dscm) of combustion gas, then the limit of detection from the study of U.S. power plants can be used to estimate an emission factor (NATO,1988). From the reported limit of detection of dioxin at the stack, an upper-bound emission factor for total CDD/CDF of 3.11E-02 g/kg of coal combusted and an upper-bound emission factor for TEQ of 4.22E-04 g/kg of coal combusted can be derived. If it is assumed that 700 billion kg of coal is combusted each year by power plants in the United States, then the upper-bound emission factors indicate an annual emission to the air of less than 2.2E+04 grams of total CDD/CDF and less than 3.0E+02 grams of TEQ/yr.

The emission factors are assigned a "low" confidence rating because the stack emission of dioxin-like compounds from coal-fired utility boilers operating in the United States has yet to be determined. Emissions tests reported to date have not detected these compounds. The estimated emissions must, therefore, be considered the upper-bound of possible emissions from the source category based on available data.

3.6.18. Combustion of Polychlorinated Biphenyls (PCBs)

The accidental or intentional combustion of PCBs in incinerators and boilers not approved for PCB burning (40 CFR 761) may produce CDDs and CDFs. At elevated temperature, such as those in transformer fires, PCBs can undergo reactions to form CDF and other by-products. Several accidental fires in the U.S. and Sweden which involved the combustion of PCBs and the generation of CDDs and CDFs are discussed in Hutzinger and Fiedler (1991b).

For example, analyses of soot samples from a Binghamton, New York office building fire detected 20 g/g of total CDDs (0.6 to 2.8 g/g of 2,3,7,8-TCDD) and 765 to 2,160 g/g of total CDFs with 12 to 270 g/g of 2,3,7,8-TCDF. At that site, the fire involved the combustion of a mixture containing PCBs (65 percent) and chlorobenzene (35 percent). Hutzinger and Fiedler (1991b) also reported that laboratory analyses of soot samples from a PCB transformer fire which occurred in Reims, France indicated total CDD and CDF levels in the range of 4 to 58,000 ng/g and 45 to 81,000 ng/g, respectively.

The use of PCBs in new transformers in the United States has been banned and their use in existing transformers is being phased out. Because of the accidental nature of transformer fires it is not possible to accurately estimate annual emissions from this source.

3.6.19. Pyrolysis of Brominated Flame Retardants

The pyrolysis and photolysis of brominated phenolic derivatives and polybrominated biphenyl ethers used as flame retardants can generate polybrominated dibenzo-p-dioxins (BDDs) and dibenzofurans (BDFs) (Hutzinger and Fiedler, 1991a; Luijk et al., 1992; Watanabe and Tatsukawa, 1987). Watanabe and Tatsukawa (1987) observed the formation of BDFs from the photolysis of decabromobiphenyl ether. Approximately 20 percent of the decabromobiphenyl ether was converted to BDFs in samples that were irradiated with ultraviolet light for 16 hours (Watanabe and Tatsukawa, 1987). Decabromobiphenyl ether is used as a flame retardant in resins, textiles, and paints.

Luijk et al.(1990) studied the formation of BDD/Fs during the compounding/ extrusion of decabromodiphenyl ether into high-impact polystyrene polymer at 275 C. HpBDF and OBDF were formed during repeated extrusion cycles, and the yield of BDFs increased as a function of the number of extrusion cycles (Luijk et al., 1990). HpBDF increased from 1.5 to 9 ppm (in the polymer matrix) and OBDF increased from 4.5 to 45 ppm after four extrusion cycles.

Thoma and Hutzinger (1989) observed the formation of BDFs during combustion experiments with polybutylene-terephthalate polymers containing 9 to 11 percent decabromodiphenyl ether. Maximum formation of BDFs occurred at 400 to 600 C with a BDF yield of 16 percent. Although Thoma and Hutzinger (1989) did not provide specific quantitative results for similar experiments conducted with octabromodiphenyl ether and 1,2-bis(tri-bromophenoxy)ethane, they did report that BDDs and BDFs were formed. Insufficient data are available upon which to derive annual BDD/BDF emission estimates from this source.

3.6.20. Carbon Reactivation Furnaces

Granular activated carbon (GAC) is an adsorbent that is widely used to remove organic pollutants from wastewater and in the treatment of finished drinking water at water treatment plants. Activated carbon is manufactured from the heat treatment of nut shells and coal under pyrolytic conditions (Buonicore, 1992a). The properties of GAC make it ideal for adsorbing and controlling vaporous organic and inorganic chemicals entrained in combustion plasmas, as well as soluble organic contaminants in industrial effluents and drinking water.

The high ratio of surface area to particle weight (e.g., 600 - 1600 m2/g), combined with the extremely small pore diameter of the particles (e.g., 15-25 Angstroms) increases the adsorption characteristics (Buonicore, 1992a). GAC will eventually become saturated and the adsorption properties will significantly degrade. When saturation occurs, the GAC usually must be replaced and discarded, which significantly increases the costs of pollution control. The introduction of carbon reactivation furnace technology in the mid-1980s created a method involving the thermal treatment of used GAC to thermolytically desorb the synthetic compounds and restore the adsorption properties for reuse (Lykins et al., 1987).

The used GAC can contain compounds that are precursors to the formation of CDDs/Fs during the thermal treatment process. The U.S. EPA measured precursor compounds in spent GAC used as a feed material to a carbon reactivation furnace tested during the National Dioxin Study (U.S. EPA, 1987). The total chlorobenzene content of the GAC ranged from 150 ppb to 6,630 ppb. Trichlorobenzene was the most prevalent species present, with smaller quantities of di- and tetra-chlorobenzenes detected. Total halogenated organics were measured to be about 150 ppm.

The U.S. EPA has stack tested two GAC reactivation furnaces for the emission of dioxin (U.S. EPA, 1987; Lykins et al., 1987). One facility was an industrial carbon reactivation plant, and the second facility was used to restore GAC at a municipal drinking water plant. The industrial carbon regeneration plant processed 36,000 kg/day of spent GAC used in the treatment of industrial wastewater effluents. Spent carbon was reactivated in a multiple-hearth furnace, cooled in a water quench and, stored and shipped back to primary chemical manufacturing facilities for reuse.

The furnace fired natural gas, and consisted of seven hearths arranged vertically in series. The hearth temperatures ranged from 480 C to 1000 C. The spent GAC contained about 40 percent weight moisture. The used GAC was fed to the top hearth. In the furnace, the spent carbon was dried and the organics adsorbed onto the carbon particles were volatilized and burned in the heated combustion atmosphere. The regenerated carbon dropped from the bottom hearth of the furnace to a quench tank to reduce the temperature. Air pollutant emissions were controlled by an afterburner, a sodium spray cooler, and a fabric filter. Temperatures in the afterburner were about 930 C.

The second GAC reactivation facility tested by U.S. EPA consisted of a fluidized-bed furnace located at a municipal drinking water treatment plant (Lykins et al., 1987). The furnace was divided into three sections: a combustion chamber, a reactivation section and a dryer section. The combustion section was fired by natural gas, and consisted of a stoichiometrically balanced stream of fuel and oxygen. These expanding gases of combustion provided heat and suspended and fluidized the carbon. Temperatures of combustion were about 1,038 C. The reactivation section outside the combustion chamber allowed for the complete volatilization of the heated GAC. Off-gasses from the reactivation/combustion section were directed through an acid gas scrubber and high-temperature afterburner prior to discharge from a stack.

The industrial GAC reaction furnace test data indicate that an average of 5.87E-02 g of CDD/CDF per kg of GAC incinerated may be emitted from the stack during operation (U.S. EPA, 1987). An average of 2.98E-03 g TEQ per kg of GAC may be released to the air during operation. A "medium" confidence rating is given to these emission factors, because only one industrial GAC reactivation furnace operating in the United States has been stack tested. In the second GAC reactivation furnace tested by EPA (Lykins et al., 1987), measurable concentrations of dioxin-like compounds were detected in the stack emissions.

When chlorine was used in pretreatment of the surface water for preliminary disinfection prior to filtration with GAC, the 2,3,7,8-TCDD congener was seen in the particulate stack emission discharges to the incinerator afterburner in low concentration [0.001-0.02 parts per trillion by volume (ppt/v)]. 1,2,6,7-TCDF was detected in two out of four stack tests in a concentration range of 0.004-0.02 ppt/v. When no chlorine was used to disinfect the surface water prior to filtration with GAC, no 2,3,7,8-TCDD was detected (<0.001 ppt/v). With the afterburner operating, no CDD congeners below HpCDD were detected in the stack emissions. Concentrations of HpCDDs and OCDD ranged from 0.001 to 0.05 ppt/v and 0.006 to 0.28 ppt/v, respectively. All congener groups of CDFs were detected in the stack emissions even with the afterburner operating. Total CDFs emitted from the stack averaged 0.023 ppt/v. Measurements of the individual CDD/CDF congeners were not performed, therefore it was not possible to derive emission factors for this facility.

The mass of GAC that is reactivated annually in carbon reactivation furnaces is not known. However, a crude estimate, which is given a "low" confidence rating, is the mass of virgin GAC shipped each year by GAC manufacturers. According to U.S. DOC (1990c), 48 thousand metric tons of GAC were shipped in 1987. Applying the emission factors developed above to this crude estimate of potential GAC reactivation volume, annual releases of 0.14 grams of TEQ and 2.8 grams of total CDD/CDF are estimated. Based on the "medium" confidence rating assigned to the emission factor, the estimated range of potential annual emissions is assumed to vary by a factor of 5 between the low and high ends of the range. Assuming that the best estimate of annual emissions (0.14 g TEQ/yr) is the geometric mean of this range, the range is calculated to be 0.06 to 0.3 g TEQ/yr.

3.6.21. Cement Kilns

Portland cement is a fine, grayish powder consisting of a mixture of four basic materials: lime (calcareous), silica (siliceous), alumina (argillaceous), and iron (ferriferous). Pyroprocessing in a rotary-type kiln plays a central role in fusing the basic raw materials into cement. The raw materials are ground into fine particles and are then either suspended in water to form a pumpable slurry (i.e., wet process) or are fed directly (i.e., dry process) into a rotary kiln for processing at elevated temperatures in an oxygen-enriched atmosphere.

In the rotary kiln, evaporation of the water, calcination of the carbonate constituents, and fusion of the minerals occurs to form clinker. Clinker is a gray-colored, glass-hard material comprised of the cement minerals, dicalcium silicate, tricalcium silicate, calcium aluminate, and tetracalcium aluminoferrite. The clinker is then ground into a fine powder and mixed with gypsum to form portland cement. Approximately 1,575 kg of dry raw materials are needed to produce about 1000 kg of cement clinker (Greer et al., 1992). In 1991, the last year in which data is available, about 66 billion kg of cement clinker was produced in the United States by 212 portland cement kilns requiring about 103 billion kg of raw materials (U.S. EPA, 1993f; Greer et al.,1992).

Because of the relatively high combustion temperature required to produce cement clinker (1400 to 1510 C), coal or petroleum coke are typically used as the primary fuel to sustain combustion in the kiln. However, some cement kilns do burn hazardous liquid and solid waste as supplemental fuel to reduce the amount of coal that is purchased. It is estimated that 34 of the 212 existing cement kilns (i.e., 16 percent) burn hazardous waste as supplemental fuel (U.S. EPA, 1993f). Other types of non-hazardous liquid and solid wastes used as supplemental fuels include tires, waste oil, and wood chips.

The most common air pollution control devices (APCD) employed on rotary kilns are those intended to control dust and particulate matter (i.e., fabric filters and/or electrostatic precipitators). Dioxins were first detected in stack emissions from portland cement kilns in the early 1980s (U.S. EPA, 1987; Peters, 1983; Branscome et al. 1984, 1985). Dioxin was detected only in low amounts and was thought to be caused by the co-firing of liquid hazardous waste with conventional fossil fuels (Peters, 1983).

The EPA gave this source category a low priority for follow-up testing in EPA's National Dioxin Study conducted in 1985-1986 (U.S. EPA, 1987). Since then, the thermolytic reactions and the conditions favoring the formation of CDDs and CDFs in combustion processes have become better understood. (See Section 3.5). Some aspects of this theory warrant investigation into the formation of dioxin in portland cement kilns, including:

Some primary combustion fuels (i.e., coal and petroleum coke) and fuel supplements (wood chips and tires) used to sustain elevated temperatures in the kiln to form clinker may also produce aromatic hydrocarbon compounds (e.g., benzene, phenol) that can later become chlorinated ring structures. The oxidation of HCl gas has been shown to provide chlorine available for ring substitution. In addition, chlorine has been measured directly in the combustion fuels to cement kilns (EER, 1993).

The chlorinated aromatic compounds may act as precursor molecules to the thermalytic formation of CDD/CDFs on the active surface of carbonaceous particulates;

De novo synthesis of CDD/CDFs on the active surface of carbonaceous particulates in the presence of a catalytic agent (e.g., a metal ion such as copper chloride);

Post-kiln temperatures of the combustion gases in the APCD system are within the range of temperatures observed in laboratory studies that promote the continued formation of CDD/CDFs (i.e., 250 to 350 C); and

Co-firing of liquid hazardous organic wastes with coal and petroleum coke may lead to an increase in the amount of CDD/CDFs formed in the post-combustion zone.

Currently, cement kilns that accept and burn hazardous waste as an auxiliary fuel are required under RCRA to characterize pollutant stack emissions, including emissions of CDD/CDFs. EPA's Office of Solid Waste is in the process of collecting and analyzing these emission reports to determine the extent and magnitude of CDD/CDF releases and the need for further regulation.

Preliminary stack test data are available from 14 of the 34 cement kilns burning hazardous waste and from 3 of the 178 kilns not burning hazardous waste (EER, 1993; RTI, 1993). Table 3-35 is a summary of the available emissions data. For kilns accepting and burning hazardous waste as supplemental fuel, it appears that the concentration of dioxin in the stack gas (grams/dscm at 7 percent O2) is highly variable. For example, the average stack emissions of total CDD/CDF for individual kilns range from 2 to 2000 ng/dscm, a thousand-fold difference.

There appears to be no consistent pattern to the relationship of total CDD/CDFs to the estimated dioxin TEQ, indicating a wide variability in the distribution of toxic congeners of CDDs and CDFs in the emissions. For example, the ratio of total CDD/CDFs to the TEQ ranges from about a factor of 5:1 to a factor of 1000:1, indicating that some kilns have a congener distribution skewed toward the lower chlorinated more toxic congeners and others are skewed toward the higher chlorinated less toxic congeners. Cement kilns which do not burn hazardous waste as supplemental fuel appear to be less variable. However, this observation must be tempered by the fact that fewer of these kilns have been stack tested.
...

table Table 3-35Concentrations of Total CDD/CDFs and Dioxin TEQ(grams/dscm) Measured at the Stack of Portland Cement Kilns Burning and Not Burning Hazardous Waste As Supplemental Fuel
... The limited emission data suggests that the average stack concentrations of CDD/CDFs are about eight times higher among the kilns burning hazardous waste than those that do not.

As discussed below, a similar relationship was seen in the cement kiln dust samples from these two categories of kilns.

On this basis, it was decided that separate emission factors should be developed for the kilns burning hazardous waste and those that do not.

Given the limited emission test data, especially among the kilns that do not burn hazardous waste, clearly more testing is needed to confirm this difference in emission factors.
expand Table V2 3-35

National estimates of air emissions of dioxin TEQ/yr from all operating cement kilns were made using two different methods:

1. Dioxin emissions correlate with the total mass of materials processed and burned at the kiln to form clinker (i.e., related to the kiln throughput); and

2. Dioxin emissions correlate with the total energy content of the fuel (including hazardous waste) used to sustain combustion in the kiln.

Although these two methods would likely generate different emission estimates if site-specific data (i.e., throughput and energy consumption data) were available, such data were not available for this report and, therefore, the use of generic industry average data resulted in identical estimates.

For the first method, three values must be known in order to calculate annual dioxin TEQ emissions: (a) the average concentration of dioxin TEQ in the stack gas (g TEQ/dscm); (b) the average volume of combustion gas evolved per kg of material fed to the kiln; and (c) an estimate of the total dry weight of materials processed by all operating cement kilns (kg/yr). Average dioxin TEQ stack emission concentrations are presented in Table 3-35 for kilns burning and not burning hazardous waste as supplemental fuel.

The averaging gave equal weighting to all tested kilns. The average dioxin TEQ stack concentrations are 7.1 ng/dscm and 0.9 ng/dscm for kilns burning hazardous waste and not burning hazardous waste, respectively. A reasonable estimate of combustion gas volume/kg of materials processed in the kiln is 1.75 dscm/kg (RTI, 1993). This approach suggests an emission factor of 12.4 ng TEQ/kg and 1.6 ng TEQ/kg for kilns burning and not burning hazardous waste, respectively.

Greer et. al (1992) estimated that the ratio of the dry weight of materials charged into the kiln to the weight of dry clinker produced is 1.575:1. Therefore, if 66 billion kg of clinker were produced by 212 cement kilns in 1991 (U.S. EPA, 1993f), then 104 billion kg of raw materials were consumed in that year. A final assumption is that the annual throughput of raw materials processed is roughly proportional to the number of kilns in the class of cement kilns (i.e., the number burning hazardous waste versus the number of facilities not burning hazardous waste). Multiplying the emission factors by the raw material throughput yields annual emission estimates of 210 g TEQ for kilns burning hazardous waste and 140 g TEQ for kilns not burning hazardous waste.

For the second method, it is assumed that the dioxin emissions are better calculated on the basis of annual fuel consumption rather than the annual amount of materials processed by the kiln. Given the diversity and mixtures of fuel types typically used (coal, coke, liquid hazardous waste, natural gas, oil, wood chips, tires), a good measurement of total fuel consumption is the amount of total energy consumed to produce the clinker (U.S. EPA, 1993f). Johnson (1992) estimated that 71 trillion kcals were consumed in the year 1991 by all operating portland cement kilns.

Dividing this value by the total kg of raw material processed (is 104 billion kg, as reported above) yields an average energy usage of 680 kcal/kg of raw material. Dividing this value into the emission factors derived above (in units of ng of dioxin TEQ/kg of raw material) yields emission factors in terms of ng of dioxin per kcal. This approach suggests an emission factor of 2.3 pg of TEQ/kcal for kilns not burning hazardous waste and 18 pg of TEQ/kcal for kilns which do burn hazardous waste.

Finally, it is assumed that the total energy usage can be apportioned between kilns burning hazardous waste (34 of 212 kilns) and those that do not (178 of 212 kilns) on the basis of the number of kilns in each group. Multiplying these energy use estimates by the energy based emission factors yields the following estimates of annual emissions: 210 g TEQ for kilns burning hazardous waste and 140 g TEQ for kilns not burning hazardous waste.

Both of these methods suggest that annual dioxin emissions to air from all cement kilns combined is about 350 grams TEQ. The estimated TEQ emission factors used to derive this best estimate of annual TEQ emissions are given a "low" confidence rating because the mechanisms giving rise to dioxin emissions from cement kilns are largely unknown; very few of the existing facilities have been stack tested for emissions; and because of the apparent high variability in the ratio of total CDD/CDFs to dioxin TEQ and the apparent high variability in emissions between tested facilities as shown by the large standard deviation in Table 3-35.

The "production" estimate of annual raw material throughput is given a "high" confidence rating because it is based on recent survey data. Based on these confidence ratings, the estimated range of potential emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the lowest estimate of annual emissions (350 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 110 to 1,100 g TEQ/yr.

In a recent Report to Congress (U.S. EPA, 1993f), EPA's Office of Solid Waste establishes the factual basis for its decision making regarding the appropriate regulatory status, under RCRA, of cement kiln dust (CKD) waste. To aid EPA in their study, the Portland Cement Association (PCA) conducted a survey in 1991 of cement manufacturers. Survey responses were received from 64 percent of the active cement kilns in the United States.

Based on the survey responses, EPA estimated that the U.S. cement industry generated about 12.9 million metric tons of gross CKD and 4.6 million metric tons of "net CKD", of which 4.2 million metric tons was land disposed, in 1990. The material collected by the APCD system is called "gross CKD" (or "as generated" CKD). The gross CKD is either recycled back into the kiln system or is removed from the system for disposal (i.e., "net CKD" or "as managed" CKD) (U.S. EPA, 1993f).

Also in support of the Report to Congress, EPA conducted sampling and analysis during 1992 and 1993 of CKD and clinker. The purposes of the sampling and analysis efforts were:

(1) to characterize the CDD/CDF content of clinker and CKD ;
(2) to determine the relationship, if any, between the CDD/CDF content of CKD and the use of hazardous waste as fuel; and
(3) to determine the relationship, if any, between the CDD/CDF content of CKD and the use of wet versus dry process cement kilns. Clinker samples from 9 kilns and CKD samples from 11 kilns (six of which burn hazardous waste) were analyzed (U.S. EPA, 1993f).

CDD/CDFs were not detected in any of the clinker samples. Tetra- through octa-chlorinated CDDs and CDFs were detected in the "gross CKD" samples obtained from 10 of the 11 kilns and in the "net CKD" samples obtained from 8 of the 11 kilns. The CDD/CDF content of "gross CKD" ranged from 0.008 to 247 ng TEQ/kg and for "net CKD" the content ranged from 0.045 to 195 ng TEQ/kg. Analyses for seven PCB congeners were also conducted but no congeners were detected in any clinker or CKD sample. TCLP leachate testing of the CKD samples from six kilns showed no leaching of CDD/CDFs (detection limits ranged from 3 to 37 pg/L) except for OCDD in two samples (110 and 170 pg/L).

Statistical analysis of the results indicated that mean CDD/CDF concentrations in "net CKD" generated by the sampled kilns burning hazardous waste are higher (35 ng/kg) than in "net CKD" generated by the sampled facilities not burning hazardous waste (3.0E-02 ng/kg). These calculations of mean values treated not detected values as zero. If the not detected values had been excluded from the calculation of the means, then the mean value for "net CKD" from kilns burning hazardous waste would increase by a factor of 1.2 and the mean value for "net CKD" from kilns not burning hazardous waste would increase by a factor of 1.7. One sampled kiln had CDD/CDF concentrations more than two orders of magnitude greater than the TEQ levels found in samples from any other kiln. If this kiln is considered to be atypical of the industry (U.S. EPA, 1993f) and is not included in the calculation, then the mean "net CKD" concentration for hazardous waste burning kilns decreases to 2.9 ng/kg.

From these data, an estimate of dioxin TEQ emissions to land in the form of land-disposed "net CKD" can be made. The estimate of land-disposed CKD from the 1991 PCA Survey, 4.2 million metric tons per year (basis year is 1990), was divided among kilns burning hazardous waste (34 kilns) and those which do not (178 kilns) on the basis of the number of kilns in each category. The average TEQ concentration in the net CKD from kilns burning hazardous waste, including the potentially non-typical kiln, was 35 ng TEQ/kg. For kilns which do not have hazardous waste the average concentration in the "net CKD" was 3.0E-02.

Multiplying these average concentrations by the annual "net CKD" production, yields estimates of 24 g TEQ/yr for kilns burning hazardous waste and 0.1 g TEQ/yr for kilns not burning hazardous waste, yielding a total of 24.1 g TEQ/yr for all kilns. The "production" estimate was assigned a "high" confidence rating because it is based on recent EPA survey data. The "emission factor" estimates are assigned a "low" confidence rating because the sampling data upon which they are based showed high variability among the 11 kilns sampled (out of 212 kilns in the United States). Based on these confidence ratings, the estimated range of potential annual emissions is assumed to vary by a factor of 10 between the low and high ends of the range. Assuming that the best estimate of annual emissions (24.1 g TEQ/yr) is the geometric mean of this range, then the range is calculated to be 7.6 to 76 g TEQ/yr.

 

3.6.22. Additional Combustion and High Temperature Sources

Although discussed as potential sources in the preceding sections, insufficient data are available upon which to develop emission factors for the primary ferrous and primary nonferrous metal refining/smelting industries. In addition, although emission estimates were developed in Section 3.6.13 for diesel-powered on-highway vehicles (i.e., the largest use of diesel fuel in the United States), because of insufficient data no estimates were developed for off-highway transportation diesel engine fuel use (including railroad engine fuel and fuel for agricultural machinery) or for diesel fuel use by the commercial and industrial sectors of the economy. Also, although no estimated emissions could be developed for residential oil, gas, and charcoal combustion because of lack of emission rate factors, these have been identified as potential sources by Harrad et al. (1992a, 1992b) and Fiedler and Hutzinger (1991b).

3.7. RESERVOIR SOURCES
It is very difficult to estimate CDD/CDF releases that may be occurring from reservoir sources. However, some idea of the potential magnitude of these emissions can be gained by estimating the size of the overall reservoir. Equation 1 calculates the concentration of a contaminant in a reservoir given the deposition rate of the contaminant into the reservoir and the rate of dissipation from that reservoir:
Formula V2 3-1

where C is the concentration after time t, DEP is the deposition rate (in units of mass/area-time), k is the first order dissipation rate (time-1), and MIX is the mass of the reservoir into which DEP mixes (mass units, corresponding to area of DEP).

Consider the case where DEP has been occurring for a number of years at a steady rate. A question that might be asked is, what is the contribution of a year's worth of deposition to the amount that is already there. This can be estimated with a ratio of estimated concentrations C1/C2 using Equation (1), where C1 is the concentration after the number of years, and C2 is the year's worth of deposition. Assuming DEP, MIX, and k are constant, the ratio of C1/C2 reduces to:

Formula V2 3-2

where t1 is the number of years that DEP has been occurring, and t2 is equal to 1 for the one year's worth of deposition. For the sake of this discussion, if one assumes that DEP for dioxin-like compounds has been steady since the 1940s, and one wants to evaluate a year's worth of deposition in the 1990s, then t1 equals about 50 years. A dissipation rate of 0.0693, corresponding to a 10-year half-life was used for atmospheric deposition onto soils for the methodologies described in Volume III. A t1 of 50 and k of 0.0693 applied to Equation (2) yields a ratio of about 14.5. This means that there would be about 14.5 times more contaminant in the reservoir than one year's contribution.

However, the half-life of 10 years is probably too low for this exercise. This half-life was generated from data on 2,3,7,8-TCDD applied to experimental plots as 2,4,5-T in Agent Orange testing (Young, 1983). This might be appropriate for dissipation from a bounded area of high soil contamination, where dissipation mechanisms such as soil erosion, dust resuspension, or volatilization might be occurring. However, if MIX is considered as the total reservoir of soil and surface vegetation, then losses from a bounded area are unlikely to translate to losses from the larger system. In other words, a more appropriate half-life for this discussion might be more like 50 years than 10 years. If 50 years is assumed in the above exercise, than the ratio increases to 36.

This analysis suggests that dioxin-like compounds already in the reservoir source may exceed annual contributions to the reservoir source by 15 or more times. The potential for emissions from this large source is uncertain. Dioxin-like compounds that accumulate in deep sediments or become buried in the soil are not likely to contribute to current emissions. However, those located near the surface could become re-entrained into the air or water bodies.